Natural_and_Enhanced.pdf

NATURAL andENHANCEDREMEDIATION

SYSTEMS

L1282_fm_frame Page 2 Monday, June 18, 2001 10:20 AM

LEWIS PUBLISHERSBoca Raton London New York Washington, D.C.

NATURAL andENHANCEDREMEDIATION

SYSTEMSSuthan S. Suthersan

This book contains information obtained from authentic and highly regarded sources. Reprinted materialis quoted with permission, and sources are indicated. A wide variety of references are listed. Reasonableefforts have been made to publish reliable data and information, but the author and the publisher cannotassume responsibility for the validity of all materials or for the consequences of their use.

Neither this book nor any part may be reproduced or transmitted in any form or by any means, electronicor mechanical, including photocopying, microfilming, and recording, or by any information storage orretrieval system, without prior permission in writing from the publisher.

The consent of CRC Press LLC does not extend to copying for general distribution, for promotion, forcreating new works, or for resale. Specific permission must be obtained in writing from CRC Press LLCfor such copying.

Direct all inquiries to CRC Press LLC, 2000 N.W. Corporate Blvd., Boca Raton, Florida 33431.

Trademark Notice:

Product or corporate names may be trademarks or registered trademarks, and areused only for identification and explanation, without intent to infringe.

Visit the CRC Press Web site at www.crcpublications.com

© 2002 CRC Press LLC Lewis Publishers is an imprint of CRC Press LLC

No claim to original U.S. Government worksInternational Standard Book Number 1-56670-282-8

Library of Congress Card Number 2001029566Printed in the United States of America 1 2 3 4 5 6 7 8 9 0

Printed on acid-free paper

Library of Congress Cataloging-in-Publication Data

Suthersan, Suthan S.Natural and enhanced remediation systems / by Suthan S. Suthersan.

p. cm. — (Arcadis Geraghty & Miller science and engineering)Includes bibliographical references and index.ISBN 1-56670-282-81. Soil remediation. 2. Groundwater–Purification. 3. Hazardous wastes–Natural

attenuation. 4. Bioremediation. I. Title. II. Geraghty & Miller environmental science andengineering series.TD878.S873 2001628.5—dc21 2001029566

CIP


Sincere Thanks To:

Sumathy, Shauna, and Nealon for their enthusiastic support and unending patience.

STP, MTP, MLM, and SBB for their insight, support, inspiration, and trust.

Dedicated with utmost humility to the heroes and heroines of Eelam who have put their lives in the line of fire to express

their intellectual freedom.



Foreword

I have worked with Dr. Suthersan for the past 13 years and have seen firsthandthe impact he has had on the evolution of our business. Over this period, environ-mental remediation has moved from a world of standard operation and applicationof proven technology to one where more innovative concepts can be applied, tested,and developed for the benefit of the environment, the regulatory community, andindustry. Dr. Suthersan has worked assiduously to develop new remediation tech-nologies, move them to pilot testing in cooperation with industry, and make themdemonstrated techniques.

As our industry has matured, the pressures on all parties have increased: pressureto assure protection of human health and the environment, to remediate faster, torapidly return sites to beneficial use, to reduce costs, etc. Finding a solution to thesecompeting objectives has become more and more intricate and must include theimpacts of social, economic, business, and environmental factors. Dr. Suthersan isone of the most talented purveyors of remediation technology as a tool to solve thesecomplex problems in a world where competing priorities are the rule not the excep-tion. The author has focused on finding these total business solutions for our industry,using the innovative technical solutions he or others have created. Finding totalbusiness solutions to multifaceted environmental problems is one of the hallmarksof Dr. Suthersan’s career.

In this book, Dr. Suthersan explains some of the pioneering remediation tech-nologies developed over the past few years. The focus is on those techniques thatmodify or enhance the natural environment to aid in the remediation of contaminants.When applied correctly, these engineered, natural systems have proven to be moreefficient and cost effective than their more intrusive predecessors. Assuring that thesetechniques are applied correctly and tailored to each particular setting is a keycomponent of any system’s success. The impact of biological, chemical, and hydro-geologic settings on these technologies is thoroughly discussed. Dr. Suthersandescribes each technique in detail: its processes, the science behind it, its application,and the constraints. This book will be an invaluable resource to the practicingremediation engineer, the regulatory community charged with evaluating these tech-niques, and the industry applying them.

It has been a privilege to have worked with Dr. Suthersan for these past yearsand to have seen the influence of his knowledge and skill in our industry. I believethat those who read this book will gain from his wisdom.

Steve Blake

Executive Board, ARCADIS, N.V.Denver, Colorado

L1282_fm_frame Page 7 Tuesday, June 19, 2001 12:56 PM


Preface

Remediation of hazardous wastes present in the subsurface has evolved withtime and has been influenced by various factors over the years. During the earlyyears, direction and efforts were mostly influenced by the regulations in place andthe need for compliance and protection of human health and environment. Thecontaminants primarily focused upon during this time were the petroleum-relatedcontaminants stemming from leaking underground storage tanks (USTs).

In later years, remediation efforts were driven by a combination of economicand regulatory factors. During this time contaminants that caught most of the atten-tion were the chlorinated solvents, heavy metals, and chlorinated and nonchlorinatedpolynuclear aromatic hydrocarbons (PAHs). The current focus seems to be taking adifferent direction: instead of focusing on the type of contaminants, emphasis is onevaluating the damage to the environment (and thus the risk) and repairing thatdamage in a cost-effective manner.

Evolution of remediation technologies was influenced not only by changingregulatory and economic factors, but also by the type and chemical characteristicsof contaminants under focus. An example is the shift in emphasis from engineeredaerobic bioremediation systems of the 1980s to engineered anaerobic bioremediationsystems of the 1990s. Significant reliance and dependence on natural remediationsystems have increased as a result of recent acceptance that landfills behave asbioreactors and the very recent focus on dealing with ecological risks and naturalresources damage (NRD) assessments. Ever increasing understanding of the behav-ior of most contaminants in the natural environment has also led to the effort ofmaximizing the remediation potential of natural systems.

The thematic focus of this book is to highlight the current phase in the evolutionof remediation technologies. All the technologies discussed in the book utilize orenhance the natural biogeochemical environment for remediation of hazardous con-taminants. The discussion throughout the book is focused towards helping practitionersof remediation to engineer remediation systems utilizing the natural environment.These natural systems or reactors still have to be properly designed and engineered tooptimize the performance and maximize contaminant removal efficiencies.

The basic understanding of environmental and contaminant characteristicsrequired to design these systems is provided in Chapter 2. I had just coined thephrase “

in situ

reactive zones (IRZ)” when I wrote my previous book in 1996 andwas able to provide only an introduction of the technology. I have made a signif-icant effort in Chapter 4 to describe the IRZ technology and its various modifiedapplications. The manner in which the application of this technology is explodingmay justify a book of its own. I am proud to see the advances and expansion ofthis technology pioneered by my colleagues and me at ARCADIS G & M, Inc.Due to the shortage of space I could not present data from all the successful sitesusing this technology. Technical advances and theoretical insights on the applica-tion of

in situ

chemical oxidation are also presented in Chapter 4 (special thanksto Dr. Fred Payne).


I also had the privilege of being involved in some of the earliest phytoremediationand phyto-cover applications. Some contributions to the science of designing phyto-covers are presented in Chapter 7 (special thanks to Dr. Scott Potter). I have providedonly a summary on the current state of the science of phytoremediation in Chapter5. Basic concepts of treatment wetlands are provided in Chapter 6. I truly believethat this technology will have more applications in the field of hazardous wasteremediation.

I wrote this book to reach a wide audience: remediation design engineers,scientists, regulatory specialists, graduate students in environmental engineering,and people from the industry who have general responsibility for site cleanups. Ihave tried to provide a general, basic description of the technologies in all chaptersin addition to detailed information on basic principles and fundamentals in mostchapters. Readers who are not interested in basic principles can skip these passagesand still receive the general knowledge they need.

Suthan S. Suthersan

Yardley, Pennsylvania


Acknowledgments

First and foremost, I would like to thank members and colleagues from theInnovative Strategies Group (ISG) of ARCADIS — Frank Lenzo, Mike Hansen, andJeff Burdick — for their enthusiasm and hard work in trying to experiment withinnovative and cutting edge technologies in the field. Insights and advice providedby Drs. Scott Potter and Fred Payne in formulating the theoretical and mathematicalfoundations behind the technical concepts are immense. In addition, the patienceand excitement exhibited by Chris Lutes and David Liles during the laboratory“proof of concept” experiments always boosted my confidence to proceed to thenext level in implementing many of the technologies. Taking these technologiesfrom the conceptual level to field scale applications would not have been possiblewithout these individuals.

I have to thank Eileen Schumacher and Ben Tufford for patiently drafting allthe figures and Amy Weinert and Gail Champlin for typing the manuscript. Themanagement of my employer ARCADIS G & M, Inc. deserves special mention forall the support given to me over the years. The opportunities and encouragementprovided to me in order to “think out of the box” are a reflection of the company’sculture. I owe a special debt to all the engineers and project managers who helpedme to implement many innovative and challenging remediation projects. This list isa long one, but special mention is due to the following: Mike Maierle, Don Kidd,Gary Keyes, Steve Brussee, Jack Kratzmeyer, Mark Wagner, Jim Drought, TinaStack, Eric Carman, Al Hannum, John Horst, Kurt Beil, Dave Vance, Nanjun Shetty,and Pat Hicks.

The encouragement, support, and feedback on the state of the science approachesin phytoremediation by Drs. Steve Rock and Steve McCutcheon, of the USEPA, arevery much appreciated.



The Author

Suthan S. Suthersan

, Ph.D., P.E., is senior vice presidentand director of Innovative Remediation Strategies atARCADIS G & M, Inc., an international environmentaland infrastructure services company. In his 12 years withthe company, Dr. Suthersan has helped make AG&M oneof the most respected environmental engineering compa-nies in the U.S., specifically in the field of

in situ

remedi-ation of hazardous wastes. Many of the technologies hepioneered have since become industry standards. His big-gest contribution to the industry, beyond the technologydevelopment itself, has been to convince the regulatorycommunity that these innovative technologies are better

than traditional ones, not only from a cost viewpoint, but also for technical effec-tiveness. His experience is derived from working on at least 500 remediation projectsin design, implementation, and technical oversight capacities during the past 15years.

Dr. Suthersan’s technology development efforts have been rewarded with sevenpatents awarded and more pending. His most important recent contributions arereflected by the following patents: Engineered

In Situ

Anaerobic Reactive Zones,US Patent 6,143,177; In Well Air Stripping, Oxidation, and Adsorption, US Patent6,102,623;

In Situ

Anaerobic Reactive Zone for

In Situ

Metals Precipitation and toAchieve Microbial De-Nitrification, US Patent 5,554,290;

In Situ

Reactive Gate forGroundwater Remediation, US Patent 6,116,816.

Dr. Suthersan has a Ph.D. in environmental engineering from the University ofToronto, a M.S. degree in environmental engineering from the Asian Institute ofTechnology, and a B.S. degree in civil engineering from the University of Sri Lanka.In addition to his consulting experience Dr. Suthersan has taught courses at severaluniversities. He is the founding editor in chief of the

Journal of Strategic Environ-mental Management

and is a member of the editorial board of the

InternationalJournal of Phytoremediation

.



Contents

Chapter 1

Hazardous Wastes Pollution and Evolution of Remediation....................................11.1 Introduction ......................................................................................................11.2 The Concept of Risk ........................................................................................2

1.2.1 The Decision Making Framework .......................................................31.3 Evolution of Understanding of Fate and Transport in

Natural Systems ...............................................................................................41.4 Evolution of Remediation Technologies .........................................................7References................................................................................................................11

Chapter 2

Contaminant and Environmental Characteristics ....................................................132.1 Introduction ....................................................................................................142.2 Contaminant Characteristics ..........................................................................18

2.2.1 Physical/Chemical Properties ............................................................182.2.1.1 Boiling Point.......................................................................182.2.1.2 Vapor Pressure ....................................................................182.2.1.3 Henry’s Law Constant ........................................................192.2.1.4 Octanol/Water Partition Coefficients..................................202.2.1.5 Solubility in Water..............................................................202.2.1.6 Hydrolysis ...........................................................................222.2.1.7 Photolytic Reactions in Surface Water...............................24

2.2.2 Biological Characteristics ..................................................................262.2.2.1 Cometabolism .....................................................................272.2.2.2 Kinetics of Biodegradation.................................................32

2.3 Environmental Characteristics .......................................................................382.3.1 Sorption Coefficient ...........................................................................38

2.3.1.1 Soil Sorption Coefficients...................................................432.3.1.2 Factors Affecting Sorption Coefficients .............................48

2.3.2 Oxidation-Reduction Capacities of Aquifer Solids ...........................512.3.2.1 pe and pH............................................................................512.3.2.2 REDOX Poise .....................................................................522.3.2.3 REDOX Reactions ..............................................................53

References................................................................................................................58

Chapter 3

Monitored Natural Attenuation ...............................................................................633.1 Introduction ....................................................................................................64

3.1.1 Definitions of Natural Attenuation ....................................................643.2 Approaches for Evaluating Natural Attenuation ...........................................653.3 Patterns vs. Protocols .....................................................................................70


3.3.1 Protocols for Natural Attenuation......................................................703.3.2 Patterns of Natural Attenuation .........................................................71

3.3.2.1 Various Patterns of Natural Attenuation.............................723.4 Processes Affecting Natural Attenuation of Compounds..............................79

3.4.1 Movement of Contaminants in the Subsurface .................................793.4.1.1 Dilution (Recharge) ............................................................793.4.1.2 Advection ............................................................................813.4.1.3 Dispersion ...........................................................................83

3.4.2 Phase Transfers ..................................................................................853.4.2.1 Sorption...............................................................................853.4.2.2 Stabilization ........................................................................883.4.2.3 Volatilization .......................................................................89

3.4.3 Transformation Mechanisms..............................................................893.4.3.1 Biodegradation ....................................................................90

3.5 Monitoring and Sampling for Natural Attenuation .....................................1093.5.1 Dissolved Oxygen (DO) ..................................................................1133.5.2 Oxidation–Reduction (REDOX) Potential (ORP)...........................1173.5.3 pH .....................................................................................................1193.5.4 Filtered vs. Unfiltered Samples for Metals .....................................120

3.5.4.1 Field Filtration and the Nature of Groundwater Particulates..................................................121

3.5.4.2 Reasons for Field Filtration..............................................1223.5.5 Low-Flow Sampling as a Paradigm for Filtration ..........................1243.5.6 A Comparison Study........................................................................125

References..............................................................................................................126

Chapter 4

In Situ

Reactive Zones...........................................................................................1314.1 Introduction ..................................................................................................1324.2 Engineered Anaerobic Systems ...................................................................135

4.2.1 Enhanced Reductive Dechlorination (ERD) Systems .....................1354.2.1.1 Early Evidence..................................................................135

4.2.1.1.1 Biostimulation vs. Bioaugmentation ................1364.2.1.2 Mechanisms of Reductive Dechlorination .......................1384.2.1.3 Microbiology of Reductive Dechlorination .....................142

4.2.1.3.1 Cometabolic Dechlorination .............................1424.2.1.3.2 Dechlorination by Halorespiring

Microorganisms.................................................1444.2.1.4 Electron Donors ................................................................147

4.2.1.4.1 Production of H

2

by Fermentation ...................1494.2.1.4.2 Competition for H

2

...........................................1524.2.1.5 Mixture of Compounds on Kinetics.................................1554.2.1.6 Temperature Effects ..........................................................1584.2.1.7 Anaerobic Oxidation.........................................................158


4.2.1.8 Electron Acceptors and Nutrients.....................................1584.2.1.9 Field Implementation of IRZ for Enhanced Reductive

Dechlorination...................................................................1604.2.1.10 Lessons Learned ...............................................................1634.2.1.11 Derivation of a Completely Mixed System for

Groundwater Solute Transport of Chlorinated Ethenes...1704.2.1.12 IRZ Performance Data......................................................177

4.2.2

In Situ

Metals Precipitation .............................................................1834.2.2.1 Principles of Heavy Metals Precipitation.........................1874.2.2.2 Aquifer Parameters and Transport Mechanisms ..............1954.2.2.3 Contaminant Removal Mechanisms.................................196

4.2.3

In Situ

Denitrification.......................................................................1974.2.4 Perchlorate Reduction ......................................................................199

4.3 Engineered Aerobic Systems .......................................................................2004.3.1 Direct Aerobic Oxidation.................................................................200

4.3.1.1 Aerobic Cometabolic Oxidation.......................................2024.3.1.2 MTBE Degradation ..........................................................204

4.4

In Situ

Chemical Oxidation Systems...........................................................2054.4.1 Advantages .......................................................................................2064.4.2 Concerns...........................................................................................2074.4.3 Oxidation Chemistry ........................................................................208

4.4.3.1 Hydrogen Peroxide ...........................................................2114.4.3.2 Potassium Permanganate ..................................................2134.4.3.3 Ozone ................................................................................216

4.4.4 Application .......................................................................................2184.4.4.1 Oxidation of 1,4-Dioxane by Ozone................................2224.4.4.2 Biodegradation Enhanced by Chemical

Oxidation Pretreatment.....................................................2234.5 Nano-Scale Fe (0) Colloid Injection within an IRZ ...................................223

4.5.1 Production of Nano-Scale Iron Particles .........................................2284.5.2 Injection of Nano-Scale Particles in Permeable Sediments............2314.5.3 Organic Contaminants Treatable by Fe (0) .....................................231

References..............................................................................................................233

Chapter 5

Phytoremediation ...................................................................................................2395.1 Introduction ..................................................................................................2405.2 Chemicals in the Soil–Plant System............................................................241

5.2.1 Metals ...............................................................................................2415.2.2 Organics............................................................................................242

5.3 Types of Phytoremediation ..........................................................................2445.3.1 Phytoaccumulation ...........................................................................2455.3.2 Phytodegradation..............................................................................2485.3.3 Phytostabilization .............................................................................2505.3.4 Phytovolatilization............................................................................251


5.3.5 Rhizodegradation..............................................................................2525.3.6 Rhizofiltration...................................................................................2565.3.7 Phytoremediation for Groundwater Containment ...........................2595.3.8 Phytoremediation of Dredged Sediments ........................................260

5.4 Phytoremediation Design.............................................................................2615.4.1 Contaminant Levels .........................................................................2655.4.2 Plant Selection..................................................................................2655.4.3 Treatability .......................................................................................2665.4.4 Irrigation, Agronomic Inputs, and Maintenance .............................2665.4.5 Groundwater Capture Zone and Transpiration Rate .......................267

References..............................................................................................................267

Chapter 6

Constructed Treatment Wetlands...........................................................................2696.1 Introduction ..................................................................................................270

6.1.1 Beyond Municipal Wastewater ........................................................2726.1.2 Looking Inside the “Black Box” .....................................................2736.1.3 Potential “Attractive Nuisances”......................................................2746.1.4 Regulatory Uncertainty and Barriers ...............................................275

6.2 Types of Constructed Wetlands ...................................................................2766.2.1 Horizontal Flow Systems.................................................................2766.2.2 Vertical Flow Systems......................................................................277

6.3 Microbial and Plant Communities of a Wetland.........................................2786.3.1 Bacteria and Fungi ...........................................................................2786.3.2 Algae ................................................................................................2796.3.3 Species of Vegetation for Treatment Wetland Systems...................279

6.3.3.1 Free-Floating Macrophyte-Based Systems.......................2826.3.3.2 Emergent Aquatic Macrophyte-Based Systems ...............2846.3.3.3 Emergent Macrophyte-Based Systems with Horizontal

Subsurface Flow ...............................................................2856.3.3.4 Emergent Macrophyte-Based Systems with Vertical

Subsurface Flow ...............................................................2856.3.3.5 Submerged Macrophyte-Based Systems ..........................2856.3.3.6 Multistage Macrophyte-Based Treatment Systems..........287

6.4 Treatment-Wetland Soils..............................................................................2876.4.1 Cation Exchange Capacity...............................................................2896.4.2 Oxidation and Reduction Reactions ................................................2906.4.3 pH .....................................................................................................2926.4.4 Biological Influences on Hydric Soils.............................................2926.4.5 Microbial Soil Processes..................................................................2926.4.6 Treatment Wetland Soils ..................................................................293

6.5 Contaminant Removal Mechanisms ............................................................2946.5.1 Volatilization ....................................................................................2946.5.2 Partitioning and Storage...................................................................2956.5.3 Hydraulic Retention Time................................................................297


6.6 Treatment Wetlands for Groundwater Remediation....................................2996.6.1 Metals-Laden Water Treatment........................................................300

6.6.1.1 A Case Study for Metals Removal ..................................3026.6.2 Removal of Toxic Organics .............................................................306

6.6.2.1 Biodegradation ..................................................................3066.6.3 Removal of Inorganics .....................................................................3096.6.4 Wetland Morphology, Hydrology, and Landscape Position............309

References..............................................................................................................310

Chapter 7

Engineered Vegetative Landfill Covers .................................................................3137.1 Historical Perspective on Landfill Practices................................................3147.2 The Role of Caps in the Containment of Wastes........................................3157.3 Conventional Landfill Covers ......................................................................3167.4 Landfill Dynamics........................................................................................3177.5 Alternative Landfill Cover Technology .......................................................3217.6 Phyto-Cover Technology..............................................................................321

7.6.1 Benefits of Phyto-Covers over Traditional RCRA Caps.................3267.6.2 Enhancing

In Situ

Biodegradation...................................................3267.6.3 Gas Permeability ..............................................................................3277.6.4 Ecological and Aesthetic Advantages..............................................3277.6.5 Maintenance, Economic, and Public Safety Advantages ................329

7.7 Phyto-Cover Design .....................................................................................3297.7.1 Vegetative Cover Soils .....................................................................3307.7.2 Nonsoil Amendment ........................................................................3317.7.3 Plants and Trees ...............................................................................331

7.8 Cover System Performance..........................................................................3327.8.1 Hydrologic Water Balance ...............................................................3327.8.2 Precipitation .....................................................................................3357.8.3 Runoff...............................................................................................3357.8.4 Potential Evapotranspiration — Measured Data .............................3377.8.5 Potential Evapotranspiration — Empirical Data .............................3397.8.6 Effective Evapotranspiration ............................................................3407.8.7 Water Balance Model.......................................................................343

7.9 Example Application....................................................................................3447.10 Summary of Phyto-Cover Water Balance....................................................3477.11 General Phyto-Cover Maintenance Activities .............................................348

7.11.1 Site Inspections ................................................................................3487.11.2 Soil Moisture Monitoring ................................................................349

7.11.2.1 Drainage Measurement.....................................................3507.11.3 General Irrigation Guidelines ..........................................................3527.11.4 Tree Evaluation ................................................................................356

7.11.4.1 Stem ..................................................................................3567.11.4.2 Leaves ...............................................................................356

7.11.5 Agronomic Chemistry Sampling .....................................................357


7.11.6 Safety and Preventative Maintenance..............................................3597.11.7 Repairs and Maintenance.................................................................359

7.12 Operation and Maintenance (O&M) Schedule............................................3597.12.1 Year 1 — Establishment ..................................................................3607.12.2 Years 2 and 3 — Active Maintenance.............................................3607.12.3 Year 4 — Passive Maintenance .......................................................361

7.13 Specific Operational Issues..........................................................................3627.13.1 Irrigation System Requirements ......................................................3627.13.2 Tree Replacement.............................................................................362

References..............................................................................................................362

Appendix A

Physical Properties of Some Common Environmental Contaminants .................365

Appendix B

Useful Information for Biogeochemical Sampling...............................................383

Appendix C

Common and Scientific Names of Various Plants ................................................405

Index

......................................................................................................................409


1

CHAPTER

1

Hazardous Wastes Pollution andEvolution of Remediation

CONTENTS

1.1 Introduction ......................................................................................................11.2 The Concept of Risk ........................................................................................2

1.2.1 The Decision Making Framework .......................................................31.3 Evolution of Understanding of Fate and Transport in Natural Systems ........41.4 Evolution of Remediation Technologies .........................................................7References................................................................................................................11

The earth was made so various that the mind of desultory man, studious ofchange and pleased with novelty, might be indulged.

1.1 INTRODUCTION

Among the many environmental problems that have received attention in recentdecades is subsurface contamination caused by hazardous wastes. This has been dueto the growing concern over short and long term health and environmental effectsof toxic substances released into the environment.

The public policy maker is faced with particular difficulties in regulating haz-ardous pollutants, most notably because of the high levels of uncertainty surroundingthe issue. Such uncertainty exists in determining the precise impacts in relation toboth human health effects and long term effects on the environment, especially withrecalcitrant pollutants, or pollutants with extremely slow degradation rates. Never-theless, policy makers have been required to formulate environmental regulationsusing some dependable basis. While theoretical methods of decision making suchas dose-response and risk-benefit analysis may be employed to assist regulators,

L1282_C01_frame Page 1 Monday, June 18, 2001 10:18 AM

2 NATURAL AND ENHANCED REMEDIATION SYSTEMS

they are still faced with conflicting pressures — for example, between political andeconomic priorities or between public demands and technical expertise.

Estimating the damage function of a pollutant is an exercise that underpinsregulatory formulation for hazardous waste management. Figure 1.1 outlines thebasic steps involved in this estimation. Determining the transfer function of a haz-ardous pollutant raises several problems. Persistent, nondegradable substances willtend to accumulate in the environment often becoming concentrated in the foodchain. In the past, the potential life span of persistent substances in the subsurfacewas considered to be decades or even centuries. Research performed and scientificadvancements, specifically in the last decade, indicate that compounds deemed tobe persistent or nondegradable in the past are considered to be less persistent or atleast partially degradable under natural conditions.

1.2 THE CONCEPT OF RISK

Many of the problems associated with hazardous waste management, such as uncer-tainty, irreversibility, and persistence make the concept of risk relevant to this discussion.From an engineering or scientific standpoint, “risk” may be defined in quantitative termsby applying probabilistic measures. If “hazard” is defined as the potential for adverseconsequences of some event, then “risk” may be defined as the chance of a particularhazard occurring. It combines two aspects — the probabilistic measure of the occurrenceof the event with a measure of the consequences of the event (in this case the level oftoxicity of the pollutant). Further aspects of risk are highlighted by social scientists whoexamine risk perception in recognition that a particular risk or hazard may meandifferent things to different people in different contexts.

The concept of risk is not without problems, particularly in relation to the issueof hazardous pollutants. For example, an initial problem is determining the proba-bility of such risks; there have been only a few decades of experience in dealingwith many pollutants. Their effects on human beings had been largely unknown,and thus the probabilistic calculations of risk on exposure and associated health andecological impacts were mostly conservative.

In relation to the assumed, perceived, or calculated risks associated with haz-ardous pollutants until recently, it is important to highlight two significant features:1) the subjective probability of the hazard (caused by toxicity of the releasedpollutant) occurring may be very low, but, 2) the consequence of the hazard wasassumed or perceived to be very high, often as irreversible because of assumptions

Figure 1.1

Evaluation of pollution damage.

Release ofPollutants

Rate and Massat a Particular

Place and Time

AmbientConditions

Concentrationsof Pollutants inDifferent Media

DamageEffects

Physical, Ecological,Health, Property,

Natural Resources

ExposurePathways

Dose-ResponseFunction

TransferFunction

Monetary


HAZARDOUS WASTES POLLUTION AND EVOLUTION OF REMEDIATION 3

of persistence and negligible degradation of many pollutants in the natural environ-ment. Thus the low probability tends to cancel out the assumed or perceived impactsassociated with the risk.

Recent research shows that pollutants and other organic chemicals present in thesubsurface become less available or create lesser levels of hazard (in other wordsbecome less toxic) due to interactions between the compound and the subsurfaceenvironment. This drop in availability and toxicity lowers the risk of these chemicalsto human and ecological receptors. Furthermore, the availability of an organicchemical in the subsurface is not a function of its measured concentration; rather,it depends upon the geologic and biogeochemical characteristics of the subsurface,the physicochemical properties of the chemical itself, and the time of contact betweenthe chemical and the subsurface media, i.e., aging, as well as the type and extent oftreatment, natural or anthropogenic, to which it has been subjected.

1.2.1 The Decision Making Framework

In the face of the many uncertainties surrounding hazardous waste managementwith respect to the assumed, perceived, or calculated risks, the regulatory authoritiesare faced with an initial decision about the appropriate framework for decision-making: should it be a balancing approach such as cost-benefit or risk-benefitanalysis, or should it be an approach which emphasizes the protection of humanhealth and natural resources regardless of costs?

Three types of approaches have been utilized to implement hazardous wastemanagement policies in the U.S. during the past three decades:

1. Health-based approaches — zero risk, significant risk, or acceptable risk2. Balancing approaches — cost-benefit, risk-benefit, or decision analysis3. Technology-based approaches — best available technology, risks as low as rea-

sonably practicable

Environmental threats, rather than the scientific evidence and theory from whichthey may be deduced, have been ill-defined during the past three decades. Theevidence from which a threat is deduced has been challenged by conflicting evidenceor placed into a context of associations which heightens its significance. A scenariofor an exposure pathway typically used in the past, where a kid climbing an eight-foot fence and eating a few grams of soil every day for a decade is an example ofsuch an association. For many years, there was a widespread but unfounded assump-tion that some toxic pollutants stemming from industrial releases and/or accidentsand landfills would not be degraded in the natural environment. The measurementof damage, and thus the risk, requires an understanding of the physical processesof transportation and of the distribution and deposition of pollutants, including theirchemical and biological transformations on the way.

The creation of new knowledge usually involves institutions very different fromthose concerned with its acceptance, application, and dissemination. A genuinescience-based environmental policy should be a dynamic one and evolve via con-tinuous monitoring of pollutants in many media, as well as of their impacts on theecosystem and human health (or any other selected target organisms). The technical



means for such monitoring must, of course, be available, as must the baselines forthe establishment of a time series so that change can be observed over time in thenatural environment. There must be agreement on which pollutants to monitor andhow to synthesize and use the masses of data that will accumulate.

Even complete understanding of how the subsurface works as a bio-geo-physi-co-chemical system cannot give ready answers as to the proper regulatory response,i.e., how to use the earth in the common interest of humanity and without degradingit for future generations. This is probably why the government, more out of frustra-tion than intent, has come to rely less on science, engineering, and economics andmore on caution and law.

Figure 1.2 describes the shortcomings of the health-based, conservative approachof the past and the more credible, balanced approach still evolving. Understandingthe contribution by Mother Nature towards a natural remediation process has had asignificant influence on this evolution.

1.3 EVOLUTION OF UNDERSTANDING OF FATE AND TRANSPORT IN NATURAL SYSTEMS

Predicting the hazard of an organic contaminant to humans, animals, and plantsrequires information not only on its toxicity to living organisms but also on thedegree of exposure of the organisms to the compound. The mere release or dischargeof a pollutant does not, in itself, constitute a hazard; the individual human, animal,or plant must also be exposed to it. In evaluating exposure, the transport of the

Figure 1.2

A hypothetical analysis of cost to risk reduction benefit ratios during remediationactivities.

Remediation expenditurewhich justifies reasonablerisk reduction

Cos

t of R

emed

iatio

n ($

)

Ass

ocia

ted

Ris

k

10 -7

-610

-510

-410

-310



chemical and its fate must be considered. A molecule that is not subject to environ-mental transport is not a health or environmental problem except to species at thespecific point of release.

Information on dissemination of the chemical from the point of its release to thepoint where it could have an effect is of great relevancy for risk calculations.However, the chemical may be modified structurally or totally destroyed during itstransport, and the fate of the compound during transport, that is, its modification ordestruction, is crucial to defining the exposure. A compound modified to yieldproducts that are less or more toxic, or totally degraded to harmless end products,or bio-magnified — factors associated with the fate of the molecule — representsgreater or lesser hazard to the species potentially exposed to injury.

At the specific site of discharge or during its transport, the pollutant moleculeor ion may be acted on by abiotic mechanisms. Photochemical transformations occurin the atmosphere and at or very near the surfaces of water, soil, and vegetation,and these processes may totally destroy or appreciably modify a number of differenttypes of organic chemicals. Nonenzymatic, nonphotochemical reactions are alsoprominent in soil, sediment, and surface and groundwater, and these may bring aboutsignificant changes; however, such processes rarely, if ever, totally convert organiccompounds to harmless end products or mineralized compounds in nature. Many ofthese nonenzymatic reactions only bring about a slight modification of the moleculeso that the product is frequently similar in structure, and often in toxicity, to theprecursor compound.

However, biological processes may modify organic molecules at the site of theirrelease or during their transport. Such biological transformations, which involveenzymes as catalysts, frequently bring about extensive modification in the structureand toxicological properties of pollutants or potential pollutants. These biotic pro-cesses may result in the complete conversion of the organic molecule to inorganicproducts, cause major changes resulting in new organic products, or occasionallylead to only minor modifications. The available body of information suggests thatthe major agents causing the biological transformations in soil, sediment, surfaceand groundwater, and many other sites are the microorganisms that inhabit theseenvironments.

The earth is thought to be about 4.6

×

10

9

years (4.6 eons) old.

2

The originalatmosphere surrounding the earth was reducing and probably included the gasesCH

4

, CO

2

, CO, NH

3

and H

2

O. Although abiotic organic synthesis probably occurredsince the earth’s beginnings, life probably did not appear until about 0.5–1 billionyears later, according to present thinking.

The first form of life that was established on the “infant” earth was anaerobic.

1

As anaerobic life became more firmly established, the organic nutrients must havebegun to be depleted at a faster rate than they could be replenished by abioticsynthesis. Hence, an alternative mechanism for producing organic matter wasrequired to sustain life. The subsequent evolutionary developments led to the emer-gence of photosynthesis and eventually resulted in the emergence of aerobic het-erotrophic organisms. These aerobic organisms ended up much more efficient thantheir anaerobic counterparts in sustaining life.



Billions of years of evolution by Mother Nature have shown that the naturalcommunities of microorganisms in the various habitats have an amazing physiolog-ical versatility. These communities are adaptable, flexible, versatile, and robust. Theyare able to metabolize and often mineralize an enormous number of organic mole-cules. Probably every natural product, regardless of its complexity, is degraded byone or another species in some particular environment; if not, this long after theappearance of life on earth, such compounds would have accumulated in enormousamounts.

The compounds that caught the most attention in the remediation industry,initially during the 1970s and 1980s, were the BTEX compounds released viapetroleum spills — natural products formed as the result of decomposition of plantsand other organic materials over millions of years. It has been proven during thelast decade that the BTEX compounds will naturally attenuate in the groundwaterthrough microbial degradation. Although certain bacteria and fungi act on a broadrange of organic compounds, no organism known to date is sufficiently omnivorousto destroy a very large percentage of the natural chemicals.

2

Bioremediation is now a widely accepted technique for contaminant cleanup.But a few short decades ago, its use for anything as effective as the

in situ

cleanupof groundwater contamination was considered laughable. “At that time, there was amyth, widely held by the geological and hydrological community, that the subsurfacewas sterile, that there were no bacteria, and therefore no biological processes ofconsequence”.

3

This thinking was mainly due to the information available from thetextbooks at that time.

Microbial ecology is the study of interrelationships between different microor-ganisms; among microorganisms, plants, and animals; and between microorganismsand their environment. Microbial biogeochemistry is the study of microbially cata-lyzed reactions and their kinetics with emphasis on environmental mass transfer andenergy flow.

In subsequent chapters, this book summarizes and systematizes current under-standing of abiotic and biotic transformations of organic and inorganic pollutants inthe natural environment. Knowledge of abiotic transformations can provide a con-ceptual framework for understanding biologically mediated transformations. Mostabiotic transformations are slow, but they can still be significant within the timescales commonly associated with groundwater movement. In contrast, biotic trans-formations typically proceed much faster, provided the biogeochemical environmentis conducive to mediate such transformations.

The ability to predict the behavior of a chemical substance in a biological orenvironmental system largely depends on knowledge of the physical-chemical prop-erties and reactivity of that compound or closely related compounds. Chemicalproperties frequently used in environmental fate assessment include melting/boilingtemperature, vapor pressure, various partition coefficients, water solubility, Henry’sLaw Constant, sorption coefficient, and diffusion properties. Reactivities by pro-cesses such as biodegradation, hydrolysis, photoysis, and oxidation/reduction arealso critical determinants of environmental fate. Unfortunately, measured valuesoften are not available and, even if they are, the reported values may be inconsistentor of doubtful validity. In this situation it may be appropriate or even essential to



use estimation methods. The evolution of understanding of half-lives of chlorinatedaliphatic compounds and the refinement of those values by precisely measured valueshave been remarkable during the last decade. Half-lives which were estimated to bein the two to three year range have been measured in the field to be in the range ofthree to six months. Later chapters describe the marked difference in the acceptedhalf-lives during the last decade.

1.4 EVOLUTION OF REMEDIATION TECHNOLOGIES

Remediation technologies have undergone many changes over the last two decadesduring which they have been applied to clean up subsurface and hazardous wastecontamination problems. These changes have occurred at a relatively rapid pace; duringthis period some of the most profound changes have occurred in how we apply remedialtechnologies as a result of pressure from the industry to continuously improve technicalefficiency and cost effectiveness of the preferred technologies.

Initially contaminated groundwater was a driving concern because it was mobileand, as a result, transported the liability off site. Also, the need to contain thecontamination on site led to universal application of pump and treat systems forsource control and mass removal. A decade of experience has taught that pump andtreat is not the solution and, in fact, is an inefficient technology for fast and cheapsite cleanup.

The realization that mass removal efficiencies can be significantly enhancedusing air as an extractive media instead of water led to the development and appli-cation of

in situ

extractive technologies such as soil vapor extraction and

in situ

airsparging. While it can be argued that the initial motive for applying these technol-ogies has been one of saving money, the end result is much quicker cleanup timesto more acceptable cleanup levels (Figure 1.3). This win-win situation for the entireremediation industry fostered continuous innovation, which led to 1) faster, cheapersolutions, 2) less invasive

in situ

technologies, and 3) technologies complementaryto the natural environment which took advantage of nature’s capacity to degrade thepollutants. Thus holistically, environmentally, and economically sound and sustain-able solutions were provided.

Figure 1.4 illustrates the evolutionary reduction in remediation costs from thelate 1970s to the present time.

Ex situ

extractive techniques such as pump and treatsystems were replaced by

in situ

extractive techniques, namely, soil vapor extraction(SVE) and

in situ

air sparging. Subsequently these

in situ

extractive techniques gaveway to

in situ

nonextractive techniques such as funnel and gate systems and, even-tually, to

in situ

mass destruction techniques such as

in situ

reactive zones (IRZ) asthe preferred remediation technologies. This evolutionary pattern has focusedtowards more natural solutions and/or enhancing existing subsurface biogeochemicalconditions that contribute to remediation.

The most recent shift occurred approximately 5 years ago, with the recognitionand demonstrated value of natural mechanisms that contributed towards the contain-ment, control, and mass reduction of contaminants in soil and groundwater. Under ahost of names — including natural attenuation, bioattenuation, natural remediation,



monitored natural attenuation (MNA) — this remediation approach has taken root asa viable remediation approach at the appropriate site and under the right biogeochem-ical conditions. Used in conjunction with already ongoing remediation systems or asa stand-alone remedy, MNA can increase significantly the probability of a successful,cost-effective, and well-documented restoration of a contaminated site.

The development of

in situ

reactive zones (IRZ), which are engineered

in situ

anaerobic or aerobic systems, is essentially an outgrowth of the efforts to enhance thenatural processes which contribute towards degradation of many contaminants. Forexample, the use of an engineered IRZ to reductively dechlorinate chlorinated aliphatichydrocarbons, such as PCE and TCE, in essence, enhances the rate of natural degra-dation by providing the optimum biogeochemical conditions (Figure 1.5).

At many contaminated sites, the bulk of the contaminant mass may still bepresent in what remediation professionals call “source areas.” Even though the plumelength has reached a stable equilibrium and the contaminant concentrations havereached steady or declining concentrations at the compliance points, it may bedesirable to enhance the rates of natural degradation if the plume has crossed theproperty boundary (Figure 1.6a). Surgical reduction of the mass at the source areasand enhancement of natural degradation along the property boundary will enablesuch properties to be restored within a reasonable time frame (Figure 1.6b). Theduration of the containment IRZ at the property boundary will be significantly longerif mass removal is not accomplished at the source area.

Figure 1.3

Evolution of

in situ

remediation technologies and improvements in efficiencies.

Clean-Up Standards

Only WhenContaminantsAre AerobicallyBiodegradable

Time

Con

cent

ratio

n

MNA with Source Reduction

Conventional Pump and Treat

In Situ Reactive Zones (IRZ)

In Situ Air Sparging

MNA - Monitored Natural Attenuation



Figure 1.4

Evolution reduction in remediation costs.

Figure 1.5

Implementation of an IRZ for enhanced biodegradation has an impact on the timetowards closure in comparison to reliance on MNA.

Ex Situ ExtractiveTechniques

Late 1970s - Early 1980s

Cos

t ($)

Capital Costs

O&M Costs

Monitored Natural Attenuation

In Situ ExtractiveTechniques

Early 1980s - Late 1980s

In Situ ExtractiveTechniques

Early 1990s - Present

In Situ Mass DestructionTechniques

Mid 1990s - Present

MNACurrent

MNA

Cos

t ($)

Time

In Situ Reactive Zones


IRZ

MNA

Creation of IRZ

Natural Rate of Decline

Enhanced Rate of Decline



Figure 1.6a

Implementation of a containment IRZ and a source reduction IRZ to reduce thecleanup time.

Figure 1.6b

Reduction in cleanup time as a result of enhanced rate of mass removal.

Engineered IRZ

Stable Plume

CompliancePoints

Property Boundary

Source Area

In Situ Reactive ZonesIRZC

once

ntra

atio

n at

Com

plia

nce

Poi

nt(s

)

Time

In Situ Reactive ZonesIRZ

Creation of IRZ



The evolutionary trend towards more natural remediation solutions should not besurprising due to the recognition for decades, even centuries, of the value of naturalwetlands to buffer the effects of human activity in waterways. Today engineeredwetlands, phyto-covers and phytoremediation are our attempts to mimic natural sys-tems by “engineering” remediation solutions using nature as the material of construc-tion — trees and microorganisms instead of pumps and air compressors.

REFERENCES

1. Ehrlich, H.L.,

Geomicrobiology

, Marcel Dekker Inc., New York, 1981.2. Alexander, M.,

Biodegradation and Bioremediation

, 2nd ed., Academic Press, NewYork, 1999.

3. Harvey, M.A., Quotation by John Wilson in Germ Warfare,

Environ. Prot.

, 23-26,October 1999.



13

CHAPTER

2

Contaminant and EnvironmentalCharacteristics

CONTENTS

2.1 Introduction ....................................................................................................142.2 Contaminant Characteristics ..........................................................................18

2.2.1 Physical/Chemical Properties ............................................................182.2.1.1 Boiling Point.......................................................................182.2.1.2 Vapor Pressure ....................................................................182.2.1.3 Henry’s Law Constant ........................................................192.2.1.4 Octanol/Water Partition Coefficients ..................................202.2.1.5 Solubility in Water..............................................................202.2.1.6 Hydrolysis ...........................................................................222.2.1.7 Photolytic Reactions in Surface Water...............................24

2.2.2 Biological Characteristics ..................................................................262.2.2.1 Cometabolism .....................................................................272.2.2.2 Kinetics of Biodegradation.................................................32

2.3 Environmental Characteristics .......................................................................382.3.1 Sorption Coefficient ...........................................................................38

2.3.1.1 Soil Sorption Coefficients...................................................432.3.1.2 Factors Affecting Sorption Coefficients .............................48

2.3.2 Oxidation-Reduction Capacities of Aquifer Solids ...........................512.3.2.1 pe and pH............................................................................512.3.2.2 REDOX Poise .....................................................................522.3.2.3 REDOX Reaction ...............................................................53

References................................................................................................................58



Water is scientifically very different in comparison to other liquids. With its rareand distinctive property of being denser as a liquid than as a solid, it is different.Water is different in that it is the only chemical compound found naturally in solid,liquid, or gaseous states at ambient conditions. Water is sometimes called theuniversal solvent. This is a fitting name, especially when you consider that water isa powerful reagent, which is capable in time of dissolving everything on earth.

2.1 INTRODUCTION

The primary management goal during remediation of a contaminated site is toobtain closure, that is, to achieve a set of conditions that is considered environmen-tally acceptable and which will ensure that no future action will be required at thesite. A substantial ongoing national debate associated with site closure centers onthe definition of “how clean is clean” for contaminated subsurface media. The keyissue in this debate is, “What concentration of residual contaminant in the subsurface,particularly adsorbed to the soil, is environmentally acceptable?”

In this context, the term

contaminant availability

becomes an important concept;it refers to the rate and extent to which the chemical will be released from thesubsurface into the environment and/or is bioavailable to ecological and humanreceptors. The dissemination of a contaminant after its release into the environmentis determined by its partition among the water, soil and sediment, and atmosphericphases, and its degradability via biotic and/or abiotic means. These processes deter-mine both the impact and the extent of its dissemination.

Within the context of overall site management, measurements of

contaminantavailability

are not intended to replace other approaches, required regulatorily, toachieve site closure; rather, they are meant to broaden the range of options or toolsavailable to environmental professionals. This chapter will discuss the basis andparameters for the development of procedures and determination of partitioning,transport, and fate of various types of contaminants in the subsurface. These param-eters will also provide the basis for the development of the tools to determinecontaminant availability and incorporate those estimations into a decision frameworkto define

environmentally acceptable endpoints

for the different media. In addition,how these parameters and characteristics influence contaminant fate and transportand how they impact remediation system design are woven together in the discus-sions in subsequent chapters.

The reactions that contaminants undergo in the natural environment, such as sorp-tion, desorption, precipitation, complexation, biodegradation, biotransformation,hydrolysis, oxidation-reduction, and dissolution, are critical in determining their fateand mobility in the subsurface environment. Reaction time scales can vary from micro-seconds for many ion association reactions microseconds and milliseconds for someion exchange and sorption reactions, to days, weeks, or months for some microbiallycatalyzed reactions, or years for many mineral solution and crystallization reactions.

Both transport and chemical reaction processes can affect the reaction rates inthe subsurface environment. Transport processes include: (1) transport in the solutionphase, across a liquid film at the particle/liquid interface (film diffusion), and in


CONTAMINANT AND ENVIRONMENTAL CHARACTERISTICS 15

liquid-filled macropores, all of which are nonactivated diffusion processes and occurin mobile regions; (2) particle diffusion processes, which include diffusion of sorbateoccluded in micropores (pore diffusion) and along pore-wall surfaces (surface dif-fusion) and diffusion processes in the bulk of the solid, all of which are activateddiffusion processes (Figure 2.1).

1

Pore and surface diffusion within the immediateregion can be referred to as intra-aggregate (intraparticle) diffusion and diffusion inthe solid can be called interparticle diffusion. The actual chemical reaction at thesurface, e.g., adsorption, is usually instantaneous. The slowest of the chemicalreaction and transport process is the ratelimiting reaction.

As an introduction to the various organic compounds which end up as contam-inants once discharged into the environment, Table 2.1 gives the basic structure ofthe different compounds.

Figure 2.1

Transport processes in solid-liquid soil reactions (adapted from Sparks, 1998).

Liquid (Groundwater)

Film Solid (Soil Grain)

Transport in the Soil Solution (Macro Pores)Transport Across a Liquid Film at the Solid-Liquid InterfaceTransport in a Liquid-Filled MacroporeDiffusion of a Sorbate at the Surface of the SolidDiffusion of a Sorbate Occluded in a MicroporeDiffusion in the Bulk of the Solid

123456

1 2

3

44 5

6 6



C

Non

e

Fun

ctio

nal

Gro

upG

ener

alF

orm

ula

Gen

eral

Nam

eF

orm

ula

IUPA

C N

ame

Com

mon

Nam

e

Exa

mpl

e

C CC

Cl

Br

OH

O NH

NR

X2 2+-

C H

2n+

2n n

2nC

H n2n

-2C

H

Cl

R R R R

Br

OH

OR

RN

H

-+

4R

N X

1

Alk

ane

Alk

ene

Alk

yne

Chl

orid

e

Bro

mid

e

Alc

ohol

Eth

er

Am

ine

Qua

tern

ary

2

amm

oniu

msa

lt

Eth

ane

Eth

ene

Eth

yne

Chl

oroe

than

e

Bro

mom

etha

ne

Eth

anol

Eth

oxy e

than

e

1-A

min

opro

pane

Dec

yltr

imet

hyl-

Am

mon

ium

chlo

ride

chlo

ride

Am

mon

ium

Dec

yltr

imet

hyl-

Pro

pyla

min

e

Die

thyl

eth

er

Eth

yl a

lcoh

ol

Met

hyl b

rom

ide

Eth

yl c

hlor

ide

Ace

tyle

ne

Eth

ylen

e

Eth

ane

3

3C

H C

H

2

3 CH

H C

2

CH

HC

CH

CH

Cl

22

3C

H B

r

23

CH

CH

OH

23

CH

CH

OC

H C

H

23

CH

CH

CH

NH

23

CH

(C

H )

N(C

H )

Cl

23

22

+-

93

3

OC H

HCO

R

CC

RO

CR

O CO

H

O CR

OH

Ald

ehyd

e

Ket

one

Car

boxy

lic a

cid

Pro

pana

l

2-B

utan

one

Eth

anoi

c ac

id

Pro

pion

alde

hyde

keto

neM

ethy

l eth

yl

Ace

tic a

cid

CH

CH

CH

O

H

CH

OC

H C

H C

OH

CH

CO

32

32

3

3

Tab

le 2

.1 S

om

e C

om

mo

n F

un

ctio

nal

Gro

up

s.

(Con

tinue

d)



Tab

le 2

.1 (

Co

nt.

)

O CC

HO

C H

Ace

tic a

cid

Eth

yl e

than

oate

Est

er

OR

'R

CO

OR

'CO

Ace

toni

trile

Nitr

omet

hane

Met

hyl m

erca

ptan

Dim

ethy

l dis

ulfid

eD

imet

hyl d

isul

fide

Met

hane

thio

l

Nitr

o met

hane

Eth

ane n

itrile

(sul

fide)

Dis

ulfid

e

Thi

ol

Nitr

o

Nitr

ile

1

RRR

SH

NO

Exa

mpl

e

Com

mon

Nam

eIU

PAC

Nam

eF

orm

ula

Gen

eral

Nam

eG

ener

alF

orm

ula

Fun

ctio

nal

Gro

up

25

Ace

tam

ide

Eth

anam

ide

2

O CN

H

O CR

NH

Am

ide

NH

CH

CO

32

2

3

O CC

HC

l

Aci

d ch

lorid

e

Cl

RCO

Cl

COE

than

oyl c

hlor

ide

Ace

tyl c

hlor

ide

3

O CC

H

Aci

d an

hydr

ide

RCO

COE

than

oic

anhy

drid

eA

cetic

anh

ydrid

e

O

O CO

O CR

OC

HCO

3

CN

SS

NC

22

NO

SH

RT

hioe

ther

Dim

ethy

l thi

oeth

erD

imet

hyl s

ulfid

eS

SR

RS

SR

N3 3

NO

CH

2

CH

SH

3 3C

HS

3C

H

CH

3S

CH

3S

O SR

OH

O SO

H

Sul

foni

c ac

id

OH

CH

SO

3

Met

hane

sulfo

nic

acid

Met

hane

sulfo

nic

acid

OO

O

O S

O SR

Sul

foxi

deD

imet

hyl s

ulfo

xide

Dim

ethy

l sul

foxi

de

R3

CH

CH

SO

3

OSO

O

RSO

Sul

fone

O3

O SC

H

Dim

ethy

l sul

fone

Dim

ethy

l sul

fone

RC

H3

The

ital

iciz

ed p

ortio

n in

dica

tes

the

grou

p.A

prim

ary

(1°)

am

ine;

ther

e ar

e al

so s

econ

dary

(2°

), R

NH

, and

tert

iary

(3°

), R

N, a

min

es.

Ano

ther

nam

e is

pro

pana

min

e.

1 2 32

3



2.2 CONTAMINANT CHARACTERISTICS

2.2.1 Physical/Chemical Properties

2.2.1.1 Boiling Point

The boiling point is defined as

the temperature at which a liquid’s vapor pressureequals the pressure of the atmosphere on the liquid

.

2

If the pressure is exactly 1atmosphere (101,325 Pa), the temperature is referred to as “the normal boiling point.”Pure chemicals have a unique boiling point, and this fact can be used in somelaboratory investigations to check on the identity and/or purity of a material. Mixturesof two or more compounds have a boiling point range.

For organic compounds, boiling points range from –162 to over 700

°

C, but formost chemicals of interest the boiling points are in the range of 300 to 600

°

C.

2

Having a value for a chemical’s boiling point, whether measured or estimated, issignificant because it defines the uppermost temperatures at which the chemical canexist as a liquid. Also, the boiling point itself serves as a rough indicator of volatility,with higher boiling points indicating lower volatility at ambient temperatures. Theboiling point is associated with a number of molecular properties and features. Mostimportant is molecular weight; boiling points generally increase with this parameter.Next is the strength of the intermolecular bonding because boiling points increasewith increasing bonding strength. This bonding, in turn, is associated with processesand properties such as hydrogen bonding, dipole moments, and acid/base behavior.

2.2.1.2 Vapor Pressure

The vapor pressure of a chemical is the pressure its vapor exerts in equilibriumwith its liquid or solid phase.

2

Vapor pressure’s importance in environmental workresults from its effects on the transport and partitioning of chemicals among theenvironmental media (air, water, and soil). The vapor pressure expresses and controlsthe chemical’s volatility. The volatilization of a chemical from the water surface isdetermined by its Henry’s Law Constant, which can be estimated from the ratio ofa chemical’s vapor pressure to its water solubility. The volatilization of a chemicalfrom the soil surface is determined largely by its vapor pressure, although this istempered by its sorption to the soil matrix and its Henry’s Law Constant betweenthe soil water content and air.

A substance’s vapor pressure determines whether it will occur as a free moleculein the vapor phase or will be associated with the solid phase. For volatile substancesthat boil at or below 100

°

C, the vapor pressure is likely to be known, but, for manyhigh-boiling substances with low vapor pressure, the value may be unknown orpoorly known. An estimation procedure may be needed to help convert the knownvapor pressure at the normal boiling point (i.e., 1 atmosphere) to the vapor pressureat the lower temperatures of environmental importance. For some of these highboiling compounds, the actual boiling point may also be unknown, since the sub-stance may decompose before it boils.



2.2.1.3 Henry’s Law Constant

Along with the octanol-water and octanol-air partition coefficients, the Henry’sLaw Constant determines how a chemical substance will partition among the threeprimary media of accumulation in the environment, namely air, water, and organicmatter present in soils, solids, and biota. Volatile organic compounds (VOCs) withlarge values of Henry’s Law Constant evaporate appreciably from soils and water,and their fate and effects are controlled primarily by the rate of evaporation and therate of subsequent atmospheric processes. For such chemicals, an accurate value ofthis parameter K

AW

is essential. Even a very low value of K

AW

for example, 0.001,can be significant and must be known accurately, because the volume of the acces-sible atmosphere is much larger than that of water and soils by at least a factor of1000; thus even a low atmospheric concentration can represent a significant quantityof chemical. Further, the rate of evaporation from soils and water is profoundlyinfluenced by K

AW

because that process involves diffusion in water and air phasesin series, or in parallel, and the relative concentrations which can be established inthese phases control these diffusion rates.

2,3

Accurate values of K

AW

are thus essential for any assessment of the behavior ofexisting chemicals or prediction of the likely behavior of new chemicals. Air-waterpartitioning can be viewed as the determination of the solubility of a gas in wateras a function of pressure, as first studied by William Henry in 1803. A plot ofconcentration or solubility of a chemical in water expressed as mole fraction x, vs.partial pressure of the chemical in the gaseous phase P, is usually linear at low partialpressures, at least for chemicals which are not subject to significant dissociation orassociation in either phase. This linearity is expressed as

Henry’s Law

. The Henry’sLaw Constant (H) which in modern SI units has dimensions of Pa/(mol fraction).For environmental purposes, it is more convenient to use concentration units in waterC

W

of mol /m3 yielding H with dimensions of Pa m

3

/mol.

P (Pa) = H (Pa m

3

/mol) C

W

(mol/m

3

) (2.1)

The partial pressure can be converted into a concentration in the air phase C

A

byinvoking the ideal gas law:

C

A

= n/V = P/RT (2.2)

Where n is mols, V is volume (m

3

), R is the gas constant (8.314 Pa m

3

/mol K) andT is absolute temperature (K).

C

A

= P/RT = (H/RT) C

W

= K

AW

C

W

(2.3)

The dimensionless air-water partition coefficient K

AW

(which can be the ratio in unitsof mol/m

3

or g/m

3

or indeed any quantity/volume combination) is thus (H/RT).A plot of C

A

vs. C

W

is thus usually linear with a slope of K

AW

as Figure 2.2illustrates. For organic chemicals which are sparingly soluble in water, these con-centrations are limited on one axis by the water solubility and on the other by the



maximum achievable concentration in the air phase which corresponds to the vaporpressure, as Figure 2.2 shows. To the right of or above the saturation limit, a separateorganic phase is present. Strictly speaking, this saturation vapor pressure is that ofthe organic phase saturated with water, not the pure organic phase.

2,3

2.2.1.4 Octanol/Water Partition Coefficients

The usefulness of the ratio of the concentration of a solute between water andoctanol as a model for its transport between phases in a physical or biological systemhas long been recognized.

2,4,5

It is expressed as

P

OCT

=

C

O

/C

W

=

K

OW

. This ratio isessentially independent of concentration, and is usually given in logarithmic terms(log P

OCT

or log K

OW

). The importance of bioconcentration in environmental hazardassessment and the utility of this hydrophobic parameter in its prediction led to anintense interest in the measurement of P

OCT

and also its prediction from molecularstructure. (So many calculation methods have been published in the last five yearsthat it is not possible to examine them all in detail.)

2.2.1.5 Solubility in Water

Solubility in water is one of the most important physical chemical properties ofa substance, having numerous applications to the prediction of its fate and its effectsin the environment. It is a direct measurement of hydrophobicity, i.e., the tendencyof water to

exclude

the substance from solution. It can be viewed as the maximumconcentration which an aqueous solution will tolerate before the onset of phaseseparation.

Figure 2.2

Description of Henry’s Law Constant.

Con

cent

ratio

n in

Air

CA

Concentration in Water Cw

(Vapor Pressure/RT)

[Solubility ofCompound]

Slope = Kaw= H/RT



Substances which are readily soluble in water, such as lower molecular weightalcohols, will dissolve freely in water if accidentally spilled and will tend to remainin aqueous solution until degraded. On the contrary, sparingly soluble substancesdissolve more slowly and, when in solution, have a stronger tendency to partitionout of aqueous solution into other phases. They tend to have larger air–water partitioncoefficients or Henry’s Law Constants, and they tend to partition more into solidand biotic phases such as soils, sediments, and fish. As a result, it is common tocorrelate partition coefficients from water to those media with solubility in water.

Solubility normally is measured by bringing an excess amount of a pure chemicalphase into contact with water at a specified temperature, so that equilibrium isachieved and the aqueous phase concentration reaches a maximum value. It is rareto encounter a single compound as the contaminant present in the groundwater at acontaminant site.

(2.4)

where,

C

i*

= equilibrium solute concentration for component

i

in the mixtureC

i0

= equilibrium solute concentration for component

i

as a pure compoundx

i

= mole fraction of compound

i

in the mixture

γ

i

= activity coefficient of compound

i

in the mixture.

Possible equilibrium situations may exist, depending on the nature of the chem-ical phase, each of which requires separate theoretical treatment and leads to differentequations for expressing solubility. These equations form the basis of the correlationsdiscussed later.

Single compound is an immiscible liquid (e.g., Benzene)

C

*

= C

o

x

γ

(2.5)

In this case, C

*

is also C

°

. Thus the product x

γ

is 1.0 and x is 1/

γ

. Sparinglysoluble substances act in such a way because the value of

γ

is large.

2

For example, at 25

°

C benzene has a solubility in water of 1780 g/m

3

or 22.8mol/m.

3

Since 1 m

3

of solution contains approximately 10

6

/18 mol water (1m

3

is10

6

g and 18 g /mol is the molecular mass of water), the mole fraction x is22.8/(10

6

/18) or 0.00041. The activity coefficient

γ

is thus 2440; i.e., a benzenemolecule in aqueous solution behaves as if its concentration were 2440 timeshigher.

Substances such as polychlorinated biphenyls (PCBs) can have activity coeffi-cients exceeding 1 million. Hydrophobicity thus is essentially an indication of themagnitude of

γ

. Some predictive methods focus on estimating

γ

, from which solu-bility can be deduced.

C C xi i i i* = 0 γ



Single compound is a miscible substance (e.g., Ethanol)

If the activity coefficient is relatively small, i.e., < 20, it is likely that the liquidis miscible with water and no solubility can be measured. The relevant descriptorof hydrophobicity in such cases is the activity coefficient. Correlations of otherenvironmental partitioning properties with solubility are then impossible.

2

Solubility is a function of temperature because both vapor pressure and

γ

aretemperature dependent. Usually

γ

falls with increasing temperature, thus solubilityincreases. This implies that the process of dissolution is endothermic. Exceptionsare frequent and in some cases, such as benzene, there may be a solubility minimumas a function of the temperature at which the enthalpy of dissolution is zero.

2

Under natural conditions, dissolved organic matter such as humic and fulvicacids frequently increases the apparent solubility. This is the result of sorption ofthe chemical to organic matter which is sufficiently low in molecular mass to beretained permanently in solution. The

true

solubility or concentration in the pureaqueous phase probably is not increased. The apparent solubility is the sum of thetrue or dissolved concentration and the quantity which is sorbed.

The solubility of substances such as carboxylic acids, which dissociate or formions in solution, is also a function of pH, a common example being pentachlorophe-nol. Data must thus be at a specific pH. Alternatively, the solubility of the parent(nonionic) form may be given, and pK

a

or pK

b

given, to permit the ratio of ionic tononionic forms to be calculated as

Ionic/non-ionic = 10

(pH–pKa)

(2.6)

The total solubility is then that of the parent and ionic forms.

2.2.1.6 Hydrolysis

Hydrolysis is a bond-making, bond-breaking process in which a molecule, RA,reacts with water, forming a new R–O bond with the oxygen atom from water andbreaking the R–A bond in the original molecule. One possible pathway is the directdisplacement of –A with –OH, as Equation 2.7 shows.

RA + H

2

O

→

ROH + HA (2.7)

Hydrolytic processes provide the baseline loss rate for any chemical in anaqueous environment. Although various hydrolytic pathways account for significantdegradation of certain classes of organic chemicals, other organic structures arecompletely inert. Strictly speaking, hydrolysis should involve only the reactantspecies water provides — that is, H

+

, OH

–

and H

2

O — but the complete pictureincludes analogous reactions and thus the equivalent effects of other chemical speciespresent in the local environment, such as HS

–

in anaerobic bogs, Cl

–

in seawater,and various ions in laboratory buffer solutions.

Hydrolysis results in reaction products that may be more susceptible to biodeg-radation, as well as more soluble. The likelihood that a halogenated solvent will



undergo hydrolysis depends in part on the number of halogen substituents. Morehalogen substituents on a compound will decrease the chance for hydrolysis reactionsto occur and will therefore decrease the rate of the reaction. Hydrolysis rates cangenerally be described using first-order kinetics, particularly in groundwater wherewater is the dominant nucleophile. Bromine substituents are more susceptible tohydrolysis than chlorine substituents. As the number of chlorine atoms in the mol-ecule increases,

dehydrohalogenation

may become more important.

12,47

Dehydrohalogenation

is an elimination reaction involving halogenated alkanesin which a halogen is removed from one carbon atom, followed by subsequentremoval of a hydrogen atom from an adjacent carbon atom. In this two-step reactionan alkene is produced. Although the oxidation state of the compound decreases dueto the removal of a halogen, the loss of a hydrogen atom increases it. This resultsin no external electron transfer, and there is no net change in the oxidation state ofthe reacting molecule.

47

Contrary to the patterns observed for hydrolysis, the like-lihood of dehydrohalogenation increases with the number of halogen constituents.Under normal environmental conditions, monohalogenated aliphatics apparently donot undergo dehydrohalogenation. The compounds 1,1,1-TCA and 1,1,2-TCA areknown to undergo dehydrohalogenation and are transformed to 1,1-DCE, which isthen reductively dechlorinated to VC and ethene. Tetrachloroethanes and pentachlo-roethanes are transformed to TCE and PCE via dehydrohalogenation pathways.

47

Methods to predict the hydrolysis rates of organic compounds for use in theenvironmental assessment of pollutants have not advanced significantly since thefirst edition of the Lyman Handbook.

8

Two approaches have been used extensivelyto obtain estimates of hydrolytic rate constants for use in environmental systems.

2

The first and potentially more precise method is to apply quantitative structure/activ-ity relationships (QSARs).

2,9

To develop such predictive methods, one needs a setof rate constants for a series of compounds that have systematic variations in structureand a database of molecular descriptors related to the substituents on the reactantmolecule. The second and more widely used method is to compare the targetcompound with an analogous compound or compounds containing similar functionalgroups and structure, to obtain a less quantitative estimate of the rate constant.

Predictive methods can be applied for assessing hydrolysis for simple one-stepreactions where the product distribution is known. Generally, however, pathways areknown only for simple molecules. Often, for environmental studies, the investigatoris interested in not only the parent compound but also the intermediates and products.Therefore, estimation methods may be required for several reaction pathways.

Some preliminary examples of hydrolysis reactions illustrate the very wide rangeof reactivity of organic compounds. For example, triesters of phosphoric acid hydro-lyze in near-neutral solution at ambient temperatures with half-lives ranging fromseveral days to several years,

10

whereas the halogenated alkanes such as tetrachlo-roethane, carbon tetrachloride, and hexachloroethane have half-lives of about 2hours, 50 years, and 1000 millennia (at pH = 7, and 25ºC), respectively.

11,12

On theother hand, pure hydrocarbons from methane through the PAHs are not hydrolyzedunder any circumstances that are environmentally relevant.

Hydrolysis can explain the attenuation of contaminant plumes in aquifers wherethe ratio of rate constant to flow rate is sufficiently high. Thus 1,1,1-trichloroethane



(TCA) has been observed to disappear from a mixed chlorinated hydrocarbon plumeover time, while trichloroethene and its biodegradation product cis-1,2-dichloroet-hene persist. The hydrolytic loss of organophosphate pesticides in sea water, asdetermined from both laboratory and field studies, suggests that these compoundswill not be long-term contaminants despite runoff into streams and, eventually,the sea.

2.2.1.7 Photolytic Reactions in Surface Water

Photolysis (or photolytic reaction) can be defined as

any chemical reaction thatoccurs only in the presence of light

. Environmental photoreactions necessarily takeplace in the presence of sunlight, which has significant photon fluxes only above295 nm in the near ultraviolet (UV) range, extending into the infrared region of theelectromagnetic spectrum.

2,13

Environmental photoreactions occur in surface waters,on solid ground, and in the atmosphere, sometimes rapidly enough to make themthe dominant environmental transformation processes for many organic compounds.In the atmosphere, for example, photooxidation, mediated by hydroxyl radical (OH

•

),is the dominant removal process for more than 90% of the organic compoundsfound there.

Photolytic reactions are often complex reactions that not only control the fateof many chemicals in air and surface water, but also often produce products withchemical, physical, and biological properties quite different from those of their parentcompounds: more water soluble, less volatile, and less likely to be taken up by biota.Photooxidation removes many potentially harmful chemicals from the environment,although occasionally more toxic products form in oil slicks and from pesticides.

14

Biogeochemical cycling of organic sulfur compounds in marine systems involvesphotooxidation on a grand scale in surface waters, as well as in the troposphere.

2

Environmental photoreactions can be divided into two broad categories of reac-tions: direct and indirect. A direct photoreaction occurs when a photon is absorbedby a compound leading to formation of excited or radical species, which can reactin a variety of different ways to form stable products. In dilute solution, rate constantsfor these reactions are the products of the rate constants for light absorption and thereaction efficiencies. An indirect photoreaction occurs when a sunlight photon isabsorbed by one compound or group of compounds to form oxidants of excitedstates, which then react with or transfer energy to other compounds present in thesame environmental compartment to form new products. For example, NO2 and O3

in air form hydroxyl radicals (OH•), and humic acids in water form singlet oxygenand oxyradicals, when they absorb sunlight photons. These oxidants react with otherchemicals in thermal (dark) reactions, and the rates for these processes follow simplebimolecular kinetics.

Direct Photoreactions: Only a small proportion of synthetic organic compoundsabsorb UV light in the sunlight region of the spectrum (above 295 nm) and thenphotolyze at significant rates.13 Most aliphatic and oxygenated compounds, such asalcohols, acids, esters, and ethers, absorb only in the far UV region (below 220 nm),



and simple benzene derivatives with alkyl groups or one heteroatom substituentabsorb strongly only in the far and middle UV region. Nitro or polyhalogenatedbenzenes, naphthalene derivatives, polycyclic aromatics and aromatic amines,nitroalkanes, azaolkanes, ketones, and aldehydes absorb sunlight between 300 and450 nm; polycyclic and azoaromatics (dyes), as well as quinones, also absorb visiblelight, in some cases to beyond 700 nm.2,13

The rate of a direct photoprocess depends only on the product of the rate of light(photon) absorption by compound C,(IA) and the efficiency with which the absorbedlight is used to effect reaction (quantum yield, Ø):13

(2.8)

Under most environmental conditions, chemicals are present in surface water orair at low concentrations, so their light absorbing properties lead to simple kineticexpressions for direct photolysis in water.13

Indirect Photoreactions: Indirect photolysis is most important for compoundsthat absorb little or no sunlight. Light absorption by chromophores (sensitizers) otherthan the compound of interest begin the process, forming intermediate (and transient)oxidants or excited states that affect chemical changes in the compound of inter-est.2,15,16 Examples of sensitizers that serve this purpose are dissolved organic matter(DOM or humic acid) and nitrate ion in water, and ozone and NO2 in the atmosphere.Transient species formed by indirect photoreactions in water include singlet oxygenand peroxy radicals, both of which are relatively selective and electrophilic. As aresult, only electron-rich compounds, such as phenols, furans, aromatic amines,polycyclic aromatic hydrocarbons (PAHs), and alkyl sulfides can undergo relativelyrapid indirect photoprocesses with these oxidants. Nitroaromatics, though not oxi-dized, appear to be sensitized by triplet DOM or scavenged by solvated electrons.Many of these compounds (e.g., PAHs, nitroaromatics, and aromatic amines) alsoundergo rapid direct photoreactions.2,16

By contrast, OH• radical, which dominates tropospheric photochemistry, oxidizesall classes of organic compounds (except perhalogenated compounds such as PCE),including alkanes, olefins, alcohols, and simple aromatics.160,166 Aqueous OH•. rad-ical, derived mainly from the photolysis of nitrate ion, plays an important role inconverting marine DOM to simpler carbonyl compounds, even though the averageconcentration is extremely low (<2 × 10–8).17 OH• also appears important in degradingsynthetic chemicals in a variety of nitrate-bearing freshwaters, where the OH• con-centrations appear to be one to two orders of magnitude higher.2,13

In many cases, detailed pathways for forming these oxidants and reductantsremain unclear, but identities of several of the transients are fairly well established.2,13

Transient species are transient because they react rapidly with themselves or with avariety of natural organic and metal species in natural waters,2 balancing formationrates to give low average concentrations.

Ratedcdt

= × ∅= Efficiency Photons Absorbed / time = IA



2.2.2 Biological Characteristics

It is generally conceded that biological reactions are of the greatest significancein determining the fate and persistence of organic compounds in most natural aquaticecosystems. It is essential at the start to make a clear distinction between biodegra-dation and biotransformation. Biodegradation is a process in which the destructionof a chemical is accomplished by the action of a living microorganism. Duringbiotransformation, on the other hand, only a restricted number of metabolic reactionsis accomplished, and the basic framework of the molecule remains essentiallyintact.18 Even though biodegradation and biotransformation, considered as alterna-tives, are not mutually exclusive.

Biodegradation can be categorized into three types that have importance in anecosystem setting:

Primary Biodegradation: biodegradation to the minimum extent necessary to changethe identity of the compound.

Ultimate Biodegradation: biodegradation to water, carbon dioxide, and inorganic com-pounds (if elements other than C, H, and O are present). This is also called miner-alization. Under anaerobic conditions, methane may be formed in addition to carbondioxide during fermentation reactions.

Acceptable Biodegradation: biodegradation to the minimum extent necessary to removesome undesirable property of the compound, such as toxicity. Conversion of vinylchloride to ethene is an example and in many instances this can also be consideredbiotransformation.

Although biological degradation conceivably might be accomplished by anyliving organism, available information indicates that, by far, the most significantbiological systems involved in ultimate biodegradation of contaminants are bacteriaand fungi. Critical and necessary conditions necessary for biodegradation of con-taminants to take place are summarized below:

• A microbial population must exist that has the necessary enzymes to bring aboutthe biodegradation.

• This population must be present in the environment where the contaminant ispresent.

• The contaminant must be accessible to the microorganisms having the requisiteenzymes, and most of the time this requires the contaminant to be available inthe dissolved phase.

• If the initial enzyme bringing about the degradation is extracellular, the bondsacted upon by that enzyme must be exposed for the catalytic enzyme to function.

• Should the enzyme catalyzing the initial degradation be intracellular, the moleculemust penetrate the surface of the cell to the internal sites where the enzyme acts.Alternatively, for the transformation to proceed further, the products of an extra-cellular reaction must penetrate the cell.

• Because the population or biomass of bacteria or fungi acting on many syntheticcompounds is initially small, conditions in the environment must be optimum toallow for proliferation of the potentially active microorganisms.



The initial concentration of the microbial population and the contaminant some-what affects the growth and proliferation; a lag period often occurs between theaddition of a chemical and the onset of biodegradation. This lag period, usuallyattributed to the need for acclimation,19,20 could result from enzyme induction, genetransfer or mutation, predation by protozoa, or growth in the population of respon-sible organisms.

The initial species present, their relative concentrations, the induction of theirenzymes, and their ability to acclimate once exposed to a chemical are likely to varyconsiderably, depending upon environmental parameters such as temperature, salin-ity, pH, oxygen concentration (aerobic or anaerobic), redox potential, concentrationand nature of various substrates and nutrients, concentration of heavy metals (tox-icity), and effects (synergistic and antagonistic) of associated microflora.21 Many ofthese parameters affect the biodegradation of contaminants in the environment.

One important parameter is the chemical substrate concentration. A number ofchemicals have biodegradation rates proportional to substrate concentration, butthere are also examples of thresholds and inhibitions.22 Recently, the bioavailabilityof the chemical to the catalytic enzyme has been identified as a major factor indetermining biodegradability in nature. Several studies have demonstrated that,although a chemical freshly added to soil is biodegraded at a moderate rate, thebiodegradation rate for some chemicals present in the soil sample for a long timeis very low.19 Thus, depending on the chemical, the longer a chemical remains inthe soil, the greater the potential for it to become sequestered and less bioavailable.

Some microorganisms are capable of biodegrading contaminants without popu-lation growth. In this process, known as “cometabolism,”19 the microorganismdegrades the contaminant from which it derives no carbon or energy; instead, it issustained on other organic substrates and nutrients.

2.2.2.1 Cometabolism

The transformation of an organic compound by a microorganism that is unableto use the substrate as a source of energy or as one of its growth substrate is termedcometabolism. The active populations thus derive no nutritional benefit from thesubstrates they cometabolize. Energy sufficient to fully sustain growth is not acquiredeven if the conversion is an oxidation and releases energy. In addition, the C, N, S,or P that may be in the molecule is not used as a source of these elements for growthand energy deriving purposes. Because of the prefix co, which often is appended toa word to indicate that something is done jointly or together (as in copilot orcooperate), there has been some debate regarding the use of the term cometabolism.Specifically, some classical microbiologists argue that the term should be appliedonly to circumstances in which a substrate that is not used for growth is metabolizedin the presence of a second substrate that is used to support multiplication.19 Accord-ing to this view, the transformation of a substance that is not used as a nutrient orenergy source but which occurs in the absence of a chemical supporting growthshould be designated by another term, for example, fortuitous metabolism. However,the prefix co also has another meaning, namely, “the same or similar.” The latterusage implies that the cometabolic transformation is similar to some other metabolic



reaction, which is consistent with one explanation for the phenomenon. Fortuitousmetabolism is, indeed, a more attractive term because it suggests an explanation forcometabolism, but the term will be used here as in the original definition, if for noother reason than it has gained wide acceptance.

The term cooxidation is sometimes used in studies of pure cultures of bacteria,referring specifically to oxidations of substrates that do not support growth in thepresence of a second compound that does support multiplication. Cooxidation hashistorical precedence in the debate but since it is restricted to oxidation, the worddoes not have sufficient breadth to include many reactions that are not oxidations.19

In summary, two types of reactions called cometabolism take place in the envi-ronment. In one, the cometabolized compound is transformed only in the presenceof a second substrate, which indeed may be the compound that supports growth.For heterotrophs, the energy-providing substrate is organic; for autotrophs, it isinorganic. In the other type, the compound is metabolized even in the absence of asecond substrate.

Important reasons for using the more general definition, and even for maintainingcometabolism as a term apart from bioconversion or biotransformation, are theenvironmental consequences of cometabolism. Cometabolic reactions have impactsin nature that are different from growth-linked biodegradations, and when the trans-formations take place, it is usually totally unclear whether the microorganisms door do not have a second substrate available on which they are growing.

A large number of chemicals are subject to cometabolism in nature. Amongcometabolic conversions that appear to involve a single enzyme, the reactions maybe hydroxylations, oxidations, denitrations, deaminations, hydrolyses, acylations, orcleavages of ether linkages; however, many of the conversions are complex andinvolve several enzymes. Some of the unique cometabolic reactions brought aboutby bacteria and fungi in nature come as no surprise in view of the vast array ofgrowth linked biological transformations that heterotrophic bacteria and fungi arecapable of in nature. An example which has no significant importance in contaminantremoval is the methane monooxygenase of methanotrophic bacteria which is ableto oxidize alkanes, alkenes, secondary alcohols, methylene chloride, chloroform,dialkyl ethers, cycloalkanes, and various aromatic compounds.19

Caution needs to be exercised in concluding that cometabolism is occurringmerely because an organism cannot be isolated from an environment in which achemical is undergoing a biological reaction.19 The isolation of bacteria acting onspecific substrates is usually performed by enriching the organism in a medium whenthe only C source is the test chemical, and the agar medium used to plate theenrichments contains that single organic supplement. Yet, many bacteria that areable to grow at the expense of that substrate will not develop in such simple mediabecause they require amino acids, B vitamins, or other growth factors. These essentialgrowth factors are not routinely included in such liquid media, and hence bacteriaand fungi needing them fail to proliferate. If the only organisms in the environmentable to metabolize a contaminant need these growth factors, no isolate will beobtained, and an erroneous conclusion will be reached that the compound is come-tabolized. If a chemical supports the growth of many species, some will undoubtedlyrequire no growth factors (these organisms are called prototrophs), and they will be



enriched and ultimately can be isolated. If the compound is acted on by only onespecies, in contrast, it is likely that the responsible organism will need amino acids,B vitamins, or other growth factors; these species are termed auxotrophs. Hence,the failure to isolate a bacterium or fungus capable of using the contaminant as thesole C source for growth is not sufficient evidence for cometabolism.

Mechanisms of Cometabolic Reactions: Several reasons have been advanced toexplain cometabolism, that is, why an organic chemical that is a substrate does notsupport growth but is converted to products that accumulate. Some of these reasonshave experimental support: (1) the initial enzyme converts the substrate to a productthat is not further transformed by other enzymes in the microorganism to yield themetabolic intermediates ultimately used for biosynthesis and energy production; (2)the initial substrate is transformed to products that inhibit the activity of later enzymesin mineralization or that suppress growth of the organisms; and (3) the organism needsa second substrate to bring about some particular reaction.

It could be speculated that the first explanation is the most common. The basisfor this explanation is the fact that many enzymes act on several structurally relatedsubstrates; thus, an enzyme naturally present in the cell will catalyze reactions andalter synthetic chemicals that are not typical cellular intermediates. These enzymesare not absolutely specific for their substrates. Consider a normal metabolic sequenceinvolving the conversion of A to B by enzyme a, B to C by enzyme b, and C to Dby enzyme c in a sequence that ultimately yields CO2 energy for biosynthesisreactions and intermediates that are converted to cell constituents.19

(2.9)

The first enzyme a may have a low substrate specificity and act on a moleculestructurally similar to A, namely, A.1 The product (B1) would differ from B in thesame way that A differs from A1. However, if enzyme b is unable to act on B1

(because the structural features controlling which substrate it modifies differ fromthose controlling the substrate specificity of enzyme a), B1 will accumulate:19

(2.10)

In addition, CO2 and energy will not be generated and, because cellular carbonis not formed, the organisms do not multiply. The formation of B1 is thus entirelyfortuitous.19

In instances where the contaminant concentration is high, cometabolism mayresult from the conversion of the parent compound to toxic products. In the sequencejust depicted, if the rate of reaction catalyzed by enzyme a is faster than the processcatalyzed by enzyme b, B will accumulate because it is not destroyed as readily asit is generated. For example, a strain of pseudomonas that grows on benzoate butnot 2-fluorobenzoate converts the latter to fluorinated products that are toxic.23 Theinhibitor that accumulates may affect a single enzyme important for the furthermetabolism of the toxin.

A B C D CO energy cell Ca b c

→ → → →→→ + +2 –

→ →a

A B1 1



In some instances, an organism may not be able to metabolize an organiccompound because it needs a second substrate to bring about a particular reaction.The second substrate may provide something that is present in insufficient supplyin the cells for the reaction to proceed — for example, an electron donor for thetransformation.19

The above explanation is linked to the existence of enzymes acting on more thana single substrate. Many enzymes are not absolutely specific for a single substrate.As a rule, they act on a series of closely related molecules, but some carry out asingle type of reaction on a variety of somewhat dissimilar molecules. The followingare examples of single enzymes acting on a range of substrates:

• Methane monoxygenase of methanotrophic bacteria: When grown on methane,methanol, or formate, these aerobic bacteria are able to cometabolize a large arrayof organic molecules, including several major pollutants. In each instance, methanemonooxygenase is the responsible catalyst. Other chlorinated aliphatic hydrocar-bons transformed by one such methanotroph, Methylosinus trichosporium, are cis-and trans-1,2-dichloroethylene, 1,1-dichloroethylene, 1,2-dichloropropane, and1,3-dichloropropylene.24 Apparently the same enzyme in other bacteria, aftergrowth on methane, will catalyze the oxidation of n-alkanes with two to eight Catoms, n-alkenes with two to six C atoms, and mono- and dichloroalkanes withfive or six C atoms, as well as dialkyl ethers and cycloakanes.19

• Toluene dioxygenase of a number of aerobic bacteria: This enzyme incorporatesboth atoms of oxygen from O2 (hence, it is a dioxygenase) into toluene as itcatalyzes the first step in the degradation of toluene by bacteria grown on thataromatic hydrocarbon (Figure 2.3a). However, that same enzyme has very lowspecificity and also is able to bring about the degradation of TCE,19,25,26 to convert2- and 3-nitrotoluene to the corresponding alcohols, and to hydroxylate the ringof 4-nitrotoluene.19,27

• Toluene monooxygenase of several aerobic bacteria: Differing from the dioxyge-nase, this enzyme incorporates only one atom of oxygen from O2 into toluene togive o-cresol (Figure 2.3b). However, because of this enzyme, bacteria can come-tabolize TCE, convert 3- and 4-nitrotoluenes to the corresponding benzyl alcoholsand benzaldehydes, and add hydroxyl groups to other aromatic compounds.19,28,29

• Oxygenase of propane-utilizing bacteria: Aerobes using propane as C and energysource for growth also have an oxygenase of broad specificity. This enzymecometabolizes TCE, vinyl chloride, and 1,1-di- and trans- and cis-1,2-dichloroet-hylene and has been recently known to degrade MtBE.19,30

• Ammonia monooxygenase of Nitrosomonas europaea: This bacterium, which is achemoautotroph whose energy source in nature is NH4

+ and whose C source isCO2, cometabolizes TCE, 1,1-dichloroethylene, various mono- and polyhaloge-nated ethanes, and a variety of monocyclic aromaticcompounds and thioethers, aswell as methyl fluoride and dimethyl ether.19,31

• An alkane hydroxylase hydroxylates a number of alkylbenzenes and linear,branched, and cyclic alkanes.19,23

• An alkane monooxygenase degrades TCE, vinyl chloride, and dichloroethylenesand propylenes.19

• Naphthalene dioxygenase acts on xylene, isomers of nitrotoluene, and ethylben-zene.19,33

• Biphenyl dioxygenase transforms several PCB congeners.19,34



Figure 2.3a Reactions catalyzed by toluene dioxygenase (adapted from Alexander, 1999).

CH3

TOLUENE

O2

2H

OH

OH

H

H

CH2OH

CH2OH

CH3

NO2NO2

CH3

2-NITROTOLUENE

NO2

NO2

CH3

3-NITROTOLUENE

CH3

4-NITROTOLUENE

CCI2 CICHTCE

OH

NO2

NO2

CH3

NO2

OH

OH

+

HC COOH + HCOOH + CI-

O



The organism containing these enzymes may be able to use one of several ofthe enzyme’s substrates for growth. However, many of the substrates are transformedbut do not support growth. The product of the reactions then accumulates.

Because cometabolism generally leads to a slow degradation of the substrate,attention has been given to enhancing its rate.19 The addition of a number of organiccompounds to the contaminated zone promotes the rate of cometabolism of a numberof chlorinated aliphatic and aromatic compounds and chlorinated phenols, but theresponses to such additions are not predictable. No relation is known to exist betweenthe metabolic pathways involved in the degradation of the added mineralized sub-strate and the compound that is cometabolized in these circumstances. The addedsubstrates were randomly chosen in these trials, and sometimes they do and some-times they do not stimulate cometabolism. In instances in which stimulation occurs,the benefit probably results from an unpredicted increase in the biomass of organ-isms, some of which fortuitously cometabolize the compound of interest.

An alternative approach is to add mineralizable compounds that are structurallyanalogous to the compound whose cometabolism one wishes to promote. Presum-ably, the microorganism that grows on the mineralizable compound containsenzymes transforming the analogous molecule that is cometabolized. This largerbiomass thus has more of the degradative enzyme than is present in the unsupple-mented water or soil. This method of analogue enrichment has been used to enhancethe cometabolism of PCBs by additions of biphenyl. The unchlorinated biphenylwas selected for addition to soil since it is mineralizable, nontoxic, and serves as aC source for microorganisms that are able to cometabolize PCBs.19

Analogue enrichment is a procedure that is similar to the usual means of isolatingbacteria that can cometabolize a compound. The enrichment culture contains a Csource that supports growth, and the pure cultures thus obtained also cometabolizestructurally related compounds that would not support growth. For example, bacteriaisolated on diphenylmethane and containing enzymes to degrade it also cometabolizechlorinated diphenylmethanes. Many of the latter do not sustain growth.19

2.2.2.2 Kinetics of Biodegradation

Impacts of the Environment: Soil, water, sediment, and wastewater environ-ments have different microbial populations and different available nutrients whichmay affect considerably the rate of biodegradation. For example, wastewater

Figure 2.3b Reactions catalyzed by toluene monooxygenase (adapted from Alexander,1999).

CH3CH3

TOLUENE o-CRESOL

OH



treatment plants may have high levels of nutrients and high microbial populationsthat may have been pre-exposed to a contaminant (acclimated microbial population),but the contact (retention) time is relatively short.

In contrast, marine waters are usually fairly low in nitrogen and phosphorouswhich may limit the biodegradation of chemicals (e.g., oil spills). Sediment oftenhas high levels of organic nutrients, but often is anaerobic, while surface waterstend, by comparison, to have low levels of organic nutrients.18 Digestor sludge fromwastewater treatment plants has high organic nutrients and is anaerobic. Surfacesoils have high concentrations of organic nutrients (depending on the type of soil),but this usually decreases with depth.

In the past, it was believed that groundwater aquifers were devoid of microbiallife. However, a number of studies have demonstrated that microorganisms are quiteplentiful in certain aquifers, and, in some instances, the bacterial concentration andactivity in aquifers may be higher than those in surface waters.19,35 In addition,availability of the chemical to the microbial population can be affected considerablyby the conditions of the microenvironment (e.g., organic concentration or claycontent may bind the chemical tightly).

A contaminant may become less available or essentially unavailable for biodeg-radation if it enters or is deposited in a micropore that is inaccessible even to themicroorganisms. These micropores may be filled entirely with water, as in sedimentsor groundwater aquifers, and the contaminant would have to move out of a microporeby diffusion to be accessible to bacteria for its destruction. The tortuous path thecontaminant molecule must traverse before it gets destroyed dramatically affectsbioavailability if the contaminant not only is physically remote from potentiallyactive microorganisms, but also is strongly sorbed to solid surfaces associated withthat remote micropore.19

Some organic compounds that persist in the subsurface often undergo a timedependent decline in bioavailability. Since this process is slow and time dependentit is appropriately called aging. This modification in bioavailability to microorgan-isms as a result of aging is also called sequestration.19 In the initial period, thecompound gradually disappears as a result of biodegradation, and possibly by othermass removal mechanisms, but little or none of the compound is destroyed after ithas resided in the soil or sediment for some time. Witness the finding that although80% of hexachlorobenzene deposited in the early 1970s in a lake bottom sedimentwas dechlorinated in the succeeding 20 years, all sediment cores still contained atleast 40 ppb of hexachlorebenzene. This time-dependent change in the rate ofdegradation, which has been observed with a number of insecticides, was the firstline of evidence for sequestration.19 Because these aged molecules are solventextractable, albeit by vigorous treatment, they are presumed to be present in anuncomplexed form and thus considered to be contaminants and subject to the reg-ulations and cleanup standards. In addition to the above mentioned contaminants,PAHs with three or more rings, such as phenanthrene, anthracene, fluorene, pyrene,chrysene, and others will also undergo sequestration.

It has been known for some time that it is increasingly difficult to remove stronglyhydrophobic compounds from soil with mild extractants as the residence time of



those compounds in the soil increases. This phenomenon is not restricted to soils;a similar decline in extractability is witnessed also from sediment samples.36

The amount of a contaminant that is sequestered increases with time. Expressedin another way, the percentage of the chemical that is bioavailable diminishes withincreasing persistence. This presumably occurs because more of the contaminant isdiffusing into inaccessible sites. However, after a period of time that varies with thesoil and the compound, sequestration of additional quantities slows and possiblystops. The reason for this rate of decline is not presently known.36

Based on the preceding discussion, it can be seen that the rates of biodegradationare likely to vary considerably, depending on the environment to which a contaminantis released, the type of contaminant(s), and the age of contamination. Also, the ratesunder different conditions may vary depending upon the type of chemical structure.For example, nitro aromatic compounds are usually fairly resistant to biodegradationunder aerobic conditions but are reduced rapidly to amines under anaerobic condi-tions. In contrast, degradation of benzene takes place significantly faster underaerobic conditions than under anaerobic conditions.

Structural Effects on Biodegradation: In addition to contaminant concentra-tion, chemical structure and physical/chemical properties have considerable impacton the rate and pathways of biodegradation. The chemical structure determines thepossible pathways that a substrate may undergo, generally classified as oxidative,reductive, hydrolytic, or conjugative. Figure 2.4 provides some examples of commonmicrobial degradation pathways.37 Recently, a computer program was developed thatwill predict the most probable metabolites, and another computer program was alsodeveloped that simulates the biodegradation of synthetic chemicals through thesequential application of plausible biochemical reactions.38,39

Over the years, structure/biodegradability “rules of thumb” have been devel-oped.40,41 Figures 2.4a and b summarize these. Some of these structure/biode-gradibility relationships have some biochemical mechanistic underpinings. Forexample, highly branched compounds frequently are resistant to biodegradationbecause increased substitution hinders β-oxidation, the process by which alkylchains and fatty acids usually are biodegraded. This structural relationship wasdiscovered in the 1950s when detergent scientists found that alkylbenzene sul-fonate (ABS) detergents passed through wastewater treatment plants causing foam-ing problems in rivers and streams. This problem was solved by switching fromthe highly branched ABS detergents to linear alkylbenzene sulfonate (LAS) deter-gents, thus illustrating the importance of understanding the relationship betweenstructure and biodegradability.

Few other rules of thumb have such mechanistic bases, but there are some generaltrends. Functional groups commonly seen by microorganisms in natural productsusually are degraded easily, probably because the microbes have had eons to developthe required enzyme systems in order to gain carbon and energy from the metabolism.Conversely, functional groups less common in nature or newly synthesized by manusually make a chemical more resistant to biodegradation. Aromatic substituentsthat are electron withdrawing (e.g., nitro groups and halogens) increase the persis-tence of a chemical, possibly by making it more difficult for enzymes to attack thearomatic ring, whereas electron donating functionalities (e.g., carboxylic acids,phenols, amines) generally increase biodegradation rates.



Physical/chemical properties affect the rate of biodegradation mostly by affectingbioavailability. Compounds which are sparingly soluble in water tend to be moreresistant to biodegradation, possibly due to an inability to reach the microbial enzymesite, a reduced rate of availability due to solubilization, or sequestration due toadsorption or trapping in inert material.

19,40

Biodegradation Rates:

The study of the kinetics of biodegradation in naturalenvironments is often empirical, reflecting the rudimentary level of knowledge aboutmicrobial populations and activity in these environments. An example of an empiricalapproach is the power rate model.

19

–dC/dt = kC

n

(2.11)

Figures 2.4a

Common microbial degradation pathways (after Boethling and MacKay, 2000).

OH

OHCI

CH3[CH2]x CO2H CH3[CH2]x CO2H

R CH3 R CH2OH

RR

R CH2HR CHO

R NH2 R NHOH R NO2R N=O

R NO2

R CN

R S R

R CO2H

R N=O R NH2R NHOH

RR

O

OOH

R R

R RS

R RS

R R R R

O

R CH2CI R CH2OH

R CCI3 R CO2H

R NH2

OO

O

S,OP S

O

S,O

S,OP O

S

S,O

S,OP S

S

S,O

S,OP OH

S,ORH2O

H2OR

S,O

S,OP S,O

S,OR R

RS,O

O

R O

HO

R O

R

β-oxidation

Methyl oxidation

Epoxide formation

Hydroxylation and ketone formation

Type of Reactions (not all steps are given)

Nitrogen oxidation

Nitro reduction

Nitrile/amide metabolism

Sulfur oxidation

Thiophosphate ester oxidation

Dehalogenation

Hydrolysis

Fatty acids and straight chainhydrocarbons (after oxidationof chain to carboxylic acid - see methyl oxidation)

Aromatic and aliphaticmethyl groups

Olefins

Aromatic to form phenolsand hydrocarbons to alcoholsand then ketones

Aromatic amines tonitroaromatic

Nitroaromatics aromaticamines (e.g., parathion)especially fast underanaerobic conditionsBromoxynil, Dichlobenil

Example of ChemicalsSubject to Reaction

Sulfides such as aldicarb

Thiophosphate pesticides

Aromatic and aliphatichalogens

Phosphate and carboxylicesters

L1282_C02_frame Page 35 Tuesday, June 19, 2001 12:57 PM


where

C

is substrate concentration,

t

is time,

k

is the rate constant for chemicaldisappearance, and n is a fitting parameter. This model can be fit to substrate-disappearance curves by varying n and k until a good fit is achieved. It is evidentfrom this equation that the rate is proportional to a power of the substrate concen-tration. The power-rate law provides a basis for comparison of different curves, butit gives no insight into the reasons for the shapes. Therefore, often it may have nopredictive ability. Moreover, investigators interested in kinetics do not always statewhether the model they are using has a theoretical basis or is simply empirical, andwhether constants in an equation have physical meaning or are only fitting param-eters.

19

An appropriate introduction to the kinetics of biodegradation is to consider

Figure 2.4b

Relationship between chemical structure and biodegradability (after Boethlingand MacKay, 2000).

≤3 rings ≥3 rings

OHOH

CI

CI

R RN

R RN

R RN

R

O

R

RR

H

R

RR

H

RR

R

H R R

R CH2OH

R CIR CO2H

R NH2

OH

CH3

CO2H

OMe

NH2

SO3H

CF3

CI

NO2

Br

R SO3R

R RN

R

R HN

H

R HN

R O

N

CH2H

3

OHOH

CI CI

Br

OHOH

CI

CI

CI

OH

Br

OH

CI

CI

CI

OH

CI

CI

CI

CICI

OH

CI

CI

CI

CI

N

CH3

N SR

N

N CH3

CH2H

3

N

CH3

N NCI

N

N CH3

More Biodegradable(Less Perisistent)

Less Biodegradable(More Persistent)

Branching

Aliphatic functional groups

Aliphatic amines

Halophenols

Polycyclic aromatics

Triazines

Aromatic functional groups (benzene, naphthalene, pyridine rings)



a pure culture of a single bacterial population growing on and degrading a single,soluble organic chemical, and to assume that no barriers exist between the substrateand the cells.

Biodegradation kinetics, under the conditions described previously, have beenreviewed in detail and a number of kinetic models proposed, including the use ofscreening tests for generating biodegradation kinetics. 19,42,43 Biodegradation ratestypically are interpreted through the Monod equation (Equation 2.12), which isanalogous to the Michaelis-Menten equation used in enzyme kinetics:

(2.12)

The parameters µ and µm refer to the growth rate and maximum growth rate,respectively, in the presence of substrate concentration S. Ks is the half-velocitycoefficient, i.e., the value of S at which µ = 0.5µm (Figure 2.5). Equation 2.13 canexpress the degradation rate of a substrate:

(2.13)

Figure 2.5 Relationship between the growth rate of bacteria vs. the substrate concentrationas described by the Monod kinetics model.

µ µ= [ ]+ [ ]m

s

S

K S

Gro

wth

Rat

e (µ

)

Substrate Concentration (s)

µmax

0.5 µmax

Ks

− [ ] = [ ][ ]+ [ ]( ) −d S

dt

S B

Y K Sm

s

µ



where B represents biomass and Y is the growth yield factor. The Monod equationassumes that the compound of interest sustains growth and is the only source ofcarbon and, thus, its applicability to a naturally contaminated environment may belimited. For example, for cometabolic processes with µm and Y defined as zero, theequation would not apply. The equation also ignores toxicity and makes no provisionfor acclimation. Experimentally, rates are measured either at low substrate concen-trations where Ks > [S] and Equation 2.13 simplifies to Equation 2.14, or athigh substrate concentrations where [S] > Ks and Equation 2.15 follows fromEquation 2.13:

(2.14)

(2.15)

For the former case (Equation 2.14), which is environmentally more relevant forlow contaminant concentrations typical of many sites, the rate obeys first-orderkinetics with respect to substrate and biomass (second-order overall), whereas in thelatter case (Equation 2.15), the kinetics have a first-order relationship to biomassbut are independent of substrate concentration (Figure 2.6). Methods for measuringbiomass, B, have varied widely, and, for studies involving mixed populations, inwhich only a fraction of the organisms can degrade the substrate, a means forquantifying the responsible fraction is not available.

The kinetics of cometabolism have received scant attention. If the microbialpopulations are neither growing nor declining and the concentration of substrate forcometabolism is below the Km of the active organisms, it is likely that the conversionwould be first-order. In a biofilm bioreactor inoculated with methane-oxidizingbacteria, the cometabolism of TCE, 1,1,1-trichloroethane, and cis- and trans-1,2-dichloroethylene is first-order at concentrations up to 1 mg/liter.19 However, inenvironments in which the transformations are slow, the C source for growth prob-ably is being depleted, so the kinetic patterns may change with time. Other modelshave been developed for cometabolism by nongrowing or growing populations.19

2.3 ENVIRONMENTAL CHARACTERISTICS

2.3.1 Sorption Coefficient

Sorption processes play a major role in determining the environmental fate andimpact of contaminants, affecting a variety of specific fate processes, includingsolubilization, volatilization, bioavailability, biodegradability, and hydrolysis. Sorp-tion coefficients quantitatively describe the extent to which an organic contaminantis distributed at equilibrium between an environmental solid (i.e., soil, sediment,

− [ ] = [ ][ ]d S

dt YKS Bm

s

µ

− [ ] = [ ]d S

dt YBmµ



suspended sediment, wastewater solids) and the aqueous phase with which it is incontact. Sorption coefficients depend on (1) the variety of interactions occurringbetween the solute and the solid and aqueous phases and (2) the effects of environ-mental variables such as organic matter quantity and type, clay mineral content andtype, clay to organic matter ratio, particle size distribution and surface area of thesorbent, pH, ionic strength, suspended particles or colloidal material, temperature,dissolved organic matter (DOM) concentration, solute and solid concentrations, andphase separation process.

Adsorption, absorption, and sorption are terms used to describe the uptake of asolute by another phase. Adsorption describes the concentration of a solute at theinterface of two phases, while absorption describes the process when a solute istransferred from the bulk state of one phase into the bulk state of the other phase.The term sorption is used frequently in environmental situations to denote the uptakeof a solute by a solid (soil or sediment or component of soil) without reference toa specific mechanism, or when the mechanism is uncertain.44

Sorption occurs when the free energy of the interaction between an environmen-tal solid sorbent and contaminant sorbate is negative. The sorption process can beeither enthalpy or entropy driven, depending on the properties of the solid sorbentand chemical solute. Enthalpy-related forces include van der Waals interactions,electrostatic interactions, hydrogen bonding, charge transfer, ligand exchange, directand induced dipole-dipole interactions, and chemisorption, while hydrophobic bond-ing or partitioning is considered the primary entropy driven force.2,44 Figure 2.7shows the polarity of the H2O molecule.

Figure 2.6 Concentration of substrate vs. time for zero-, first-, and second order biodegra-dation reactions.

Log

[Cs]

Time

Second Order

First OrderZero Order



The complex and heterogeneous nature of environmental solids makes it difficult,if not impossible, to identify specific sorption mechanisms for most solid-chemicalcombinations; in most situations, several mechanisms operate simultaneously. Inmost soils, and under most conditions, organic chemicals are sorbed on both organicand inorganic constituents. The relative importance of organic vs. inorganic constit-uent depends on the amount, distribution, and properties of those constituents andthe properties of the organic chemical. As the polarity, number of functional groups,and ionic nature of the organic chemical increases, so too does the number ofpotential sorption mechanisms (Figure 2.8). Fortunately, for many solid-organicchemical interactions, one or two mechanisms dominate the sorption process andgeneralizations regarding sorption behavior can be made.44

For instance, the sorption of most neutral, hydrophobic organic chemicals byenvironmental solids correlates highly with the organic matter content of the solid.The extent to which clay minerals contribute to sorption depends on both the ratioof clay mineral to organic carbon fractions of the soil or sediment and on the natureof the organic sorbate. A ratio of 40 has been suggested as the cutoff for organiccarbon dominated sorption.45 Among the various inorganic soil constituents, smec-tites have the greatest potential for sorption of organic chemicals, due to their largesurface area and abundance in agricultural soils.44,46

Soil is a dynamic and life-sustaining system composed of solids, liquid, and gas,with solids typically accounting for about one-half to two-thirds by volume. Living

Figure 2.7 The polarity of the H2O molecule. Because of the non-linear position of H+s,water is polar. The H2O molecule has one portion that is more negative thanpositive and an opposite side that has two hydrogens which are more positivethan negative.

Positiveside ofH Omolecule

105°

Hydrogen

Hydrogen

2

"+"

"+"

"-"

2

Negativeside ofH Omolecule

Oxygen



organisms are also very important parts of soil and contribute greatly to its generalproperties and behavior. The solid phase of soil comprises fragmented mineralmatter, derived from the weathering of hard rock at the earth’s surface, and fromsoil organic matter (SOM) consisting of a mixture of plant and animal residues invarious stages of decomposition and substances synthesized microbiologically.1

In its broadest sense, the term SOM encompasses all organic materials containedin soil and is made up of live organisms, their decomposed and partly decomposedremains, and microbially and/or chemically resynthesized products resistant to fur-ther biological attack. More specifically, the term SOM refers to the nonlivingorganic components, which are largely composed of products resulting from micro-bial and chemical transformations of organic debris. Some scientists have definedSOM as the total of organic components in soil, excluding undecayed plant and

Figure 2.8 Some examples of polar and nonpolar chemical species. Note that unbalancedelectrical charge, asymmetry, and the presence of oxygen all tend to makechemicals more polar.

NONPOLAR POLAR

Chloride ion

Tetrachloroethylene (PCE)

Benzene

H

H

H

HH

H

CC

Cl

ClCl

Cl Cl -

H

O

Water

H

d+

+HHydrogen ion

C

HH

HHHH

HH

CC

HH

HH HH

HH

C HH H OHC

HH

H

C C

H

H

H

C

Propane Propanol

H

H

H

H

Naphthalene

H

H

H

HAcetate Ion

CC

H

H

H

H

O

δ −

δ +



animal tissues, their “partial decomposition” products (the organic residues), the soilbiomass (living microbial tissues), and macrofauna and macroflora. The terms SOMand humus are thus generally interchangeable.1

To simplify this very complex system, SOM is generally divided into two groupsdesignated as nonhumic and humic substances.1 The nonhumic substance groupcomprises organic compounds that belong to chemically recognizable classes andare not unique to the soil. These include polysaccharides and simple carbohydrates,amino sugars, proteins and amino acids, fats and waxes, lignin, resins, pigments,nucleic acids, hormones, a variety of organic acids, and so on. Most of thesesubstances are relatively easily degradable and can be utilized as substrates by soilmicroorganisms, and as such have a transient existence in the soil. In contrast, humicsubstances comprise a heterogeneous mixture of chemically unidentifiablemacromolecules that are distinctive to and synthesized in the soil, and are relativelyresistant to chemical degradation and microbial attack.

Recent estimates of the average composition of SOM are the following: carbo-hydrates, 10%; N components, 10%; lipids (including alcanes, fatty acids, waxes,and resins), 15%; humic substances, 65%. However, different soils may containwidely different amounts of nonhumic and humic substances. The amount of car-bohydrates can range from 5 to 25%; proteins may vary from 15 to 45%; lipids from2% in forest SOM to 20% in acid peat soils, and humic substances from 33 to 75%of the total SOM.1

Humic substances are the most widespread and ubiquitous natural nonlivingorganic materials in all terrestrial and aquatic environments and represent a signif-icant proportion of total organic C in the global C cycle. They constitute the majorfraction of SOM (up to 80%) and the largest fraction of natural organic matter(NOM) in aquatic systems (up to 60% of dissolved organic C).1

Soil humic substances comprise a physically and chemically heterogeneousmixture of naturally occurring, biogenic, relatively high-molecular weight, yellow-to-black colored, amorphous, colloidal, polydispersed organic polyelectrolytes.These polyelectrolytes are of mixed aliphatic and aromatic natures, formed bysecondary synthesis reactions (humification) during the decay process and transfor-mation of organic matter originating from dead organisms and microbial activity.These materials are distinctive of the soil system and exclusive of undecayed plantand animal tissues, their partial decomposition products, and the soil biomass.

Soil water acts both as a solvent for the organic chemical and as a solute withwhich the organic chemical has to compete for sorption sites on the solid surface.Typically, soil water is a solution comprising mainly Ca+2, Mg+2, Na+, K+, SO4

–2 ,CO3

–2 , and HCO3– . Ionic strengths are typically 0.5 mol/L or higher; pH values of

5–8.5 are common.44 The characteristics of the solution phase determine the reactionchemistry and the dissolution/precipitation reactions, and they influence ion activity,ion pairing, and speciation. All these potentially can influence a chemical’s sorptivebehavior (Figures 2.9a, b, and 2.10).

In large lakes and estuaries, the natural organic material in sediments and sus-pended sediments is derived from a mixture of the remains of terrestrial and plank-tonic organisms. Generally, soils and sediments differ in the amount and type oforganic matter they contain. Soils typically contain higher percentages of cellulose



and hemicellulose, while sediments contain higher percentages of lipid-like material.For neutral organic compounds, sorption is generally greater in sediments than insoils, even when normalized to organic carbon content.

2.3.1.1 Soil Sorption Coefficients

Sorption coefficients quantitatively describe the extent to which an organic chem-ical distributes itself between an environmental solid (i.e., soil, sediment, suspendedsediment, wastewater solids, etc.) and the aqueous phase that it is in contact with atequilibrium. Sorption coefficients generally are determined from an isotherm, a dia-gram that depicts the distribution of the test chemical between a solid sorbent and thesolution in equilibrium with it over a range of concentrations at constant temperature(Figure 2.11). These isotherms can be linear or nonlinear, depending on the propertiesof the chemical and solid and on the aqueous phase concentration, but tend to becomenonlinear (sorption tends to decrease) as the concentration of chemical in the aqueousphase increases, especially for polar or ionizable chemicals or soils that are low inorganic carbon and high in clay. Linear sorption isotherms often are observed if theequilibrium aqueous phase organic compound concentrations are below 10–5 M or

Figures 2.9a A cross section of a soil pore and the solid particles that make up its walls.Water is held strongly as the distance from the soil particle decreases; at somedistances from the surface, water is held so weakly that the pull of gravity causessome of it to drain.

WATER

WATER

WATER

WATER

SOILPARTICLE

SOILPARTICLE SOIL

PARTICLE

SOILPARTICLE

Water is HeldStrongly nearthe Soil ParticleSurface

AirWhere

Water hasDrained

MatricPotentialIncreases

towardParticleSurface

Air WhereWater hasDrained



one-half the aqueous phase solubility (whichever is lower) and the organic content ofthe solid is greater than 0.1%:

Kd = CS/CW (2.16)

where, CS and CW are the concentrations of the organic chemical sorbed by the solidphase (mg/Kg) and dissolved in aqueous phase (mg/L), respectively. Units of Kd

typically are given as L/kg, mL/g, or cm3/g.For nonlinear isotherms, the Freundlich model most often is used to describe

the relationship between the sorbed (CS) and the solution phase concentrations (CW):

CS = Kf CWN (2.17)

where, Kf is the Freundlich sorption coefficient and N (values of N are less than oneand typically range between 0.75 and 0.95) generally is a constant.47 However, insome cases, N has been observed to exceed one. When N is equal to one, a linearequation results, and Kf and Kd are equivalent.

The Langmuir and Brumnauer, Emmett, and Teller (BET) models also have beenused to describe nonlinear sorption behavior for environmental solids, particularly

Figure 2.9b Trapped gas in saturated soil.

Trapped gas bubble

Solids

PoreSpace

Advection - dispersion

C = H'C

w

wg

Aqueous phaseconcentration

Gas phaseconcentration

Dimensionless Henry'sLaw Constant

C =

C =

H =

w

g



for mineral dominated sorption.

44

The Langmuir model assumes that maximumadsorption corresponds to a saturated monolayer of solute molecule on the adsorbentsurface, that there is no migration of the solute on the surface, and that the energyof adsorption is constant. The BET model is an extension of the Langmuir modelthat postulates multiplayer sorption. It assumes that the first layer is attracted moststrongly to the surface, while the second and subsequent layers are more weaklyheld.

47

The most commonly used method for expressing the distribution of an organiccompound between the aquifer matrix and the aqueous phase is the distributioncoefficient, K

d

, which is described by Equation 2.16.The transport and partitioning of a contaminant are strongly dependent on the

chemical’s soil-water distribution coefficient and water solubility. The distributioncoefficient is a measure of the sorption/desorption potential and characterizes thetendency of an organic compound to be sorbed to the aquifer matrix. The higher thedistribution coefficient, the greater the potential for sorption to the aquifer matrix.The distribution coefficient is the slope of the sorption isotherm at the contaminantconcentration of interest. The greater the amount of sorption, the greater the valueof K

d

. For systems described by a linear isotherm, K

d

is a constant.

47

In general

Figure 2.10

Soil water under three different values of water content. At high water content(condition C) porewater pressure is rendered negative by the force of surfacetension acting over a meniscus of relatively large area. The meniscus may bethought of as a flexible diaphragm that is under tension, thus pulling on the wateron its convex side. As water content is decreased and the meniscus retreatsinto smaller pore spaces (B, then A), surface tension forces act over a smallerarea of water, and the resulting water pressure is more negative. The sameeffect occurs in capillary tubes, where the most suction (more negative pressure,and thus more capillary rise) is developed in the tube of smallest diameter.

Meniscus

A B C

Meniscus

A B CB

CAir

Air

SoilGrain

A

B

C

0

Porewater Pressure- +

A and B

A only

A, B, C

Water Pressure

Hei

ght

0- +



terms the distribution coefficient is controlled by the hydrophobicity of the contam-inant and the total surface areas of the aquifer matrix available for sorption. Thusthe distribution coefficient for a single compound will vary with the composition ofthe aquifer matrix. Because of their extremely high specific surface areas (ratio ofsurface area to volume), the organic carbon and clay mineral fractions of the aquifermatrix generally represent the majority of sorption sites in an aquifer.

Based on literature reports, it appears that the primary adsorptive surface fororganic chemicals is the organic fraction of the aquifer matrix.47 However, there isa critical level of organic matter below which sorption onto mineral surfaces is thedominant sorption mechanism.47,48 This critical level of organic matter, below whichsorption appears to be dominated by mineral-solute interactions, and above whichsorption is dominated by organic carbon-solute interactions, is given by47,48

(2.18)

where

focc= critical level of organic matter (mass fraction)

As = surface area of mineralogical component of aquifer matrixKow = octanol-water partitioning coefficient

Figure 2.11 Characteristic adsorption isotherm shapes.

Langmuir

Freundlich

LinearA

dsor

bed

Con

cent

rat io

n C

s(µg

/g)

Dissolved Concentration Cw(µg/mL)

fA

Kocs

owc

=200

10 84.



From this relationship it is apparent that the total organic carbon content of theaquifer matrix is less important for solutes with low octanol-water partitioningcoefficients (Kow).47 Also apparent is the fact that the critical level of organic matterincreases as the surface area of the mineralogic fraction of the aquifer matrixincreases. The surface area of the mineralogic component of the aquifer matrix ismost strongly influenced by the amount of clay. For compounds with low Kow valuespresent in materials with a high clay content, sorption to mineral surfaces could bean important factor causing retardation of the chemical.

Several researchers have found that if the distribution coefficient is normalizedrelative to the aquifer matrix total organic carbon (TOC) content, much of thevariation in observed Kd values between different soils is eliminated.49 Distributioncoefficients normalized to total organic carbon content are expressed as Koc. Thefollowing equation gives the expression relating Kd to Koc:

(2.19)

where

Koc = soil sorption coefficient normalized for total organic carbon contentKd = distribution coefficientfoc = fraction of total organic carbon (mg organic carbon/mg soil)

In areas with high clay concentrations and low TOC concentrations, the clayminerals become the dominant sorption sites. Under these conditions, the use of Koc

to compute Kd might result in underestimating the importance of sorption in retar-dation calculations, a source of error that will make retardation calculations basedon the total organic carbon content of the aquifer matrix more conservative. In fact,aquifers that have a high enough hydraulic conductivity to spread organic chemicalcontamination generally have a low clay content. In these cases the contribution ofsorption to mineral surfaces is generally trivial.

Sorption coefficients also have been expressed on an organic matter basis (Kom)by assuming that the organic matter content of a soil or sediment equals some factor,usually between 1.7 to 1.9, times its organic carbon content on a mass basis.47,50

Often 1.724 is used as this factor, implying that the carbon content of organic matteris 1/1.724 or 60%. However, Koc is considered a more definite and less ambiguousmeasure than Kom.47

Assumptions inherent in the use of a Koc (or Kom) are that: sorption is exclusivelyto the organic component of the soil, all soil organic carbon has the same sorptioncapacity per unit mass, equilibrium is observed in the sorption–desorption process,and the sorption and desorption isotherms are identical.45 Both Koc and Kd have unitsof L/kg or cm3/g.

Numerous studies have been performed using the results of batch and columntests to determine if relationships exist that are capable of predicting the sorptioncharacteristics of a chemical based on easily measured parameters. The results of

KK

focd

oc

=



these studies indicate that the amount of sorption is strongly dependent on the amountof organic carbon present in the aquifer matrix and the degree of hydrophobicityexhibited by the contaminant.47 These researchers observed that the distributioncoefficient, Kd, was proportional to the organic carbon fraction of the aquifer timesa proportionality constant. This proportionality constant, Koc, is defined as given byEquation 2.19. Because it is normalized to organic carbon, values of Koc are depen-dent only on the properties of the compound (not on the type of soil). Values of Koc

have been determined for a wide range of chemicals.By knowing the value of Koc for a contaminant and the fraction of organic

carbon present in the aquifer, the distribution coefficient can be estimated usingthe relationship

Kd = Koc foc (2.20)

The fraction of soil organic carbon must be determined from site-specific data.Representative values of the fraction of organic carbon (foc) in common sediments isavailable in the literature. When predicting sorption of organic compounds, total organiccarbon concentrations obtained from the most transmissive aquifer zone unaffected bycontamination should be averaged and used for predictions. This is because the majorityof dissolved contaminant transport occurs in the most transmissive portions of theaquifer. In addition, because the most transmissive aquifer zones generally have thelowest total organic carbon concentrations, the use of this value will give a conservativeprediction of contaminant sorption and retardation. Determination of the coefficient ofretardation using sorption coefficients is described in Chapter 3.

2.3.1.2 Factors Affecting Sorption Coefficients

Many factors potentially can affect the distribution of a contaminant between anaqueous and solid phase. These include environmental variables, such as temperature,ionic strength, dissolved organic matter concentration, and the presence of colloidalmaterial, surfactants, and cosolvents. In addition, factors related specifically to theexperimental determination of sorption coefficients, such as sorbent and solid concen-trations, equilibration time, and phase separation technique, can also be important. Abrief discussion of several of the more important factors affecting sorption coefficientsfollows.

Temperature: The effect of temperature on sorption equilibrium is a directindication of the strength of the sorption process. The weaker the interaction betweensorbent and sorbate, the less the effect of temperature.47,50 While temperature caninfluence sorption, the strength and direction of the effect depends on the propertiesof the sorbent and sorbate and on the sorption mechanism. Adsorption processes aregenerally exothermic, so the higher the temperature, the less the adsorption. Hydro-phobic sorption, however, has been shown to be relatively independent of tempera-ture. Other reviews also indicate that the influence of temperature on equilibriumsorption and have found that, in most cases, equilibrium sorption decreases withincreasing temperature.47



pH: For neutral chemicals, sorption coefficients usually are unaffected by pH.However, for ionizable organic chemicals, sorption coefficients can be affectedgreatly, since pH affects not only the speciation but also the surface characteristicsof natural sorbents. Typically, for weak acids the free acid form (HA) is more stronglysorbed than the anionic form (A–). For example, pentacholorophenol (PCP) sorptiondecreased with increasing pH over the entire pH range tested (2 to 12). For weakbases the cationic form dominates at low pH and is more highly sorbed than thefree base.44

Ionic Strength: Salts can affect sorption of organic compounds by displacingcations from the soil ion exchange matrix, by changing the activity of the sorbatein solution, and by changing the charge density associated with the soil sorptionsurface. Salt effects are most important for basic sorbates in the cation state, wherean increase in salinity can significantly lower the sorption coefficient. Salt effectsare least important for neutral compounds, which may show either increases ordecreases in sorption as salinity increases.44

Dissolved or Colloidal Organic Matter: The presence of dissolved or colloidalorganic matter has been shown to influence sorption depending on the nature of thechemical and the organic matter. Some compounds were found to be associatedextensively with the dissolved organic matter; sorption by soil decreased significantlyin the presence of dissolved organic matter. Some have characterized several sizefractions of water soluble organic carbon and found that the effect of dissolvedorganic matter on the sorption of pyrene may be limited, but the presence of colloidalorganic matter suspended in the soil solution may have significant impact on thesorption of pyrene.44,51

Cosolvents: The effect of nonpolar cosolutes (trichloroethylene, toluene), polarcosolutes (1-octanol, chlorobenzene, nitrobenzene, o-cresol) and polar cosolvents(methanol and dimethyl sulfoxide) on sorption of several polycyclic aromatic hydro-carbons (PAHs) has been investigated.44,52 The nonpolar cosolutes did not signifi-cantly influence PAH sorption, while the polar cosolutes (nitrobenzene, o-cresol),having sufficiently high aqueous solubilities, caused a significant decrease in PAHsorption.

Miscible organic solvents, such as methanol and ethanol, have been shown toincrease solubility of hydrophobic organics and to decrease sorption. This is pre-sumably the result of reducing the activity coefficient of the sorbate chemical in theaqueous phase, and competition for sorbing sites.

Competitive Sorption: At concentrations normally encountered in environmen-tal situations, sorption often has been observed to be relatively noncompetitive. Forexample, it was found that there is no competition in the sorption of binary solutesm-dichlorobenzene and 1,2,4-trichlorobenzene and between parthion and lindane.53

The sorption of methyl and dimethyl naphthalene, individually and as componentsof JP-8 and synthetic jet fuel mixture, on two sediments and montomorillonite clayin water was measured.54 The sorption coefficients of the naphthalenes generallyvaried by less than a factor of two. However, there are reports of competitive sorptiontaking place that is thought to be the result of site-specific sorption occurring in soilorganic matter.



Organic Matter Type and Origin: While the constancy of Koc values suggestsa uniformity of organic matter with regard to sorption behavior, it is becomingincreasingly apparent that organic matter type can be an important sorption variablefor some sorbent/sorbate combinations. For example, it was found that the sorptionof naproamide, a nonionic herbicide, was greater in the sediment than in soils, evenon an organic carbon basis.44 The increased sorption in sediment was attributed tothe fact that soils contained a higher percentage of cellulose and hemicellulosematerial, whereas the sediments contain a higher lipid-like fraction.

Kinetic Considerations: Sorption generally is regarded as a rapid process and, inmany laboratory sorption experiments, equilibrium often is observed within severalminutes or hours. An equilibration time of 24 hours often is used for convenience.True sorption equilibrium under natural conditions, however, may require weeks tomonths to achieve depending on the chemical and environmental solid of interest. Inmany instances, an early period of rapid and extensive sorption, followed by a longslow period, is observed. Experimental determination of sorption coefficients requirespreliminary kinetic experiments to determine the time to reach equilibrium.

Two processes govern rate-limited or nonequilibrium sorption: transport ofthe substance to the sorption sites and the sorption process itself.44,50 Transport-related nonequilibrium typically results from the existence of a heterogeneousflow domain. Sorption-related nonequilibrium, caused by rate-limited interactionsbetween the sorbate and sorbent, may be the result of chemical nonequilibrium(i.e., chemisorption) or diffusive mass transfer limitations (i.e., diffusion of solutewithin pores of microporous particles or molecular diffusion into macromolecularorganic matter). Sorption kinetics are likely to be environmentally important inshort contact situations such as sediment resuspension, soil erosion, and infiltrat-ing ground water.44

In general, adsorption processes tend to be rapid and nearly instantaneous,whereas nonsurface sorption tends to be slower. For neutral organic chemicals, themore hydrophobic the compound, the larger the sorption coefficient, and the longerit takes to reach equilibrium between the solid and aqueous phases. This is becausethe sorbent must remove a chemical from a larger volume of water.

Generally, sorption estimates are based on equilibrium conditions only; however,incorporation of kinetic considerations into sorption estimation techniques is likelyto be an important area of future work. For example, the assumption of equilibriumsorption in dynamic field systems may result in calculating too much pesticide inthe sorbed state.

Ionizability: For neutral organic compounds, in soils having a low clay/organiccarbon ratio, sorption coefficients tend to increase as the hydrophobicity of thecompound increases. Aqueous solubility or octanol/water partition coefficients oftenare used as indicators of a compound’s hydrophobicity. An increase on polarity,number of functional groups, and ionic nature of the chemical will increase thenumber of potential sorption mechanisms for a given chemical. For ionizable com-pounds, pKa is of particular importance because it determines the dominant form ofa chemical at the specific environmental pH.



The entropy change is largely due to the destruction of the highly structuredwater shell surrounding the solvated organic. The term “partitioning” was used todenote an uptake in which the sorbed organic chemical permeates the network ofan organic medium by forces common to the solution, analogous to the extractionof an organic compound from water with an organic liquid. By either description,hydrophobic sorption or partitioning should increase as compounds become lesswater soluble or more hydrophobic.

Additional characteristics typically associated with hydrophobic sorption or par-titioning include sorption isotherms that are linear over a relatively wide range ofconcentrations, and sorption coefficients that are not strongly temperature dependent,and lack a competition between sorbates.44,53

2.3.2 Oxidation-Reduction Capacities of Aquifer Solids

There has been considerable research activity focused on the characterization ofREDOX-potential or intensity (Eh) conditions in groundwater systems defined asthe REDOX activity of dissolved chemical species. Early observations of significantEh trends along groundwater flow paths led to hypotheses of successive REDOXzones characterized by the activity of specific thermodynamically favored electronacceptors. These REDOX zones may be classified as oxic (i.e., detectable dissolvedO2), suboxic or postoxic (i.e., no detectable O2 or sulfide, detectable Fe2+), andreducing (i.e., detectable Fe2+ and sulfide, no detectable O2).1 Further investigationscorrectly postulated that oxidation-reduction processes were mediated by naturalmicrobial populations that catalyze electron-transfer reactions. More recent worknoted considerable temporal and spatial variability in measured subsurface REDOXconditions and that the succession of electron acceptors under oxic, suboxic, orreducing conditions was not strictly predictable by either chemical equilibriumcalculations or platinum electrode measurements.

2.3.2.1 pe and pH

A pH is the negative log (p for power) of proton (H+) activity and pe, its energyor work analog, is the negative log of the electron potential. An electron is not afull-fledged analog of a proton. Together, two equal but opposite charges make upa hydrogen atom, but that is about the extent of the equality between an electronand a proton. Without its proton, an electron is no longer an analog of H+, and it nolonger has any claim to being part of a hydrogen atom. An electron does not bounceabout by itself in the manner of an H+, and therefore it is probably not correct totry to characterize its “activity.” It always is either attached to an atom or radical orin the process of being transferred from one to another.

A proton is a cation. It can replace or be replaced by other cations and it is asgood as any other cation when it comes to balancing a chemical equation. Electronsreceive no recognition in balanced chemical equations because the donated andaccepted electrons must always cancel one another on opposite sides of an equation.



Electrons do not have anion status. They cannot trade places with other negativelycharged species.

Usually we see release of H+ when metals are oxidized and consumption of H+

with their reduction. Oxidation is furthered in a subsurface environment whereprotons and electrons are deficient; that is, where acidity and levels of easily degraded(labile) electron donors are low. But there must be a ready supply of availableelectron acceptors. Reduction is favored by surpluses of both protons and electrons.This means that low pH and high availability of organic substances will promotereduction in soil. Reduction of Fe or Mn oxides, or of nitrate, uses up H+, therebyincreasing pH of the soil and, theoretically, lowering the pe. Oxidation of Fe, Mn,or nitrate lowers the pH (measurable) and raises the pe (not measurable in mostsoils). Measuring changes in concentrations of REDOX species is more reliable forpredicting these things in the subsurface than is an attempted measurement of pewith the platinum electrode.

The farther apart the electrons, the more proportional work required to bringthem together and the higher the respective pe. A low pe system has a surplus ofelectrons and, therefore, a big tendency to lose some of them and become oxidized.A high pe system is hungry for electrons. As deficient electrons are replenished, thetendency for reduction to occur will increase. If we substitute pe and pH for theirdefined equivalents in a generic REDOX half-reaction in which activities of oxidizedand reduced species are equal, we see that the (pe + pH) sum is equivalent to theequilibrium constant of the half-reaction:

Oxidized species + e– + H+ = reduced species (2.21)

log K = log red – log ox – log e– – log H+ (2.22)

log K = pe = pH (2.23)

If indeed their sum is constant, then, thermodynamically, pe and pH are onopposite ends of a seesaw. If behavior follows thermodynamic theory, when onegoes up, the other will come down, like any sound seesaw. This sum is referred toas the REDOX parameter because, if a soil is at internal equilibrium, the (pe + pH)represents the sums of all of the REDOX equilibrium constants in the soil.1

2.3.2.2 REDOX Poise

In the natural environment REDOX seesaws are not so simple. This seesaw-likebehavior reflects the interaction between source/sink quantities and electron/protonintensities. If we add reducing reagents or reduced substances such as Fe(II), orMn(II) or Cr(III) to a soil poised so that its easily reduced substances are in balancewith its easily oxidized substances, some of the added reduced species will be quicklyoxidized. On the other hand, adding Fe(III), Mn(IV) or Cr(VI) will result in imme-diate reduction of a portion of the added oxidants. There appears to be a tendencyfor a soil, if disturbed, to maintain a REDOX balance, that is, poise, by donating



electrons to surplus electron acceptors or by accepting electrons from surplus elec-tron donors.1

A soil kept near field capacity moisture with occasional mixing, double baggedinside a thin polyethylene bag for several months at 15 to 25ºC, will be close tointernal equilibrium. If this metastable equilibrium is disturbed by adding an easilyoxidized substance to it, e.g., glucose, the (pe +pH) of the overall system will tendto remain fairly constant as the disturbed soil system moves back toward a newmetastable equilibrium. In this instance, the pe will tend to go down, and to theextent that it does, the pH will tend to rise.1

By adding increments of Cr3+ and HCrO4– , respectively, to separate subsamples

of the same soil and then determining the amount of Cr reduced [loss of Cr(VI)]and the amount oxidized [gain of Cr(VI)], it is possible to find a point of poise orbuffered REDOX region, where the electron donating and electron accepting tenden-cies cross. There the REDOX seesaw is balanced at dead-level.1

2.3.2.3 REDOX Reactions

REDOX is one of those catchy phrases invented by someone unhampered bycommitment to the use of scientifically correct terminology. The name is reversed(RED-OX, instead of OX-RED) for the sake of easy pronunciation. The RED standsfor reduction and it signifies gain of electrons by a chemical species called electronacceptors; the OX connotes oxidation, or electron loss by a chemical species calledelectron donors. Oxidation-reduction (REDOX) reactions, along with hydrolysis andacid-base reactions, account for the vast majority of chemical reactions that occurin aquatic environmental systems (soils, sediments, aquifers, rivers, lakes, and manyremediation operations). This section provides a survey of the environmental andsubstrate characteristics that govern REDOX transformations in aquatic systems.

The distinction between biotic and abiotic processes is a particularly importantissue in defining the scope of this section. Living organisms are responsible forcreating the conditions that determine the REDOX chemistry of most aquatic envi-ronmental systems. So, in this sense, most REDOX reactions in natural systemsultimately are driven by biological activity. Once environmental conditions areestablished, however, many important REDOX reactions proceed without furthermediation by organisms. These reactions are considered to be abiotic when it is nolonger practical (or possible) to link them to any particular biological activity.

Assigning Oxidation States: REDOX reactions involve oxidation and reduc-tion; they occur by the exchange of electrons between reacting chemical spe-cies.2,55 Electrons (or electron density) are lost (or donated) in oxidation andgained (or accepted) in reduction. An oxidizing agent (or oxidant) that acceptselectrons (and is thereby reduced) causes oxidation of a species. Similarly, reduc-tion results from reaction with a reducing agent (or reductant) that donateselectrons (and is oxidized).

To interpret REDOX reactions in terms of electron exchange, one must accountfor electrons in the various reacting species. Various textbooks provide simple rules,such as the following, for assigning oxidation states for inorganic REDOX couples:2,55



• For free elements, each atom is assigned oxidation number 0.• Monoatomic ions have an oxidation number equal to the charge of the ion.• Oxygen, in most compounds, has the oxidation number –2.• Hydrogen, in most compounds, has the oxidation number +1.• Halogens, in most environmentally relevant compounds, have the oxidation

number –1.

These rules, however, are not easily applied to organic REDOX reactions, andthis difficulty has led to a steady stream of alternative concepts for assigning oxi-dation states. For present purposes, familiarity with a method for assigning oxidationstates to organic molecules is sufficient. This method reflects the qualitative obser-vations from which the historical concepts of oxidation and reduction originated:oxidation is the gain of oxygen (O), chlorine (Cl) or double bonds, and/or the lossof H; reduction is the gain of H, saturation of double bonds, and/or loss of O or Cl.Thus, for example, mineralization of any hydrocarbon to CO2 and H2O involvesoxidation, and dechlorination of any chlorinated compound to hydrocarbon productsinvolves reduction.

Oxidations: Organic chemicals that are susceptible to oxidation and are ofconcern from the perspective of contamination and environmental degradationinclude aliphatic and aromatic hydrocarbons, alcohols, aldehydes, and ketones, phe-nols, polyphenols, sulfides (thiols), sulfoxides, nitriles, amines, diamines, nitrogenand sulfur hetercyclic compounds, mono- and di-chlorinated aliphatics and manyothers. Equations below show example half-reactions for oxidation of some of thesechemical groups.

Alkanes to alcohols

R – H + H2O → R – OH + 2H+ + 2e– (loss of H+ and e–) (2.24)

Alcohols to aldehydes

R CH2 OH → RCHO + 2H+ + 2e– (loss of H+ and e–) (2.25)

Aldehydes to acids

RCHO + H2O → RCOOH + 2H+ + 2e– (loss of H+ and e–) (2.26)

Reductions: Most interest in reductive transformations of environmental chem-icals involves dechlorination of chlorinated aliphatic and aromatic compounds andthe reduction of nitroaromatic compounds. Other examples of reductive transforma-tions that may occur abiotically in the environment include reduction of azo com-pounds, quinines, disulfides, and sulfoxides. An example of a half-reaction isdescribed by the equation:



Reductive dechlorination

R – Cl + H+ + 2e– → R – H + Cl– (gain of H+, e– and loss of Cl–) (2.27)

Dechlorination can occur by several reductive pathways. The simplest results inreplacement of a C-bonded halogen atom with a hydrogen and is known as hydro-genolysis or reductive dechlorination. The process is illustrated for trichloroethene,TCE, in Figure 2.12, where complete dechlorination by this pathway requires multiplehydrogenolysis steps. The relative rate of each step is a critical concern because thesteps tend to become slower with each dechlorination (and DCE and VC are at leastas hazardous as TCE if not more so than with VC). Aryl halogens, such as those inthe pesticide chlophyrifos, also are subject to hydrogenolysis, but this reaction rarelyoccurs abiotically. One notable exception is the rapid abiotic dechlorination of poly-chlorinated biphenyls (PCBs) by zero-valent iron with catalysis by Pd.2,55

The other major dechlorination pathway involves elimination of two chlorines,leaving behind a pair of electrons that usually goes to form a carbon-carbon doublebond. Where the pathway involves halogens on adjacent carbons, it is known asvicinal dehalogenation or reductive β-elimination. The major pathway for reductivetransformation of lindane involves vicinal dehalogenation, which can proceed bysteps all the way to benzene (Figure 2.13).2,55 Recently, data have shown that thispathway not only can convert alkanes to alkenes, but also can produce alkynes fromdihaloalkenes (see Equation 2.28).

Figure 2.12 Reductive dechlorination or hydrogenolysis of TCE.

Figure 2.13 Vicinal dechlorination or reductive-elimination of lindane.

TCE

+H+ +2e-

-CI-

CI

CI

H

CI

C C

cis-1,2-DCE

+H+ +2e-

-CI-

H

CI

H

CI

C C

VC

+H+ +2e-

-CI-

CI

H

H

H

C C

Ethene

H

H

H

H

C C

CI

CI

CI

CI

CI CI

Lindane

CI

CI

CI

CI

CI+2e-

-2CI-+2e-

-2CI-CI

Benzene

+2e-

-2CI-



Vicinal Dehalogenation

Cl – R – R1 – Cl + 2e– → R = R1 + 2Cl– (formation of double bond) (2.28)

The contaminant REDOX reactions just summarized only occur when coupledwith suitable half-reactions involving oxidants or reductants from the environment.In a particular environmental system, these REDOX agents collectively determinethe nature, rate, and extent of contaminant transformation. Under favorable circum-stances, the dominant REDOX agent(s) can be identified and quantified, therebyproviding a rigorous basis for estimating the potential for, and rate of, transformationby abiotic REDOX reactions.2,55

Such specificity is often possible with systems engineered for contaminant reme-diation. However, natural systems frequently involve complex mixtures of REDOX-active substances that cannot be characterized readily. The characterization ofREDOX conditions in complex environmental media is a long-standing challengeto environmental scientists that continues to be an active area of research.

The remainder of this section summarizes what is currently known about theidentity of oxidants and reductants relevant to environmental systems, in order toprovide a basis for estimating rates of contaminant transformations by specificpathways. With respect to natural reductants, however, a great deal remains to belearned, so substantial developments can be expected as new research in this areabecomes available.

Oxidants: The best opportunities for predicting REDOX transformations comefrom engineered systems where a known oxidant is added to achieve contaminantremediation. Well-documented examples include the use of ozone and chlorine indrinking water treatment. In natural systems, important oxidants are oxides of ironand manganese, as well as molecular oxygen and various photooxidants. In engi-neered remediation systems oxidants used include potassium permanganate, ozoneand hydrogen peroxide.1

The presence of molecular oxygen, O2 is used widely as the defining character-istic of oxidizing environments because the overwhelming supply of molecularoxygen makes it the ultimate source of oxidizing equivalents. However, O2 in itsthermodynamic ground-state (3O2) is a rather poor oxidizing agent and it is notusually the oxidant directly responsible for oxidative transformations of contami-nants. Instead, activated oxygen species may be involved where they are formed bythe action of light on natural organic matter (NOM), peroxides, or various inorganiccatalysts. Activated oxygen species include singlet oxygen (1O2), protonated super-oxide (HO2

· ) hydrogen peroxide and hydroperoxide anion (H2O2/HO2– ), hydroxyl

radical (OH•), and ozone (O3).1,2,55

O2 + e– + H+ = H2O˚ (protonated superoxide) (2.29)

O2 + 2e– + 2H+ = H2O2 (hydrogen peroxide) (2.30)

O2 + 3e– + 3H+ = H2O + HO˚ (hydroxyl free radical) (2.31)



O2 + 4e– + 4H+ = 2H2O (water) (2.32)

Equations 2.29–2.32 are half-reactions showing reduction of O2 by single elec-tron additions. Thus, superoxide and hydroxyl, produced by one and three oddelectron additions, are free radicals; whereas peroxide and water, with two and fourelectrons added, respectively, are not. Restricting conditions of interaction betweenthe availabilities of soil O2 and electron donors, for example, at the interface betweenoxygenated water and anaerobic soil in a wetland, tends to favor transfers of electronsin single steps, and thus such interfaces are likely to be sites for free radicalformation. Free radical mechanisms appear to explain why kinetically slow andseemingly unlikely REDOX transformations often occur readily at interfaces. Oxy-gen free radicals are much more reactive than O2 itself, and both superoxide andthe hydroxyl free radical are especially reactive with H2O2, each one capable ofbeing quickly transformed into the other.

Aside from oxygen and the activated oxygen species, there are several otheroxidants that cause abiotic oxidation reactions involving environmental contami-nants. In engineered systems, these include chlorine, chlorine dioxide, permanganateand ferrate. At highly contaminated sites, anthropogenic oxidants such as chromate,arsenate, and selenate may react with co-contaminants such as phenols.

In natural anoxic environments, the major alternative oxidants are Fe(III) andmanganese (IV) oxides and hydroxides. Both are common in natural systems ascrystalline or amorphous particles or coatings on other particles. In the absence ofphotocatalysis, however, iron and manganese oxides are weak oxidants. As a result,they appear to react at significant rates only with phenols and anilines.

In the dissolved phase, few alternative abiotic oxidants are available in the neutralenvironment. Nitrate, sulfate, and other terminal electron acceptors used by anaerobicmicroorganisms are thermodynamically capable of oxidizing some organic contami-nants, but it appears that these reactions almost always require microbial mediation.

Reductants: Abiotic environmental reductants are not well characterized as theoxidants because, until recently, there were fewer remediation applications of reduc-tants, and natural reducing environments are characterized by especially complexbiogeochemistry. The most familiar natural reductants are sulfide (present primarilyas HS– and H2S), Fe (II) and Mn (II), and natural organic matter (NOM). Thetransformation of contaminants by sulfur species in anaerobic environments caninvolve both reduction and nucleophilic substitution pathways. These processes havebeen studied extensively, but the complex speciation of sulfur makes routine pre-dictions regarding these reactions difficult.1,2,55

A similar situation applies for reduced forms of iron. As with oxidations, someof the best opportunities for reliably estimating rates of redox transformations areafforded by engineered systems where a reductant of known composition and quan-tity is added to achieve contaminant remediation. In addition to zero-valent iron,other methods for chemical reduction of contaminants involve dithionite and elec-trolysis (where, in effect, electrons are added directly).1,2,55

The role of natural organic reductants in environmental systems is even moredifficult to characterize than the roles of sulfur and iron because most natural organicmatter is of indeterminate composition. There are two general categories of NOM:



high molecular weight organic materials such as humic and fulvic acid, and lowmolecular weight compounds such as acids, alcohols, etc. Specific examples of thelatter include glycolate, citrate, pyruvate, oxalate, and ascorbate. These types ofcompounds have been studied extensively for their role in global cycling of carbon,but very little work has been done on whether they act as specific reductants oforganic contaminants.1,2,55

In contrast, the possibility that high molecular weight NOM acts as a reductantin environmental systems is widely acknowledged. Although most evidence for thisinvolves the reduction of metal ions, several studies have shown that the processextends to various organic contaminants. Presumably, the reducing potential of NOMis due to specific moieties such as complex metals or conjugated polyphenols. Often,REDOX reactions involving these moieties are reversible, which means that NOMmay serve as a mediator of REDOX reactions rather than being just an electrondonor (or acceptor).1,2,55

In the recent past, the addition of labile electron donors such as molasses, lactate,and methanol is gaining ground to facilitate enhanced reductive dechlorination ofchlorinated aliphatic and aromatic compounds. This technology is discussed in detailin Chapter 4.

Demonstrating that a REDOX transformation of a contaminant involves mediatedelectron transfer requires meeting several criteria: 1) the overall reaction must beenergetically favorable, 2) the mediator must have a reduction potential that liesbetween the bulk donor and the terminal acceptor so that both steps in the electrontransfer chain will be energetically favorable, and 3) both steps in the mediatedreaction must be kinetically fast relative to the direct reaction between bulk donorand terminal acceptor. Most evidence for involvement of mediators in reduction ofcontaminants comes from studies with model systems, because natural reducingmedia (such as anaerobic sediments) consist of more REDOX couples than can becharacterized readily. Although this is an active area of research, a variety of likelymediator half-reactions can be identified.

REFERENCES

1. Sparks, D. L., Soil Physical Chemistry, CRC Press, Boca Raton, FL, 1998.2. Boethling, R. S. and D. MacKay, Handbook of Property Estimation Methods for

Chemicals, Lewis Publishers, Boca Raton, FL, 2000.3. MacKay, D., W. Y. Shiu, and K. C. Ma, Henry Law Constant, in Handbook of Property

Estimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., LewisPublishers, Boca Raton, FL, 2000.

4. Leo, A. et al., Partition coefficients and their uses, Chem. Rev., 71, 525–616, 1971.5. Leo, A. J., Hydrophobicity, the underlying property in most biochemical events,

Environmental Health Chemistry, McKinney, J., Ed., Ann Arbor Science, Ann Arbor,MI, 1981, 323–336.

6. Kenage, E., Determination of bioconcentration potential, Residue Rev., 44, 73–113,1996.

7. Neely, W. B. et al., Partition coefficients to measure bioaccumulation potential oforganic chemical in fish, Environ. Sci. Technol., 8, 1113–1115, 1974.



8. Lyman, W. J., W. F. Reehl, and D. H. Rosenblatt, Handbook of Chemical PropertyEstimation Methods, McGraw-Hill, New York, 1982.

9. Wolfe, N. L., and P. M. Jeffers, Hydrolysis, in Handbook of Property EstimationMethods for Chemicals, Boethling, R.S. and D. MacKay, Eds., Lewis Publishers,Boca Raton, FL, 2000.

10. Wolfe, N. L., Organophosphate and organophosphorothioate esters: application oflinear free energy relationships to estimate hydrolysis rate constants for use in envi-ronmental fate assessment, Chemosphere, 9, 571–579, 1980.

11. Mabey, W. R. and T. Mill, Critical review of hydrolysis of organic compounds inwater under environmental conditions, J. Phys. Chem. Ref. Data, 7, 383–415, 1978.

12. Jeffers, P. M. et al., Homogeneous hydrolysis rate constants for selected methanes,ethanes, ethenes and propanes, Environ. Sci. Technol., 23, 965–969, 1989.

13. Mill, T., Photoreactions in surface waters, in Handbook of Property EstimationMethods for Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers,Boca Raton, FL, 2000.

14. Larson, R. A., L. L. Hunt, and D. W. Blankenship, Formation of toxic products froma No. 2 fuel oil by photooxidation, Environ. Sci. Technol., 11, 492–496, 1977.

15. Atkinson, R. J., A structure-activity relationship for the estimation of rate constantsfor the gas phase reactions of OH radicals with organic compounds, Int. J. Chem.Kinetics, 19, 799–828, 1987.

16. Hoag, W. R. and T. Mill, Survey of sunlight-produced transient reactants in surfacewaters, Proceedings of a workshop on effects of solar ultraviolet radiatiaon ongeochemical dynamics, Woods Hole, MA, 1989.

16a. Atkinson, R., Atmospheric Oxidation, in Handbook of Property Estimation Methodsfor Chemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton,FL, 2000.

17. Mopper, K. and X. Zhou, Hydroxyl radical photoproduction in the sea and its potentialimpact on marine processes, Science, 250, 661–664, 1990.

18. Howard, P. H., Biodegradation, in Handbook of Property Estimation Methods forChemicals, Boethling, R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton,FL, 2000.

19. Alexander, M., Biodegradation and Bioremediation, Academic Press, New York,1999.

20. Spain, J. C. and P. A. Van Weld, Adaptation of natural microbial communities todegradation of xenobiotic compounds: effects of concentration, exposure time, inoc-ulum, and chemical structure, Appl. Environ. Microbiol., 45, 428–435, 1983.

21. Howard, P. H. and S. Banerjee, Interpreting results from biodegradability test ofchemicals in water and soil, Environ. Toxicol. Chem., 3, 551–562, 1984.

22. Alexander, M., Biodegradation of organic chemicals, Environ. Sci. Technol., 19,106–111, 1985.

23. Taylor, B. F. et al., Arch. Microbio., 122, 301–306, 1979.24. Oldenhuis, R. et al., Appl. Environ. Microbiol., 55, 2816–2819, 1989.25. Nelson, M. J. K. et al., Appl. Environ. Microbiol., 54, 604–606, 1988.26. Li, S. and L. P. Wackett, Biochem. Biophy. Res. Commun., 185, 443–451, 1992.27. Rebertson, J. B. et al., J. Appl. Environ. Microbiol., 58, 2643–2648, 1992.28. Delgado, A. et al., J. Appl. Environ. Microbiol., 58, 415–417, 1992.29. Shields, M. S. et al., J. Appl. Environ Microbiol., 57, 1935–1941, 1991.30. Wackett, L. P. et al., J. Appl. Environ Microbiol., 55, 2960–2964, 1989.31. Hyman, M. R. et al., J. Appl. Environ Microbiol., 60, 3033–3035, 1994.



32. Van Beilen, J. B., J. Kingma, and B. Witholt, Eng. Microb. Technol., 16, 904–911,1994.

33. Lee, K. and D. T. Gibson, J. Appl. Environ. Microbiol., 62, 3101–3106, 1996.34. Hernandez, B. S., J. J. Arensdorf, and D. D. Focht, Biodegradation, 6, 75–82, 1995.35. Ladd, T. I. et al., Heterotropic activity and biodegradation of labile and refractory

compounds in groundwater and stream microbial population, Appl. Environ. Micro-biol., 44, 321–329, 1982.

36. Neilson, A. H., Organic Chemicals, Lewis Publishers, Boca Raton, FL, 1999.37. Alexander, M., Biodegradataion of chemicals of environmental concern, Science, 211,

132–138, 1981.38. Klopman, G. et al., Computer-automated predictions of aerobic biodegradation of

chemicals, Environ. Toxicol. Chem., 14, 395–403, 1995.39. Punch, W. F. et al., Bess, a computerized system for predicting the biodegradation

potential of new and existing chemicals, 7th Int. Workshop on QSARS in Env. Sci.,June 24-28, Elsinore, Denmark, 1996.

40. Alexander, M., Nonbiodegradable and other recalcitrant molecules, Biotechnol.Bioeng., 15, 611–647, 1973.

41. Howard, P. H. et al., Review and Evaluation of Available Techniques for DeterminingPersistence and Routes of Degradation of Chemical Substances in the Environment,EPA-560/5-75-006, U.S. NTIS PB 243825, 1975.

42. Simkins, S. and M. Alexander, Models for mineralization kinetics with the variablesof substrate concentration and population density, Appl. Environ. Microbiol., 47,1299–1306, 1984.

43. Schmidt, S. K., S. Simkins, and M. Alexander, Models for the kinetics of biodegra-dation of organic compounds not supporting growth, Appl. Environ. Microbiol., 50,323–331, 1985.

44. Doucette, W. J., Soil and Sediment Sorption Coefficients, in Handbook of PropertyEstimation Methods for Chemicals, Boethling, R. S. and D. MacKay, Eds., LewisPublishers, Boca Raton, FL, 2000.

45. Green, R. E. and S. W. Karickoff, Sorption estimates for modeling, in Pesticides inthe Soil Environment, Cheng, H. H., Ed., Soil Science Society of America, Inc.,Madison, WI, 79–101, 1990.

46. Laird, D. A. et al., Adsorption of atrazine on smectites, Soil Sci. Soc. Amer. J., 56(1), 62–67, 1992.

47. Wiedemeier T. H. et al., Natural Attenuation of Fuels and Chlorinated Solvents inthe Subsurface, John Wiley & Sons, New York, 1999.

48. McCarty, P. L., M. Reinhard, and B. E. Rittmann, Trace organics in groundwater,Environ. Sci. Techn., 15, 40–51, 1981.

49. Dragun, J., The Soil Chemistry of Hazardous Materials, Hazardous Materials ControlResearch Institute, Silver Spring, MD, 1988.

50. Hamaker, J. W. and J. M. Thompson, Adsorption in Organic Chemicals in the SoilEnvironment, Goring, C. A. I. and J. W. Hamaker, Eds., Marcel Dekker, New York,1972, 49–143.

51. Herbert, B. E. et al., Pyrene sorption by water-soluble organic carbon, Environ. Sci.Technol., 27 (2), 398–403, 1993.

52. Rao, P. S. C., L. S. Lee, and R. Pinal, Consolvency and sorption of hydrophobicorganic chemicals, Environ. Sci. Technol., 24 (5), 647–654, 1990.

53. Chiou, C. T. and T. D. Shoup, Soil sorption of organic vapors and effects of humidityon sorption mechanism and capacity, Environ. Sci. Technol., 19, 1196–1200, 1985.



54. MacIntyre, W. G., T. B. Stauffer, and C. P. Antworth, A comparison of sorptioncoefficients determined by batch, column, and box methods on a low organic carbonacquifer material, Ground Water, 29 (6), 908–913, 1991.

55. Tratnyek, P. G. and D. L. Macalady, Oxidation-reduction reactions in the aquaticenvironment, Handbook of Property Estimation Methods for Chemicals, Boethling,R. S. and D. MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000.



63

CHAPTER

3


CONTENTS

3.1 Introduction ....................................................................................................643.1.1 Definitions of Natural Attenuation ....................................................64

3.2 Approaches for Evaluating Natural Attenuation ...........................................653.3 Patterns vs. Protocols .....................................................................................70

3.3.1 Protocols for Natural Attenuation......................................................703.3.2 Patterns of Natural Attenuation .........................................................71

3.3.2.1 Various Patterns of Natural Attenuation.............................723.4 Processes Affecting Natural Attenuation of Compounds..............................79

3.4.1 Movement of Contaminants in the Subsurface .................................793.4.1.1 Dilution (Recharge) ............................................................793.4.1.2 Advection ............................................................................813.4.1.3 Dispersion ...........................................................................83

3.4.2 Phase Transfers ..................................................................................853.4.2.1 Sorption...............................................................................853.4.2.2 Stabilization ........................................................................883.4.2.3 Volatilization .......................................................................89

3.4.3 Transformation Mechanisms..............................................................893.4.3.1 Biodegradation ....................................................................90

3.5 Monitoring and Sampling of Natural Attenuation ......................................1093.5.1 Dissolved Oxygen (DO) ..................................................................1133.5.2 Oxidation–Reduction (REDOX) Potential (ORP)...........................1173.5.3 pH .....................................................................................................1193.5.4 Filtered vs. Unfiltered Samples for Metals .....................................120

3.5.4.1 Field Filtration and the Nature of Groundwater Particulates..................................................121

3.5.4.2 Reasons for Field Filtration..............................................1223.5.5 Low-Flow Sampling as a Paradigm for Filtration ..........................1243.5.6 A Comparison Study........................................................................125

References..............................................................................................................126

L1282_C03-A_frame Page 63 Monday, June 18, 2001 10:05 AM


…natural attenuation (NA) is not a “no action (NA)” alternative. Monitorednatural Attenuation (MNA) defines the required monitoring parameters to dem-onstrate that the ongoing natural processes will continue to meet the remediationobjectives…

3.1 INTRODUCTION

The term monitored natural attenuation (MNA) refers to an approach to clean upsubsurface contamination, specifically in groundwater, by relying on natural processesand monitoring. MNA is also referred to as

natural degradation

and

intrinsic

or

passiveremediation

. Natural attenuation processes include a variety of physical, chemical, orbiological processes that, under favorable conditions, act without human interventionto reduce the mass, toxicity, mobility, volume, and concentration of contaminants ingroundwater. Depending on the geologic conditions, types of contaminants, and con-taminant mass and distribution at a given contaminated site, MNA could emerge asthe preferred choice of remediation approach. Natural attenuation relies on the

assim-ilative capacity

of the ecosystem for the reduction of contaminant concentration andmass. This approach has been utilized by environmental engineers for a long time tocontrol industrial and municipal wastewater discharges into surface waterbodies andmaintain acceptable water quality standards.

3.1.1 Definitions of Natural Attenuation

A variety of organizations have espoused the following definitions of naturalattenuation due to the emerging popularity and preference of MNA as the remedi-ation method of choice at many contaminated sites across the country.

1

Environmental Protection Agency

2

:

This policy directive defines monitorednatural attenuation as the reliance on natural attenuation process (within thecontext of a carefully controlled and monitored site cleanup approach) to achievesite-specific remediation objectives within a time frame that is reasonable com-pared to that offered by other more active methods. The “natural attenuationprocesses” that are at work in such a remediation approach include a variety ofphysical, chemical, or biological processes that, under favorable conditions, actwithout human intervention to reduce the mass, toxicity, mobility, volume, orconcentration of contaminants in soil or groundwater. These

in situ

processesinclude biodegradation; dispersion; dilution; sorption; volatilization; radioactivedecay; and chemical or biological stabilization, transformation, or destruction ofcontaminants.

American Society for Testing and Materials (ASTM)

3

:

Its document titled

Standard Guide for Remediation of Groundwater by Natural Attenuation atPetroleum

Release Sites

defines natural attenuation as the “reduction in mass orconcentration of a compound in groundwater over time or distance from thesource of constituents of concern due to naturally occurring physical, chemical,and biological processes, such as biodegradation, dispersion, dilution, adsorption,and volatilization.”


MONITORED NATURAL ATTENUATION 65

Air Force

4

:

The first document, published in 1995, defines the process as resulting“from the integration of several subsurface attenuation mechanisms that areclassified as either destructive or nondestructive. Biodegradation is the mostimportant destructive attenuation mechanism. Nondestructive attenuation mech-anisms include sorption, dispersion, dilution from recharge, and volatilization.”

Army

5

:

Its report defines natural attenuation as “the process by which contam-ination in groundwater, soils, and surface water is reduced over time…[via]natural processes such as advection, dispersion, diffusion, volatilization, abioticand biotic transformation, sorption/desorption, ion exchange, complexation, andplant and animal uptake.”

In the past, the first question to be asked in consideration of the potential fornatural attenuation at a contaminated site was whether biodegradation of the chem-ical contaminant had been reported. Oftentimes the question was, “Does the bio-geochemistry exist for ongoing degradation?” due to the assumption that the respon-sible microorganisms are ubiquitous in the subsurface. However, in this chapter theterm “natural attenuation” will include all the processes that contribute towards thedecrease in contaminant concentrations.

3.2 APPROACHES FOR EVALUATING NATURAL ATTENUATION

Documenting that contaminant concentration has become very low or detectablein groundwater samples is an important piece of evidence that natural attenuationis working. However, such documentation is not completely sufficient to show thatnatural attenuation is protecting human health and the environment, for three primaryreasons:

• Monitoring of contaminant concentration reductions is not always precise due tothe complex nature of groundwater systems. In some cases the total contaminantmass may have decreased, but the contaminant may have transformed to another,more hazardous chemical form.

• In a few instances reactions that initially cause contaminants to attenuate may notbe sustainable until reasonable cleanup goals are achieved.

• Another situation of concern occurs when natural biogeochemical parameters,such as electron acceptors and electron donors that support attenuation, are usedup before the treatment of contamination is complete.

For these reasons, environmental regulators and others should not rely on simplerules of thumb (such as maximum contaminant concentration data or trends in thesedata over a relatively short time) in evaluating the potential success of naturalattenuation.

The decision to rely on natural attenuation and the confirmation that it willcontinue to work depend on linking monitoring data to a site conceptual model and“footprints” of the underlying mechanisms. Footprints are mappings of concentrationchanges in reactants (contaminant(s), electron acceptors, and donors) or products ofthe biogeochemical processes (such as Cl

–

ion, dissolved Fe

2+

) that degrade or



immobilize the contaminants (Figures 3.1a, b, and c). Footprints can be measuredto document that these transformation or immobilization processes are active at thesite. An observation of the loss of a contaminant, coupled to observation of a fewfootprints, helps to establish which processes are responsible for the decrease incontaminant mass and concentrations. The three basic steps to document naturalattenuation are as follows:

1.

Develop a conceptual model of the site

: The model should show where and howfast the groundwater flows, where the contaminants are located and at whatconcentrations, and which types of natural processes could theoretically affect thecontaminants (Figures 3.2a and b).

2.

Analyze site measurements

: Samples of groundwater should be analyzed chemi-cally to look for footprints of the natural attenuation processes and to determinewhether these processes are sufficient to control the contamination.

3.

Monitor the site

: The site should be monitored until regulatory requirements areachieved to ensure that documented attenuation processes continue to occur.

Although the basic steps are the same for all sites, the level of effort needed tocarry out these steps varies substantially with the complexity of the site. When sitecharacteristics or the controlling mechanisms are uncertain, it will be difficult todevelop the site conceptual model; thus, a large amount of data will be required todocument natural attenuation. In these complex situations, computer modeling maybe necessary, and data on footprints and site characteristics will have to be morethan adequate to develop the model.

Figures 3.1a

Initial vinyl chloride plume at a landfill site in Maryland with radial groundwaterflow from the center of the landfill.



Figures 3.1b

Natural attenuation effects on the vinyl chloride plume. Note: The significantreduction in vinyl chloride concentration and mass due to natural attenuation.

Figures 3.1c

Effects of the primary electron acceptor dissolved oxygen on the attenuation ofVC and Mn along a North-South transect through the middle of the landfill.

Three-dimensional perspective plotof observed vinyl choride concentrationsin groundwater -1996

Landfillboundary

500

200

150

100

20

5

1

0

1200

1000

800

600

400

200

0

Dissolved Oxygen, Redox, and Vinyl Chloride Distribution

LandfillVinyl ChlorideManganeseFe

RedoxDissolved Oxygen

2+

Saprolite

Bedrock

Sand/Gravel



Fig

ure

s 3.

2a

A g

ener

al s

ite c

once

ptua

l exp

osur

e m

odel

(ad

apte

d fr

om A

ST

M,

1997

).

Cur

rent

Dom

estic

Wat

er S

uppl

y W

ell

Fut

ure

Dom

estic

Wat

er S

uppl

y W

ell

Con

finin

g La

yers

?

Con

finin

g La

yers

?

Con

finin

g La

yers

?

Dis

solv

ed P

lum

e

Util

ities

Util

ities

Res

idua

l NA

PL

Aba

ndon

ed W

ell?

Cur

rent

Mun

icip

alW

ater

Sup

ply

Wel

l

Sha

llow

Wat

er T

able



Fig

ure

s 3.

2b

Site

con

cept

ual e

xpos

ure

mod

els.

Prim

ary

Sou

rces

Sec

onda

ryS

ourc

esTr

ansp

ort

Mec

hani

sms

Exp

osur

eR

oute

sR

ecep

tors

Che

mic

al S

tora

geP

ipin

g / D

istr

ibut

ion

Ope

ratio

nsW

aste

Man

agem

ent U

nit

Soi

l or W

aste

Pile

sLa

goon

s or

Pon

dsO

ther

Res

iden

tial

Com

mer

cial

/Indu

stria

lC

onst

ruct

ion

Wor

ker

Rel

evan

t Eco

logi

cal R

ecep

tor

Res

iden

tial

Com

mer

cial

/Indu

stria

lC

onst

ruct

ion

Wor

ker

Rel

evan

t Eco

logi

cal R

ecep

tor

Res

iden

tial

Com

mer

cial

/Indu

stria

lR

elev

ant E

colo

gica

l Rec

epto

r

Res

iden

tial

Com

mer

cial

/Indu

stria

l

Affe

cted

Sub

surf

ace

Soi

ls(>

3 ft

dept

h)

Affe

cted

Sur

face

Soi

ls,

Sed

imen

ts o

rS

urfa

ce W

ater

Affe

cted

Sur

face

Soi

ls(<

3 ft

dept

h)

Dis

solv

edG

roun

dwat

erP

lum

e

Non

-Aqu

eous

Pha

se L

iqui

d(N

AP

L)

Vol

atili

zatio

nan

dA

tmos

pher

icD

ispe

rsio

nA

IR

Inha

latio

n of

Vap

oror

Par

ticul

ates

GR

OU

ND

WAT

ER

Pot

able

Wat

er U

se

SU

RFA

CE

WAT

ER

Rec

reat

iona

l Use

/R

elev

ant H

abita

t

SO

IL

Der

mal

Con

tact

or In

gest

ion

Sto

rmw

ater

/S

urfa

ceW

ater

Tran

spor

t

Win

dE

rosi

on a

ndA

tmos

pher

icD

ispe

rsio

n

Dis

solv

edG

roun

dwat

erP

lum

e

Leac

hing

and

Gro

undw

ater

Tran

spor

t

Mob

ileN

AP

LM

igra

tion

L1282_C03-A_frame Page 69 Tuesday, June 19, 2001 12:59 PM


3.3 PATTERNS VS. PROTOCOLS

3.3.1 Protocols for Natural Attenuation

Within the past few years, many organizations have issued documents providingguidance on evaluating natural attenuation.

1

Among the 14 documents developed bya range of organizations from federal and state agencies to private companies andindustry associations, the available technical protocols address two classes of organiccontaminants only: fuel hydrocarbons and chlorinated solvents (with the exceptionof the Department of Energy (DOE) document). A large body of empirical evidenceand scientific and engineering studies in recent years has been developed to supportunderstanding of natural attenuation of these contaminants — mostly fuel hydro-carbons under certain conditions. However, the natural attenuation of polycyclicaromatic hydrocarbons, polychlorinated biphenyls, explosives, and other classes ofpersistent organic contaminants is not addressed in any protocol.

1

Furthermore,although the DOE document proposes a method for assessing natural attenuationprocesses for inorganic contaminants such as metals, such processes are extremelycomplex, and this document does not adequately reflect this complexity.

6

A recent effort was made to compare the guidelines currently available on naturalattenuation against a list of characteristics of a comprehensive protocol.

1

The con-sensus was that a comprehensive protocol should cover three broad areas:

•

Community concerns

: The protocol should describe a plan for involving theaffected community in decision making, maintaining institutional controls torestrict use of the site until cleanup goals are achieved, and implementing contin-gency measures if natural attenuation fails to continue as expected.

•

Scientific and technical issues

: The protocol should describe how to documentwhich natural attenuation processes are responsible for observed decreases incontaminant concentrations, how to assess the site for contaminant source andhydrogeologic characteristics that affect natural attenuation, and how to assess thesustainability of natural attenuation over the long term.

•

Implementation issues

: The protocol should be easy to follow and should describethe monitoring frequency and various monitoring procedures, in addition to thetraining and expertise required for the personnel carrying out the field implemen-tation.

None of the current documents fulfills all the criteria defined above.

1

To someextent, this reflects the various, and sometimes limited, purposes for which thesedocuments were prepared. Some are detailed technical guides; others are intendedto help ensure consistency in site evaluation within a particular organization (suchas a private corporation or a branch of the military), and others are intended to guidepolicy. Nonetheless, key gaps in the existing body of protocols have to be addressed.

The existing protocols provide little or no discussion of when and how to involvethe public in site decisions and when and how to implement institutional controls.In the few instances where these matters are mentioned, the discussion is typicallybrief, almost in passing. Although most environmental regulatory agencies haveseparate policies that specify procedures for community involvement and



institutional controls, these procedures may be inadequate in cases where naturalattenuation is selected as the remedy. Discussion of when and how to implementcontingency plans in case natural attenuation does not work is also inadequate inmany of the protocols. Further, the protocols do not provide sufficient guidance onwhen and how engineered methods to remove or contain sources of contaminationbenefit natural attenuation.

A major shortcoming of some of the protocols relates to scoring systems used forinitial screening to determine whether a site has potential for treatment by naturalattenuation. Such scoring systems yield a numeric value for the site in question. If thisvalue is above a certain level, the site is judged an eligible candidate for naturalattenuation. Frequently, such scores are used inappropriately as the key factor indeciding whether natural attenuation can be a successful remedy at the site. Moreover,these scores often lead to erroneous conclusions about whether natural attenuation willor will not succeed, due to the complexity of the processes involved and the tendencyof scoring systems to oversimplify them. In addition, the scoring systems developedfor evaluating natural attenuation at petroleum sites are erroneously used to evaluatesites with chlorinated solvents by many practitioners of remediation.

In summary, the existing body of natural attenuation protocols is limited inseveral important areas.

1

Where and how existing protocols can be used to meetregulatory requirements for documenting site cleanup — and whether such protocolsare required at all — is also unclear. Guidance on the use of natural attenuation forremediation has to be developed to cover topics not addressed in existing protocolsand to provide for the use of protocols in regulatory programs.

3.3.2 Patterns of Natural Attenuation

Instead of relying on protocols and scoring systems, an educated screening toolshould be to observe the

patterns

in reduction of contaminant concentrations. Nat-urally attenuating contaminant plumes can take a variety of forms: they might beexpanding, stable, or shrinking, depending on the trends in the spatial variations ofcontaminant concentrations with time (Figures 3.3a, b, and c). Common patterns inall attenuating plumes are a decline in the dissolved contaminant mass with time,and a decline in contaminant concentrations downgradient from the source. Oncethese patterns are observed initially, the following list of questions should be devel-oped to collect additional data to develop a platform demonstrating that MNA is anongoing and continuing process to meet the site cleanup objectives:

• What chemical, physical, and biological processes are in effect to support naturaldegradation of the site-specific contaminants?

• What site biogeochemical conditions are needed for these chemical, physical, andbiological processes to work? Which types of site conditions are optimal? Whichconditions inhibit natural attenuation?

• What level of information is needed to characterize the site fully?• What breakdown products that may be more toxic, persistent, or mobile are created

when the contaminants degrade? How does one prove that contaminants aredegrading into harmless substances?



• What kinds of specific monitoring and testing are needed to determine that thesite and the contaminants are suitable for natural attenuation? Is extensive mon-itoring necessary?

• How long is it reasonable to monitor to ensure that natural attenuation is working?• How viable are institutional controls? Can they be enforced?• Is stabilization by natural attenuation irreversible for metals or other substances?

3.3.2.1 Various Patterns of Natural Attenuation

Removal of Contaminant Sources:

At most contaminated sites, the bulk of thecontaminant mass is in what remediation professionals call “source zones.” Examplesof source zones include landfills, areas of chemical spills, buried tanks that containresidual chemicals, deposits of tars, etc. Some of these sources can be easily located

Figures 3.3a

Expanding plume.

Cross Sectional View

ContaminatedZone

Plan View

MW-3MW-2 MW-1

t1

t0t2

"Contaminant plume is continuing to grow and move downgradient from the source area"

MonitoringWell



and complete or partial removal or containment may be possible. However, othercommon types of sources often are extremely difficult to locate and remove orcontain. One example of a source in this category is chemicals that have sorbed tosoil particles but have the potential to dissolve later into groundwater that contactsthe soil. Another extremely important example is the class of organic contaminantsknown as “nonaqueous-phase liquids” (NAPLs). There are two types of NAPLs:those that are more dense than water (dense nonaqueous-phase liquids, or DNAPLs),and those that are less dense than water (light nonaqueous-phase liquids, orLNAPLs). When released to the ground, these types of fluids move through thesubsurface in a pattern that varies significantly from that of the water flow becauseNAPLs have different physical properties than water. As shown in Figure 3.4a, b,and c, LNAPLs can accumulate near the water table, DNAPLs can penetrate thewater table and form pools along geologic layers, and both types of NAPLs canbecome entrapped in soil pores. These NAPL accumulations contaminate the

Figures 3.3b

Stable groundwater plume.


Plan View

MW-3 MW-2 MW-1

t1t0t2

Contaminant plume is almost stationary over time and concentrations at points within the plume are relatively constant over time with a slight declining trend.

ContaminatedZone

MonitoringWell



groundwater that flows by them by slow dissolution. Common LNAPLs includefuels (gasoline, kerosene, and jet fuel) and common DNAPLs include industrialsolvents (trichloroethene, tetrachloroethene, and carbon tetrachloride) and coal tar.Once they have migrated into the subsurface, NAPLs are often difficult or impossibleto locate in their entirety. Normally, the total mass of a contaminant within sourcezones is significantly larger compared to the mass dissolved in the plume. Therefore,the source usually persists for a very long time. The rate at which contaminantsdissolve from a typical NAPL pool is so slow that many decades to centuries maybe needed to dissolve the NAPL completely by dissolution without any intervention.The potential for success of natural attenuation of various dissolved organic andinorganic compounds is presented in Table 3.1.

Given the persistent nature of contaminant sources, removing them would seemlike a practical way to speed natural attenuation of the contaminant plume (Figure3.4). In many cases, environmental regulators require source removal or containment

Figures 3.3c

Shrinking groundwater plume.


Plan View

MW-3MW-2 MW-1

t1t0

t2

Contaminant plume is receding back toward the source area over time and the concentrations at points within the plume are declining over time.

ContaminatedZone

MonitoringWell



Table 3.1 The Potential for Success of Natural Attenuation for Various Compounds

(adapted from NCR, 2000)

Contaminant Type Dominant Attenuation ProcessesLikelihood of

Success

Organic

Hydrocarbons

BTEX Biotransformation HighGasoline, fuel oil Biotransformation ModerateNonvolatile aliphatic compounds Biotransformation, immobilization LowPAHs Biotransformation, immobilization LowCreosote Biotransformation, immobilization Low

Oxygenated Hydrocarbons

Low-molecular-weight Biotransformation HighAlcohols, ketones, estersMtBE Biotransformation Moderate

Chlorinated Aliphatics

PCE, TCE, carbon tetrachloride Biotransformation ModerateTrichloroethane (TCA) Biotransformation, abiotic

transformationModerate to High

Methylene chloride Biotransformation HighVinyl chloride (VC) Biotransformation Moderate to

HighDichloroethene (DCE) Biotransformation Moderate

Chlorinated Aromatics

Highly chlorinated PCBs, pentachlorophenol, multichlorinated benzenes

Biotransformation, immobilization Low

Less chlorinated PCBs, dioxins Biotransformation LowMonochlorobenzene Biotransformation High

Nitroaromatics

TNT, RDX Biotransformation, abiotic transformation, immobilization

Low

Inorganic

Metals

Ni Immobilization ModerateCu, Zn Immobilization ModerateCd Immobilization LowPb Immobilization ModerateCr Biotransformation, immobilization Low to

ModerateHg Biotransformation, immobilization Low



as part of a natural attenuation remedy. Although requiring source control or removalis good policy for many sites, expert opinions conflict on whether source removalis advisable when using natural attenuation as a remedy, even when such removalis technically feasible.

Nonmetals

As Biotransformation, immobilization LowSe Biotransformation, immobilization Low

Oxyanions

Nitrate Biotransformation ModeratePerchlorate Biotransformation Moderate

Figures 3.4

Various possibilities of source zone contamination.

Table 3.1 The Potential for Success of Natural Attenuation for Various Compounds

(adapted from NCR, 2000) (continued)

Contaminant Type Dominant Attenuation ProcessesLikelihood of

Success

a)Not Enough Mass Spilled to Form an NAPL

t1

t3

t4

t2

b)LNAPL Release

DNAPL Release

MNAPL

MDISS

t1

t3t4

t2

MN > MD

AdsorbedDNAPL

DNAPL Pool

c)

MNAPLMDISS

t1

t3 t4

t2

MN > MD



Goals of source removal should be the following:

• Remove as much contaminant mass as practical to reduce the mass flux of con-taminants emanating from the source zone, thus reducing the concentration of thecontaminant plume rapidly and also reducing the longevity of the required mon-itoring period; and

• Avoid any changes that would reduce the effectiveness of natural attenuation, suchas disturbing the natural dissolution equilibrium from an NAPL source by drillingthrough it and thus increasing the mass flux.

In theory, if one can delineate the source completely and succeed in removing mostof the mass, then a significant benefit may be achieved. There are many case studiesavailable in the literature even for compounds like polycyclic aromatic hydrocarbons(PAHs) plumes in which it appears that, after removal of the source, the plumesattenuated rapidly. However encouraging this example might be, this kind of successmay not always be realized. Particularly, DNAPL sources in fractured bedrock envi-ronments cannot be delineated completely and/or cannot be removed to any significantdegree at a reasonable cost. Hence, source removal options may be rejected becausenone are anticipated to be able to warrant the expense and risks of the removal effortby removing all of the source mass without leaving a significant level of residual mass.

In some cases, source removal efforts may directly and adversely affect naturalattenuation. Most of the negative impacts will be caused mainly by the disturbanceof the equilibrium between the moving groundwater and the quiescent mass ofNAPL, particularly DNAPL. As a precautionary measure, an outside-in approach toinvestigating the source zone is recommended in contrast to an inside-out approach.

Consideration should be given when looking at removal of the source of onetype of contaminant which may adversely affect natural attenuation of another typeand thus result in minimal or no overall benefit. A good example is the removal ofa petroleum hydrocarbon source zone serving as a nutrition source for microbesinvolved in degrading a chlorinated solvent plume. Such an action could slow downor completely shut off natural attenuation of the chlorinated solvent.

Natural Attenuation Capacity (NAC):

The manner in which natural attenuationand active remediation measures (such as source removal, pump and treat, chemicaloxidation, or enhanced bioremediation) are combined depends on the

natural atten-uation capacity

(NAC) of the system. If the NAC is small, for example, activeremediation measures will need to remove or degrade a high proportion of thecontaminant source to protect downgradient receptors. Conversely, if the NAC islarge, less source removal may be required to protect downgradient receptors. Ineither case, it is necessary to quantify the NAC of the biogeochemical system tocombine contaminant source-removal methods with natural attenuation effectively.

Natural attenuation capacity is a concept that refers to the capacity of a bio-geochemical system to lower contaminant concentrations along aquifer flow paths.The NAC of groundwater systems depends on hydrogeologic (dispersion and advec-tion) and biological (biodegradation rates) factors for organic contaminants andprecipitation potential also for heavy metals.



The concept of NAC is useful because it illustrates those characteristics andparameters of a groundwater system that affect the efficiency of natural attenuation.

7

For example, if the biodegradation rate constant is small (

≅

0.001 d

–1

) relative tothe groundwater velocity (~3 ft/day) and aquifer dispersivity (30 feet), the NAC ofthe system also will be small. Because of this small NAC, contaminants will betransported relatively long distances downgradient of the source area (Figure 3.5a).Conversely, if the biodegradation rate is high relative to groundwater velocity andaquifer dispersivity, the NAC will be proportionally higher, and the transport ofcontaminants will be restricted closer to the source area.

Quantitative mathematical techniques in addition to empirical methods are avail-able to estimate NAC. In addition to NAC, the distance that contaminants aretransported in a groundwater system also depends on the contaminant concentrationsat the source area (Figure 3.5b).

Figure 3.5a

The effect of natural attenuation capacity on contaminant transport.

7

Figure 3.5b

The effect of source area concentrations on the distance required to reachcleanup standards.

High NAC

Moderate NAC

Con

cent

ratio

n

Very Low NAC

Distance Along Flow Path

Distance Along Flow Path

High concentration, low NAC

Con

cent

ratio

n

Lower concentration, low NAC

High concentration, higher NAC

Lower concentration, higher NAC

Cleanup Standards



3.4 PROCESSES AFFECTING NATURAL ATTENUATION OF COMPOUNDS

3.4.1 Movement of Contaminants in the Subsurface

Even in the absence of biotic and/or abiotic transformations of a contaminant,the contaminant always is subject to transport processes — meaning that physicalprocesses cause it to move. All important transport processes for subsurface con-taminants can be categorized as dilution, advection, dispersion, or “phase transfer”(from one type of physical medium to another, such as from an NAPL to groundwateror from water to the soil matrix).

3.4.1.1 Dilution (Recharge)

Recharge is the amount of water entering the saturated zone of the water tableat the water table surface, made available mainly by precipitation events. In rechargeareas, flow near the water table is generally downward. Recharge defined in thismanner may therefore include not only precipitation that infiltrates through thevadose zone, but also water entering the groundwater system via discharge fromsurface water bodies. Where a surface water body is in contact with or is part of thegroundwater system, the definition of recharge is stretched slightly. However, suchbodies often are referred to as

recharging lakes

or

streams

.

8

The recharge of thewater table aquifer has two effects on the natural attenuation of a dissolved contam-inant plume: 1) additional water entering the system due to infiltration of precipita-tion or from surface water will contribute to dilution of the plume and 2) the influxof relatively fresh, electron-acceptor-charged water will alter the geochemical pro-cesses and in some cases, facilitate additional biodegradation.

8,9

Recharge from infiltrating precipitation is the result of a complex series ofprocesses in the unsaturated zone. Description of these processes is beyond the scopeof this chapter; however, it is worth noting that the infiltration of precipitation throughthe vadose zone brings the water into contact with the soil and thus may allow theintroduction of electron acceptors (such as NO

3–

and SO

42–

) in addition to the DO inthe recharge water and also dissolved organic carbon (electron donor). Infiltrationtherefore provides fluxes of water, inorganic species, and possibly organic speciesinto the groundwater. In the case of surface water it may be connected as part ofthe groundwater system, or it may be perched above the water table. In either case,the water entering the groundwater system will not only aid in dilution of a con-taminant plume, but it may also add electron acceptors and possible electron donorsto the groundwater.

An influx of electron acceptors will tend to increase the overall assimilationcapacity of the groundwater system. In addition to the introduction of electronacceptors that may be dissolved in the recharge (e.g., dissolved oxygen, nitrate, orsulfate), the infiltrating water may also foster biogeochemical changes in the aquifer.For example, Fe

2+

will be oxidized back to Fe

3+

and will be precipitated out. Thisreprecipitation of Fe

3+

could be again available for reduction by microorganisms.Such a shift may be beneficial for biodegradation of contaminants utilized as electron



donors, such as fuel hydrocarbons or vinyl chloride. However, these shifts can alsomake conditions less favorable for reductive dechlorination.

Evaluating the effects of recharge can be difficult. The effects of dilution mightbe estimated if one has a detailed water budget for the system in question. However,if a plume has a significant vertical extent, it cannot be known with any certaintywhat proportion of the plume mass is being diluted by the recharge. In addition,separating the effects of dilution from other processes of mass reduction may bedifficult. After recharge, the effects of the addition of electron acceptors may beapparent due to elevated electron acceptor concentrations, differing patterns in elec-tron acceptor consumption, or by-product formation in the area of recharge. How-ever, the effects of short-term variations in such a system (which are likely due tothe intermittent nature of precipitation events in most climates) may not easily bequantified. Where recharge is from surface water, the influx of mass and electronacceptors is more steady over time. In this scenario, quantifying the effects of dilutionmay be less uncertain, and the effects of electron acceptor replenishment may bemore easily identified (although not necessarily quantified).

In some cases the effects of recharge-diluting contaminant plumes can be esti-mated with a simple relationship based on the specific discharge of groundwaterpassing through the point of interest and the amount of recharge entering the plumearea. It is imiportant to note that at most sites, recharge will not actually mix withgroundwater in an aquifer but will form a stratified layer on top due to the very lowamount of vertical dispersion characteristic of aquifer systems. Mixing can beassumed in some cases, such as a very thin, unconfined aquifer: the aquifer dis-charges into a surface water body, and the groundwater associated with the rechargeis assumed to be mixed with the original groundwater flowing past a source zone.

8-10

The relationship for estimating the amount of dilution caused by recharge is

(3.1)

Eliminating the width and rearranging gives:

(3.2)

where

C

L

= concentration at distance L from origin assuming complete mixing ofrecharge with groundwater (mg/L)

C

0

= concentration at origin or at distance L = 0 (mg/L)R = recharge mixing with groundwater (ft/yr)W = width of area where recharge is mixing with groundwater (ft)L = length of area where recharge is mixing with groundwater (ft)

C CRWL V

WT VLD

h D

= -Ê

ËÁ��̄0 exp

C CRL

T VL

h D

= -( )

Ê

ËÁÁ

�

¯��0 2exp

L1282_C03-A_frame Page 80 Thursday, March 6, 2003 11:35 AM


V

D

= Darcy velocity of groundwater (ft/yr)T

h

= thickness of aquifer where groundwater flow is assumed to mix com-pletely with recharge (ft)

3.4.1.2 Advection

Transport of a contaminant molecule occurring with the groundwater movementis called advection or convection or bulk flow. Advection occurs in any moving fluid.Thus, contaminants can advect when they are in air in soil pores or in a movingNAPL, as well as in water. Advection transport is illustrated simply by consideringa contaminant that does not react biotically or abiotically (also known as conservativecompound or tracer) in the subsurface and that moves at the average velocity of thegroundwater. Figures 3.6a and b describe this phenomenon. The contaminant movesat exactly the same velocity as the water and does not change from its initialconcentration of C

0

′

at the injection point.

9

Figure 3.6a

Dispersion of a pulse of a tracer substance in a sand column experiment.

Figure 3.6b

Concentration curves showing plug flow with an instantaneous source fromadvection only and from a combination of advection, dispersion, and sorption.

Con

cent

ratio

n (C

)

Distance (x)

t

0

1t 2t

time=t

time=t1

time=t2

Q

Advection andDispersion

Advection,Dispersion,

and Sorption

Advection Only

AdvectionOnly

Initial Contaminant Slug

Distance From Source

0

0.5

1.0

C/CO



The mass flux rate at which a dissolved contaminant moves across a verticalplane in the subsurface is the product of the contaminant concentration and thevelocity of groundwater. Groundwater velocity is governed by three key factorsspecific to each site:

• The

hydraulic gradient

includes gravity and pressure components and is thedriving force for water movement. Water always moves in the direction of higherhydraulic head (which can be thought of qualitatively as elevation) to lower head.

•

Hydraulic conductivity

is the ability of porous rocks or soil sediments to transmitfluids and is measured from field tests or samples. Hydraulic conductivity valuesfor common rocks and sediments vary over ten orders of magnitude from almostimpermeable crystalline rocks to highly permeable gravels; the hydraulic conduc-tivity values for fractured rocks, sand, and clay are between these extremes. Acontaminant plume moving with the groundwater will travel faster through sandlayers, which have high hydraulic conductivity, than through clays of low hydrau-lic conductivity, under the same hydraulic head gradient.

•

Porosity

is a measure of the volume of open spaces in the subsurfaces relative tothe total volume. Like hydraulic conductivity, it depends on the type of geologicmaterial present and can be determined from field tests or samples.

The equation for describing the rate of groundwater flow from one location toanother is known as Darcy’s equation:

(3.3)

where

K

H

= hydraulic conductivity (in units of distance per time)

= hydraulic gradient

V

D

= Darcy velocity (in units of distance per time)

To determine the seepage velocity of a contaminant that travels at the same speedas the groundwater, the Darcy velocity must be divided by the effective porosity

ε

:

(3.4)

K

H

and

ε

can be estimated using various field test methods or laboratory evaluationsof cores taken from the subsurface. Uncertainty is inherent in all such measurements,and this uncertainty must be acknowledged by developing a range of possible flowscenarios.

V KhXD H= − ∆

∆

∆∆

hX

VVD=ε



3.4.1.3 Dispersion

Spreading of contaminants from the main direction of groundwater flow takesplace as the groundwater moves, altering concentrations from those that would occurif advection were the only transport mechanism. This mixing is called hydrodynamicdispersion. The mechanisms causing dispersion within the plume include moleculardiffusion, different water velocities within individual pores, different water velocitiesbetween adjacent pores, and tortuosity of the subsurface flow path (Figure 3.7).Mixing caused by local variations in velocity is also known as mechanical dispersion.Groundwater scientists quantify the combined mixing effect using a hydrodynamicdispersion coefficient DH. Except at very low water velocities, DH increases linearlywith the average speed of groundwater.

The curve labeled “dispersion” in Figure 3.6 a and b illustrates the effects ofdispersion for a conservative contaminant that travels precisely with the water mol-ecules. The solute is detected at the observation well before it would be if advectionwere the only process affecting its movement. Dispersion causes the solute to spread,rather than moving as an unchanged “plug.”

Molecular Diffusion: Molecular diffusion takes place as a result of the contam-inant gradients created within the zones of contamination. It is significant only whenthe groundwater velocities are low, and the diffusive flux of a dissolved contaminant,at steady state, can be described by Fick’s first law.

(3.5)

where

F = mass flux of solute per unit area of timeD = diffusion coefficientC = solute concentration

= concentration gradient

Figure 3.7 Seemingly random variations in the velocity of different parcels of groundwaterare caused by the tortuous and variable route the water must follow.

Flow DirectionAverage Water

C'

B'

A'

CB

A

F Ddcdx

= −

dcdx



For systems where the dissolved contaminant concentrations are changing withtime, Fick’s second law must be applied. The one-dimensional expression of Fick’ssecond law is

(3.6)

where, is the change in concentration with time.

The process of diffusion is slower in porous media than in open water becausethe contaminant molecules must follow more tortuous flow paths. To account forthis, an effective diffusion coefficient D* is used. Fetter estimates a range of 1 ×10–9 to 2 × 10–9 m2/S for D* has been estimated.9(a)

The effective diffusion coefficient is expressed quantitatively as

D* = wD (3.7)

where w is the empirical coefficient determined by laboratory experiments. Thevalue of w ranges greatly from 0.01 to 0.5.9

Mechanical Dispersion: Mechanical dispersion occurs due to variations in flowvelocity because of varying pore throat sizes and tortuosity caused by variations inflow path lengths. An additional cause of mechanical dispersion is variable frictionwithin an individual pore, thus allowing the groundwater flowing in the center ofthe pore to move faster than groundwater flowing next to the soil particle itself.

The component of hydrodynamic dispersion contributed by mechanical disper-sion can be described as:

mechanical dispersion = ∝x V (3.8)

where

∝x = dispersivitiyV = seepage velocity

Advection dispersion equation: The advection-dispersion equation, whichincludes hydrodynamic dispersion, can be described as:8,9

(3.9)

where

c = contaminant concentrationt = timeDH = hydrodynamic dispersionx = distance along flow pathV = seepage velocity

dcdt

Dd

dxc=2

2

dcdt

∂∂

= ∂ − ∂∂

ct

Dc

OxV

cxH

2

2



3.4.2 Phase Transfers

Contaminants will be added or removed from the groundwater when they transferbetween phases. The relevant phases in the subsurface are groundwater (dissolved),soil grains (adsorbed), NAPLs (liquid), and soil gas (air) in the vadose zone. Phasetransfers can increase or decrease the contaminant concentration within the ground-water plume, depending on the transfer mechanism, the contaminant, and thegeochemistry. Although the basic concepts of phase transfer are straightforward,quantification of these transfers often is not easy.

3.4.2.1 Sorption

Many contaminants, including chlorinated solvents, BTEX and dissolved metals,are removed from solution by sorption onto the aquifer matrix, thus slowing themovement of contaminants. This slowing of contaminant transport is called retardationof the contaminant relative to the average seepage velocity of groundwater and resultsin a reduction in dissolved organic concentrations in groundwater. Sorption can alsoinfluence the relative importance of volatilization and biodegradation. Figure 3.6billustrates the effects of sorption on an advancing dissolved contaminant front.

Sorption is a dynamic and reversible reaction; thus, at a given solute concentration,some portion of the contaminant is partitioning out of solution onto the aquifer matrix,and some portion is desorbing and reentering solution. As solute concentrations change,the relative amounts of contaminant that are sorbing and desorbing will change. Forexample, as solute concentrations decrease due to other factors such as biodegradationand dilution, the amount of contaminant reentering solution will probably increase.The affinity of a given compound for the aquifer matrix will not be sufficient to isolateit permanently from groundwater, although for some compounds the rates of desorptionmay be so slow that the adsorbed mass may be considered as permanent residualwithin the time scale of interest. Sorption, therefore, does not permanently removesolute mass from groundwater; it merely retards migration.

The various mechanisms that cause sorption effects to take place within theaquifer matrix are described in detail in Chapter 2. Because of their nonpolarstructure, hydrocarbons most commonly exhibit sorption through the process ofhydrophobic bonding. When the surfaces comprising the aquifer matrix are lesspolar than the water molecule, as is generally the case, there is a strong tendencyfor the nonpolar contaminant molecules to partition from the groundwater and sorbto the aquifer matrix. This phenomenon, referred to as hydrophobic bonding, is animportant factor controlling the fate of many organic pollutants in soils. As describedin Chapter 2, two components of an aquifer have the greatest effect on sorption:organic matter and clay minerals. In most aquifers, the organic fraction tends tocontrol the sorption of organic contaminants.

Sorption Models and Isotherms: Regardless of the sorption mechanism, it ispossible to determine the amount of sorption to be expected when a given dissolvedcontaminant interacts with the materials comprising the aquifer matrix. Bench-scaleexperiments are performed by mixing water-contaminant solutions of various con-centrations with aquifer materials containing various amounts of organic carbon and



clay minerals. The solutions are then sealed with no headspace and left until equi-librium between the various phases is reached. (True equilibrium may require hun-dreds of hours of incubation, but 80 to 90% of equilibrium may be achieved in oneor two days.) The amount of contaminant left in solution is then measured.

The results are commonly expressed as a plot of the concentration of chemicalsorbed (µg/g) vs. the concentration remaining in solution (µg/L). The relationshipbetween the concentration of chemical sorbed (Ca) and the concentration remainingin solution (Cs) at equilibrium is referred to as the sorption isotherm because theexperiments are performed at constant temperature (Figure 2.11). Sorption isothermsgenerally exhibit one of three characteristic shapes, depending on the sorptionmechanism: the Langmuir isotherm, the Freundlich isotherm, and the linear isotherm(a special case of the Freundlich isotherm).

Retardation: As mentioned earlier, sorption tends to slow the transport velocityof contaminants dissolved in groundwater. When the average velocity of a dissolvedcontaminant is less than the average seepage velocity of the groundwater, the con-taminant is said to be retarded. The coefficient of retardation, R, is used to estimatethe retarded contaminant velocity. The variation between the velocity of the ground-water and that of the contaminant is caused by sorption and is quantified by thecoefficient of retardation, defined as:

(3.10)

where

R = coefficient of retardationV = average seepage velocity of groundwater parallel to groundwater flowVc = average velocity of contaminant parallel to groundwater flow

The ratio (V/Vc) describes the relative velocity between the groundwater and thedissolved contaminant. When Kd = 0 (no sorption), the transport velocities of thegroundwater and the solute are equal (V/Vc). If it can be assumed that sorption isdescribed adequately by the distribution coefficient (valid when the fraction oforganic carbon (foc) > 0.001), the coefficient of retardation for a dissolved contam-inant is described by the following equation:9

(3.11)

where

R = coefficient of retardationρb = bulk density of aquiferKd = distribution coefficientn = porosity

RVVc

=

RK

nb d= +1

ρ



The bulk density, ρb, of a soil is the ratio of the soil mass to its field volume.Bulk density is related to particle density by the following equation:

ρb = (1 – n)ρs (3.12)

where n is the total porosity and ρs is the density of soil grains comprising theaquifer. In sandy soils, ρb can be as low as 1.81g/cm3. In aggregated loams andclayey soils, ρb can be as low as 1.1g/cm3.

The sorption relationship shown above expresses the coefficient of retardationin terms of the bulk density and effective porosity of the aquifer matrix and thedistribution coefficient for the contaminant. Substitution of this equation into Equa-tion 3.10 gives

(3.13)

Solving for the contaminant velocity, Vc , gives

(3.14)

Retardation factors can be calculated for several fuel and chlorinated solvent-related chemicals as a function of the fraction of organic carbon content of the soil.The value of R can vary over two orders of magnitude at a site, depending on thechemical in question and the estimated value of porosity and soil bulk density. Earlierinvestigations reported distribution coefficients normalized to total organic mattercontent (Kom). The relationship between fom and foc is nearly constant, and assumingthat the organic matter contains approximately 58% carbon:9

Koc = 1.724 Kom (3.15)

Two methods are used to estimate the distribution coefficient and amount ofsorption (and thus retardation) for a given aquifer-contaminant system. The firstmethod involves estimating the distribution coefficient by using Koc for the contam-inants and the fraction or organic carbon comprising the aquifer matrix. The secondmethod involves conducting batches of column tests to determine the distributioncoefficient. Because numerous authors have conducted experiments to determine Koc

values for common contaminants, literature values are reliable, and it generally isnot necessary to conduct laboratory tests.9

VV

K

nc

b d= +1ρ

VV

K ncx

b d

=+1 ρ



3.4.2.2 Stabilization

The transfer of an organic compound from an NAPL source to the surroundingwater increases the contaminant concentration in groundwater. The rate of transfervaries depending on the type of NAPL. Computation of this transfer rate can becomplex because the transfer rate depends on chemical properties of the contaminantand the NAPL, as well as on resistance at the interface between the water and theNAPL.11 Diffusion of the contaminant within the NAPL itself also can affect thetransfer rate for viscous NAPLs.

DNAPLs: Dense nonaqueous phase liquids (DNAPLs) present in the form ofresidual (held under capillary forces) or free phase (mobile) product may result incontinued long-term contamination of the surrounding groundwater. The marginallysoluble organic contaminants can partition into the aqueous phase at rates slowenough to continue to exist as a nonaqueous phase, yet rapid enough to causesignificant groundwater contamination. DNAPLs can migrate to depths well belowthe water table. As they migrate, they can leave behind trails of microglobules inthe pore spaces of the soil matrix, which effectively serve as long-term sources ofgroundwater contamination.

Current conceptual DNAPL transport models suggest that, when sinking freephase DNAPL encounters a confining layer (e.g., competent clay or bedrock zone),it can accumulate, or “pool,” and spread laterally until it encounters a fracture or analternative path of relatively low flow resistance towards deeper zones.11 In addition,globules can enter pores and be held as a residual phase in capillary suspension.This complex mode of subsurface transport results in unpredictable heterogeneousdistribution of nonaqueous product that is difficult to delineate.

The current lack of appropriate methods for detecting and delineating widelydispersed microglobules of DNAPL has been identified as one of the most significantchallenges today. Investigative techniques that have been used to identify DNAPLsource zones are listed below. It should be noted that some of those techniques arewell proven and extensively field tested, while others are considered relatively new.12

• Soil gas surveys• Visual evidence of soil, rock and/or groundwater samples• Chemical analyses of soil, rock and/or groundwater samples• Enhanced visual identification — shake tests• Enhanced visual identification — UV fluorescence with portable light, dye addi-

tion with Sudan IV or Oil Red O• Accumulation within monitoring wells at target locations• Partitioning interwell tracer tests• Backtracking using dissolved concentrations in wells (the 1% rule)• Surface geophysics• Subsurface geophysics• Cone penetrometer testing (CPT) methods:

• Permeable membrane sensor, membrane interface probe (MIP)• Hydrosparse• Laser induced fluorescence (LIF) techniques• GeoVis



• Raman spectroscopy• Electrochemical sensor probe• Cosolvent injection/extraction technique• Precision injection/extraction (PIX) technique

• Flexible liner underground technologies everting (FLUTE) membrane technique

It is important to recognize that each of the methods listed presents specificadvantages and disadvantages and applicability will be determined by technical andeconomic challenges encountered at each site. Several methods can be complemen-tary in an overall site management plan, and a hybrid approach could be developedto exploit the strengths of the different techniques at the most appropriate and logicaltimes in the site management process. For example, one can initially screen a sitewith a laser induced fluoroscence (LIF) technique or with geophysical techniques,then analyze confirmation soil samples in the field visually, with Sudan IV dye, andin the laboratory for chemical constituents. After determining the location of theDNAPL source zone, discreetly screened or multilevel wells can be installed formonitoring and remediation.

CPT and/or geophysical techniques, integrated with minimally intrusive directpush technologies, can provide the framework for development of the conceptualsite model. Then the refined conceptual site model integrated with hydrogeologicconsiderations can be used for guidance on a sampling plan to define the spatialextent of the contamination.

3.4.2.3 Volatilization

Volatilization reduces the total mass of the contaminant in the groundwatersystem. The potential for volatilization is expressed by the contaminant’s Henry’sLaw Constant and described in detail in Chapter 2. Henry’s Law Constants arewidely available for common volatile contaminants (see Appendix A). Although nota destructive mechanism, volatilization does not remove contaminants from ground-water. In addition to Henry’s Law Constant, other factors affecting the volatilizationof contaminants from groundwater include the contaminant concentration, thechange in contaminant concentration with depth, diffusion coefficient of the com-pound, temperature, and sorption. Because the soil gas often advects and dispersionalso occurs in the gas phase, contaminants transferred to the soil gas often migrateaway from the location at which they volatilize. Volatilization itself does not destroycontaminant mass or permanently immobilize it. Volatilized contaminants can bio-degrade in some circumstances but also can redissolve in infiltrating groundwateror be transported to the surface, where humans may be exposed to the vapors.

3.4.3 Transformation Mechanisms

A variety of reactions transform contaminants. The possible reactions are calledbiogeochemical: all are chemical (prefix chem) and occur in a geological setting(prefix geo), but some are catalyzed by microorganisms (prefix bio). Some bio-geochemical reactions can degrade or transform a contaminant into benign and



harmless end products or immobilize it permanently. A contaminant transformed orimmobilized in these ways no longer contributes to groundwater contamination.Although other reactions do not directly lead to such positive results, they can controlwhether or not the transformation or immobilization reactions take place. Often, asuite of chemical reactions (termed a reaction network) leads to contaminant trans-formation or immobilization. In other instances, the reaction network prevents thecontaminants from being transformed or immobilized and may make natural atten-uation an ineffective remediation strategy.

3.4.3.1 Biodegradation

Microorganisms can cause major changes in the chemistry of groundwater. Theirsmall size and adaptability, as well as the diversity of nutritional requirements fordifferent microbes, enable them to catalyze a wide range of reactions that often arethe basis for natural attenuation. Chemical changes brought about by microorganismscan directly or indirectly decrease the concentrations of certain groundwater con-taminants. Microorganisms use enzymes to accelerate the rates of certain biochem-ical reactions. The most important reactions are reductions and oxidations, togetherknown as REDOX reactions. The reactions involve transfer of electrons from onemolecule to another, which allow the microorganisms to generate energy and grow(Figure 3.8). More discussions on REDOX reactions and microbial electron transfersare provided in Chapters 2 and 4.

Microorganisms reproduce by organizing chemical reactions that create daughtercells composed of cellular components (e.g., membranes, proteins, deoxyribonucleicacid [DNA], cell walls) derived from building blocks that they synthesize or scavengefrom the environment.1 The chemical reactions are made possible by enzymes —protein molecules that bring together the chemicals in a way that allows them toreact quickly (Figure 3.9). The reactions are driven to completion by the expenditureof cellular energy in the form of a chemical known as adensoine triphosphate (ATP),

Figure 3.8 Conceptual description of microorganisms gaining energy and utilizing the sub-strate for growth.

Energy

Electrons

Electrons and Carbon

2

ElectronAcceptor(e.g., O )

New Cells

+OrganicContaminant



which can be thought of as a cellular fuel. Like all living organisms, microorganismsgenerate ATP by catalyzing redox reactions: they transfer electrons from electron-rich chemicals to electron-poor chemicals. The technical term for the electron-richchemical is electron donor substrate. As an analogy, human metabolism involvestransfer of electrons from chemicals derived from ingested food (the donor substrate)to oxygen (the acceptor substrate) inhaled from the air.1

When cells remove electrons from the donor substrate, they do not transfer theelectrons directly to the acceptor substrate. Instead, they transfer the electrons tointernal electron carriers as shown in Figure 3.9. Although electrons held by thecarriers can be used for many purposes, the major purpose is to generate ATP througha process called respiration. In respiration, the electrons are passed from carrier tocarrier until they reach the electron-acceptor substrate. Since this is the last moleculeto receive the electrons, it is called the terminal electron acceptor. The need for ATPproduction forces all microorganisms to have one or more electron-donor and elec-tron-acceptor pairs, and these materials largely define the metabolism of individualmicroorganisms. The amount of energy yielded varies depending on the electrondonor and electron acceptor used.

Figures 3.9 Conceptual diagram of microbial activity to derive energy for growth andmultiplication (adapted from NRC, 2000).

ElectronDonor

Oxidized Donor Product UnicellularMicroorganism

NAD

NAD

ATP

NADH2

NADH2

ADP+Pi

Respiration

ElectronAcceptor

Reduced AcceptorProduct

SynthesisandMaintenance



Ideally, all biologically mediated reactions produce energy for microbial growthand reproduction. Biologically mediated electron transfer results in oxidation of theelectron donor, reduction of the electron acceptor, and the population of usableenergy (quantified by the Gibbs free energy of the reaction ∆Gv

o ). Table 3.2 presentsa few select electron acceptor and electron donor reactions and calculated ∆Gy

o

values.9 Negative values indicate an energy-producing reaction, otherwise called anexothermic reaction, and will proceed from left to right. The value of ∆Gy

o can beused to estimate how much free energy is consumed or produced during the reaction.Positive values indicate an endothermic reaction; for the reaction to proceed fromleft to right energy must be put into the system. Microorganisms will not investmore energy into the system than can be released and must couple an endothermicwith an exothermic reaction to derive energy and grow.

Collectively, microorganisms can use a wide range of electron donors, includingboth organic and inorganic chemicals. Electron acceptors are more limited. Commonelectron acceptors include O2, NO3

– , NO2– , SO4

2– , CO2, Fe(III), and Mn(IV). Oxygenhas a special status because of its importance in many environments and reactions.Microbial use of oxygen as an electron acceptor is called aerobic metabolism; micro-bial use of electron acceptors other than oxygen is called anaerobic metabolism.

When biotransformation of a particular contaminant leads directly to energygeneration and the growth of more microorganisms, the contaminant is known as aprimary substrate (see Figure 3.8). However, the reactions that lead to microbialmetabolism of contaminants may not be part of cell-building or energy-generatingreactions. An important category of such biotransformations is cometabolism. Come-tabolism is the fortuitous degradation of a contaminant when other materials areavailable to serve as microorganisms’ primary substrates. Cometabolic reactionsoften occur because the enzymes designed for metabolizing primary substratesfortuitously transform the cometabolic substrate.

It is important to note the historic debate on the use of the word cometabolism forthe microbially catalyzed process described above.13,14 One school of thought, propa-gated by classical microbiologists, insists that usage of either the term cometabolismor the term cooxidation to describe conversions of nongrowth substrates by nonpro-liferating microbial populations in the absence of a metabolizable cosubstrate wouldbe inappropriate. The enzymatic conversion of a substrate by a nonproliferating micro-bial population because an enzyme of broad specificity and conversion capability isin proximity to the substrate might at best be described as bioconversion. There is noco- (with or together) activity concerned with such an event.

First-Order Decay Model: One of the most commonly used expressions forrepresenting the biodegradation of an organic compound involves the use of anexponential decay relationship:

C = C0e–kt (3.16)

where

C = biodegraded concentration of the chemical at time tC0 = initial concentrationk = rate of decrease of the chemical (units of 1/time) [T–1]



Table 3.2 Half-Cell Reactions for Some of the Common Electron Acceptors and Donors (adapted from

Wiedemeier et al., 1999)

Half-Cell Reaction

∆∆∆∆

G

ro

(kcal/mol e

–

)

4e

–

+ 4H

+

O

2

⇒

2H

2

O –18.5Aerobic respiration

5e

–

+ 6H

+

+ NO

3–

⇒

0.5N

2

+ 3H

2

O –16.9Denitrification

2e

–

+ 4H

+

+ MnO

2

⇒

Mn

2+

+ 2H

2

O –8.6Manganese reduction

e

–

+ Fe

3+

⇒

Fe

2+

–17.8Fe(III) reduction

8e

–

+ 9.5H

+

+ SO

42–

⇒

0.5HS– + 0.5H

2

S + 4H

2

O 5.3Sulfate reduction

8e

–

+ 8H

+

+ CO

2

⇒

CH

4

+ 2H

2

O 5.9Methanogenesis

C

2

Cl

4

+ H

+

+ 2e

–

⇒

C

2

HCl

3

+ Cl

–

–9.9PCE reductive dechlorination

C

2

HCl

3

+ H

+

+ 2e

–

⇒

C

2

H

2

Cl

2

+ Cl

–

–9.6TCE reductive dechlorination

C

2

H

2

Cl

2

+ H

+

+ 2e–

⇒

C

2

H

3

Cl + Cl

–

–7.2

cis

-DCE reductive dechlorination

C

2

H

3

Cl + H

+

+ 2e

–

⇒

C

2

H

4

+ Cl

–

–8.8VC reductive dechlorination

C

2

H

3

Cl

3

+ H

+

+ 2e

–

⇒

C

2

H

4

Cl

2

+ Cl

–

–10.3TCA reductive dechlorination

1

/

2

H

2

⇒

H

+

+ e

–

–9.9Hydrogen oxidation

1

/

4

CH

2

O +

1

/

4

H

2

O

⇒

1

/

4

CO

2

+ H

+

+ e

–

–10.0Carbohydrate oxidation

12H

2

O + C

6

H

6

⇒

6CO

2

+ 3O H

+

+ 3Oe

–

–7.0Benzene oxidation

14 H

2

O + C

6

H

5

CH

3

⇒

7CO

2

+ 36H

+

+ 36e

–

–6.9Toluene oxidation

20H

2

O + C

10

H

8

⇒

10CO

2

+ 48H

+

+ 48e

–

–6.9Naphthalene oxidation

4H

2

O + C

2

H

3

Cl

⇒

2CO

2

+ 11H

+

+ 10e

–

+ Cl

–

–11.4Vinyl chloride oxidation

12H

2

O + C

6

H

5

Cl

⇒

6CO

2

+ 29H

+

+ 28e

–

+ Cl

–

–8.0Chlorobenzene oxidation



First-order rate constants are often expressed in terms of half-life for thechemical:

(3.17)

The first-order decay model shown in Equation 3.16 assumes that the solutedegradation rate is proportional to the solute concentration. The higher the concen-tration, the higher the degradation rate. This method is usually used to simulatebiodegradation of contaminants dissolved in groundwater. Modelers using the first-order decay model typically use the first-order decay coefficient as a calibrationparameter and adjust the decay coefficient until the model results match the fielddata. With this approach, uncertainties in a number of parameters (e.g., dispersion,sorption, biodegradation) are lumped together in a single calibration parameter.

Regression methods are commonly used to obtain approximations of site-specificdegradation rates (first-order) from log-linear plots of concentration vs. time. Thisinvolves fitting an exponential regression to approximate the trend in the data. Thistype of approximation can be used to evaluate trends at an individual well or forseveral wells along a flow path. When individual wells are being evaluated, theanalytical data should be used from multiple sampling events, and the time elementin the plot represents the temporal arrangement of the data. When multiple wellsalong a flow path are being evaluated, the analytical data from a single samplingevent can be used; the time element in the plot represents groundwater travel timebetween the wells.15

Electron-Acceptor-Limited or Instantaneous Reaction Model: The electron-acceptor-limited model (traditionally called the instantaneous reaction model) was firstproposed in 1986 for simulating the aerobic biodegradation of petroleum hydrocar-bons.9,16 It was observed that microbial biodegradation kinetics are fast in comparisonwith the transport of oxygen and that the growth of microorganisms and utilization ofoxygen and organics in the subsurface can be stimulated as an electron-acceptor-limitedor instantaneous reaction between the organic contaminant and oxygen.

From a practical standpoint, the instantaneous reaction model assumes that therate of utilization of the contaminant and oxygen by the microorganisms is veryhigh, and that the time required to biodegrade the contaminant is very short, almostinstantaneous, relative to the seepage velocity of the groundwater. Using oxygen asan electron acceptor, for example, biodegradation is calculated using the expression:

(3.18)

where

∆CR = change in contaminant concentration due to biodegradationO = concentration oxygenF = utilization factor, the ratio of oxygen to contaminant consumed

tk1 2

0 693= .

∆COFR =



The variable F is obtained from the oxidation-reduction reaction involving theorganic and the given electron acceptor.

Biodegradation of Organic Contaminants: Organic contaminants vary widelyin their susceptibility to transformation by microorganisms. Some contaminants arehighly biodegradable, while others resist degradation. In general, the more degrad-able contaminants have simple molecular structures (often similar to the structuresof naturally occurring organic chemicals), are water soluble and nontoxic, and canbe transformed by aerobic metabolism (Figure 3.10). In contrast, organic contami-nants that resist biodegradation may have complex molecular structures (especiallystructures not commonly found in nature), low water solubility or an inability tosupport microbial growth, or they may be toxic to the organisms.

Microorganisms can completely convert some organic contaminants to carbondioxide and water, while they are capable of only partial conversions of others.Complete conversion to carbon dioxide is called “mineralization.” In some cases,the products of partial conversion are more toxic than the original contaminant. Vinylchloride is an example of a highly toxic chemical that results from incompletebiodegradation of chlorinated solvents.

The following discussion explains how microbial transformations occur for var-ious organic contaminant classes. It describes all of the elements of some metabolicpathways because these illustrate the core concepts of biodegradation. Biodegrada-tion pathways for most contaminants are extremely complex, so these pathways arenot described in detail.

Petroleum hydrocarbons are a highly varied class of naturally occurring chemicalsused as fuels in a variety of commercial and industrial processes. Biodegradationpotential varies depending on the type of hydrocarbon.

Benzene, Toluene, Ethylbenzene, and Xylene (BTEX): Benzene, Toluene,Ethylbenzene, and Xylene are components of gasoline. Because of their widespreaduse and because BTEX storage tanks commonly leaked in the past, BTEX arecommon groundwater contaminants. A large body of scientific research exists onthe biodegradation and natural attenuation of BTEX. However, the effectiveness of

Figure 3.10 Schematic diagram describing the mechanisms by which a contaminantbecomes available for biodegradation.

Accessible

Dissolved

Not Accessible

Gaseous

Sorbed

Nonaqueous



MNA via intrinsic bioremediation, as with any other contaminant, depends on therelationship between the contaminant biodecay rate and the groundwater velocity.

BTEX are easily biodegraded to carbon dioxide by aerobic microorganisms andcan also biodegrade under anaerobic conditions. When the volume of BTEX is smallenough and/or the supply of oxygen is large enough, microbes can degrade all ofthe BTEX components within the aerobic zones of a contaminated site. When oxygenis depleted in an advancing contaminant plume, anaerobic conditions can developand lead to the formation of as many as five different downgradient zones, each witha different terminal electron acceptor (Figure 3.11). In these zones, BTEX degrada-tion processes are slower and less reliable than when oxygen is present.

Of the possible electron acceptors, oxygen yields the most energy. Once oxygenis depleted, nitrate is next as the most energy-yielding terminal electron acceptor.If nitrate is abundant in groundwater, zones in which microbes use nitrates as theelectron acceptor will develop. A Mn(IV)-reducing zone may develop next if Mn(IV)is present in the subsurface mineral matrix (although the coupling of Mn reductionto BTEX degradation has not been well studied). Upon depletion of the Mn(IV),Fe(III) reduction will prevail if iron oxide minerals are present. In the next zones,sulfate and CO2 will serve as electron acceptors.

Based on electron acceptor abundance, Fe3+, Mn4+, and SO42– reduction by bacteria

may play a dominant role in intrinsic bioremediation under certain geologic conditions.Both Fe3+ and SO4

2– reduction processes involve mineral phases and may not be prop-erly understood by evaluating only groundwater concentrations. Fe and S mineralanalyses, from soil samples, should be incorporated in natural attenuation studies,when the geologic conditions are appropriate. Fe and S mineral analyses may not bewidely utilized in natural attenuation studies because of the inherent difficulty in solidsample collection, preservation, and analysis of bioavailable minerals.17

Figure 3.11 Conceptualization of the dominant terminal electron acceptor process (TEAP)in advancing BTEX plume.

Source Area

4 and 5

2 and 3

1 - Aerobic Zone

- Transient Anaerobic Zones

- Core Anaerobic Zones

GroundwaterFlow Direction

12

34

2

Aerobic Zone

O

3NOReduction

-4+3+Fe /MnReduction

SOReduction

2-45Methanogenic

Encroachment of the Aerobic Fringe



Many field studies of BTEX biodegradation in the subsurface have been carriedout. For example, several lines of evidence indicated that all BTEX componentswere biodegrading mainly in the Fe(III)-reducing zone of an aquifer in Bemidji,Minnesota, that was contaminated with crude oil.

18-20

At a petroleum spill site inSouth Carolina, toluene, but not benzene, was metabolized as it moved through asulfate-reducing zone.

1,21

In a recent study of an anaerobic gasoline-contaminatedaquifer in California, researchers injected BTEX components (along with bromideas a tracer) and either sulfate or nitrate into a sandy aquifer. Periodic withdrawal ofsamples from the injected zones showed that under nitrate-reducing conditions,toluene, ethylbenzene, and

m

-xylene, (but not benzene) were transformed in lessthan ten days. Under sulfate-reducing conditions, toluene,

m

-xylene, and

o

-xylenewere completely transformed in 72 days, while benzene loss was uncertain.

1,22

During a recent study

24

in short term (< 2 weeks) incubations, addition of sulfateslightly stimulated benzene degradation and caused a small decrease in the ratio ofmethane to CO

2

production from benzene. However, in long term (>100 days)incubations, sulfate significantly stimulated benzene degradation with a completeshift to CO

2

as the end product of benzene degradation. The addition of Fe(III) andhumic substances had short- and long-term effects that were similar to the effectsof sulfate amendments.

A novel

in situ

respiration technique was reported recently to measure and predictnatural attenuation of petroleum compounds in the subsurface. Monitoring CO

2

andCH

4

produced

in situ

, and their radiocarbon (

14

C), stable carbon (

13

C), and deutrium(D) signatures provides a novel method to assess anaerobic microbial processes. The

in situ

anaerobic respiration test was conducted by injecting a large volume ofindustrial grade Argon, an inert gas, into the subsurface to replace CO

2

and CH

4

,followed by monitoring the production of CO

2

and CH

4

.

23

Figures 3.12a and b showthe formula of BTEX compounds.

Figures 3.12a

Benzene formula and simplified representations.

Figures 3.12b

Structures of single-ring aromatic hydrocarbons.

C

CC

C

CCH

H

H

H H

H

OR OR

Benzene Toluene m-Xylene Ethylbenzene

3CHCH3

CH3

CH3

L1282_C03-B_frame Page 97 Monday, June 18, 2001 10:07 AM


Polycyclic Aromatic Hydrocarbons:

In contrast to BTEX, Polycyclic AromaticHydrocarbons (PAHs) biodegrade very slowly. PAH contamination comes mostlyfrom fossil fuel use and the manufactured-gas industry.

1

Groundwater contaminationat manufactured gas plants has persisted for decades because of the slow, continuousdissolution of PAHs from subsurface coal tar. PAHs are compounds that havemultiple rings in their molecular structure. These compounds have complex molec-ular structures and low water solubility, and they tend to sorb strongly to solids inthe subsurface. However, because PAHs dissolve slowly, natural attenuation couldcontrol the contamination even if biodegradation is slow, as long as it occurs at thesame rate as or faster than dissolution.

The fate of PAHs in subsurface systems is governed largely by their hydrophobicnature (the reason for their low solubility and tendency to attach to surfaces). PAHmolecules held within NAPLs or adsorbed to surfaces cannot be biodegraded. Con-sequently, understanding dissolution and the sorption processes for PAHs often isthe key to understanding biodegradation and natural attenuation potential.

Biodegradation of PAHs depends on the complexity of the chemical structureand the extent of enzymatic adaptation. In general, PAHs that contain two or threerings such as napthtalene, anthracene, and phenanthrene are degraded at reasonablerates when O

2

is present. Studies have shown that some microorganisms can metab-olize dissolved PAHs composed of up to five benzene rings. Microorganisms gen-erally use oxygenase enzymes to initiate the biodegradation; these reactions requirethe presence of oxygen. However, microbial degradation of PAHs with lower molec-ular weights (fewer benzene rings) can occur under nitrate-reducing and sulfate-reducing conditions.

28,29

Oxygenated Hydrocarbons:

Microbiologists and remediation engineers havelong known that low molecular weight alcohols, ketones, esters and ethers biode-grade readily particularly under aerobic conditions. The polar oxygen atom in MtBE(CH

3

–O–C(CH

3

)

3

) causes the molecule to be much more hydrophilic than othergasoline constituents. However, one prominent oxygenated hydrocarbon methyl tert-butyl ether (MtBE) was thought to be resistant to biodegradation because of its stablemolecular structure and its reactivity with microbial membranes. MtBE has beenused as an octane enhancer in gasoline since the late 1970s and recently has beenused up to 15% by volume.

MtBE has relatively high water solubility (43,000–54,000 ppm in comparisonto 1780 ppm for benzene), a very low Henry’s Law Constant (0.022 in comparisonto 0.22 for benzene), and very weak sorbtion to soil (log Koc = 1 to 1.1 in comparisonto about 1.5–2.2 for benzene). Until very recently, MtBE was considered nonbiode-gradable in the subsurface; a prestigious state of the science report by the NationalResearch Council, in the year 2000, stated that “present knowledge on MtBE bio-degradation is limited ….” The report further pointed out that the process was notwell understood and therefore the likelihood of success for natural attenuation as aremediation solution for MtBE contaminated sites was low.

1

However, a number of recent studies have demonstrated natural MtBE biodeg-radation in the field.

1,30-34

It is still unclear how prevalent this biodegradation is andwhether the rates are rapid enough to restrict and eventually shrink groundwater



plumes. Also, it is unknown if the potential for natural MtBE biodegradation couldbe reasonably predicted using some indicator parameters.

A recent study, which included many field sites, suggested that natural biodeg-radation of MtBE and TBA under anaerobic subsurface conditions at some sitesmay control migration of MtBE and TBA plumes. There appeared to be a goodcorrelation between strongly anaerobic plume biogeochemistry and natural biodeg-radation of MtBE.

35

To date no one has shown MtBE biodegradation in the laboratoryunder sulfate reducing conditions. It is important to note from this study that MtBEand TBA naturally biodegrade only under strongly anaerobic conditions (preferablyunder methanogenic conditions) and the rates of biodegradation (at the sites wherethis happens) are comparable to those of benzene.

35

Chlorinated Aliphatic Compounds:

The chlorine atoms added to aliphaticorganic molecules to produce these chemicals significantly change many properties,including solubility, volatility, density, hydrophobicity, stability, and toxicity. Thesechanges are valuable for commercial products, but also can make the compoundsless biodegradable. Several good reviews have been published on the biodegradationof the small (one- and two-carbon) chloroaliphatic compounds.

25

The biodegradationpotentials of many chlorinated aliphatics are discussed extensively in Chapter 4.

Researchers first demonstrated the potential for anaerobic biotransformation ofchlorinated aliphatic hydrocarbons during the early 1980s.

36

Subsequent studies haveshown that these compounds can biotransform under a variety of environmentalconditions in the absence of oxygen. In general, the biotransformation rates, partic-ularly for chlorinated compounds with more than two chlorine atoms in the molecule,are higher under anaerobic conditions.

Exceptions to the general rule that chlorinated aliphatic hydrocarbons requirespecial environmental conditions for biodegradation to occur are methylene chloride,known also as dichloromethane, and vinyl chloride. Methylene chloride and vinylchloride can support the growth of a wide range of microorganisms (both aerobicand anaerobic) under a range of environmental conditions. Methylene chloride andvinyl chloride therefore are likely to be treated successfully by natural attenuationat a much broader range of sites than other chlorinated aliphatics compounds. Inaddition to methylene chloride and VC, there are a few other chloroaliphatic com-pounds which will degrade under aerobic conditions as growth substrates or ascometabolic substrates.

Natural biotransformation of chloroaliphatics is most likely where excess organicmaterial is available to serve as an electron donor and biogeochemical conditionssupport a reducing environment. Successful intrinsic reductive dechlorination hasbeen found to occur in the presence of other electron-donating organic pollutants,such as those from leaking sewage systems, BTEX, and phenol. Reductive dechlo-rination to VC and ethene appeared to be driven by fuel hydrocarbon co-contaminantsin the center of many mixed contaminant plumes. Down gradient, where carbonsources became depleted, VC was oxidized further by iron and aerobic oxidation.When soil organic matter serves as electron donor, reductive dechlorination mayalso be observed down gradient of the plume and dechlorination products mayaccumulate.



Chlorinated Aromatic Compounds:

Bacteria able to degrade all but the mostcomplex chloroaromatic compounds have been discovered during the past 20 years.

1,25

Chlorobenzenes including hexachlorobenzene can be sequentially dechlorinated tochlorobenzene under methanogenic conditions in soil slurries

37

(Figure 3.13). Reduc-tive dechlorination of chlorobenzene has not been reported, but chlorotoluenes aredechlorinated to toluene in the preceding methanogenic systems and it seems likelythat chlorobenzene could serve as a substrate for reductive dechlorination.

Chlorobenzenes up to and including tetrachlorobenzene are readily biodegradedunder aerobic conditions. Bacteria able to grow on chlorobenzene,

25,38

1,4-dichloroben-zene,

25,38

1,3-dichlorobenzene,

25

1,2-dichlorobenzene,

25

1,2,4-trichlorobenzene,

25

and1,2,4,5-tetrachlorobenzene

25

have been isolated and their metabolic pathways identi-fied. The pathways for aerobic degradation are remarkably similar and lead to therelease of the chlorine as HCl. Chlorobenzenes are very good candidates for naturalattenuation under either aerobic or anaerobic conditions. Aerobic bacteria able to growon chlorobenzene have been detected at a variety of chlorobenzene-contaminated sitesbut not at adjacent uncontaminated sites,

25,39

providing strong evidence that they areselected for their ability to derive carbon and energy from chlorobenzene degradation

in situ

. Removal of multiple chlorines as HCl consumes a large amount of alkalinityand produces a considerable drop in the pH of unbuffered systems which could leadto a loss of microbial activity at some sites.

Although the benzene ring that is the nucleus of chlorinated aromatic compoundsis relatively easy for microorganisms to biodegrade, the addition of chlorine atomscompletely alters the biodegradability of benzene. The number and position ofchlorine atoms on the benzene ring determine how biodegradable the compoundwill be. Compounds with many chlorine atoms may not be biodegradable at all underaerobic conditions; however, under special environmental conditions, these com-pounds can be reductively dechlorinated by the same type of microbial dechlorina-tion process that can occur for chlorinated aliphatic compounds.

25,40-42

As the reduc-tive dechlorination process removes chlorine atoms from the benzene ring, themolecules become more susceptible to biodegradation by aerobic microbes. Whenenvironmental conditions are right, natural attenuation may be able to control halo-genated aromatic compounds, but these conditions generally are uncommon.

Chlorophenols and chlorobenzoates are dechlorinated under anaerobic conditionsin sediments and subsurface material.

25,43,44

In some instances the dechlorination clearlyyields energy for the growth of the specific bacteria. In other examples the dechlori-nation is specific and enriched in the community, but has not been rigorously linked

Figure 3.13

Reductive dechlorination of hexachlorobenzene under anaerobic conditions.

Cl

Cl

Cl Cl

Cl

Cl Methanogenic Conditions

ClCl - -Cl Cl - Cl - Cl -



to energy production. Addition of small fatty acids or alcohols as electron donors orsources of carbon can enhance the process of reductive dechlorination.

Aerobic pathways for the degradation of chlorophenols and chlorobenzoates areinitiated by oxygenase-catalyzed attack on the aromatic ring and subsequent removalof the chlorine after ring fission or hydrolytic replacement of the chlorine with ahydroxyl group. Bacteria able to grow on chlorophenols and chlorobenzoates arewidely distributed and readily enriched from a variety of sources, indicating a highpotential for natural attenuation. Chlorophenols are unusual among the syntheticcompounds discussed here in that they can be very toxic to microorganisms. Theyare often used as biocides; therefore, high concentrations can dramatically inhibitbiodegradation. Inoculation with specific bacteria has been helpful in overcomingtoxicity and stimulating degradation of chlorophenols.

25,43

Pentachlorophenol deserves special consideration because it has been widelyused as a wood preservative and has been released into the environment throughoutthe world. Reductive dechlorination under methanogic conditions can lead to min-eralization.

25,43

Aerobic bacteria catalyze the replacement of the chlorine in the 4position by a hydroxyl group to form tetrachlorohydroquinone. Subsequent reductivedechlorinations lead to the formation of ring fission substrates. Bacteria able todegrade pentachlorophenol are widely distributed, and both experimental and full-scale bioremediation projects have been successful in field applications

43

(Figure3.14). Adding selected strains has been helpful in some instances; in others, indig-enous strains have been used. Wood treatment facilities typically are contaminatedwith complex mixtures of organic compounds; therefore, investigations of toxicitymust be conducted for each site under consideration. Natural attenuation of pen-tachlorophenol has been reported, because specific bacteria able to use it as a growthsubstrate are enriched at contaminated sites. However, rates seem to be low at manysites due to toxicity and bioavailability of the pentachlorophenol.

Although polychlorinated biphenyl (PCB) use has been banned, these chemicalsare still present in the environment, especially in sediment and aquatic systems, andtheir persistence is due in part to their resistance to biodegradation.

1,45

PCBs consistof up to ten chlorine and hydrogen atoms attached to a structure consisting of twobenzene rings attached by a bond between carbon atoms. Chemical synthesis cancreate various possible combinations — called “congeners” — of chlorine andhydrogen atoms in the ten positions (Figure 3.15). PCBs were marketed as mixturesof congeners called Aroclors (the Monsanto Corporation trade name), characterizedaccording to average chlorine content.

PCBs have been studied extensively because of their stability, toxicity, andbioaccumulation potential.

1,46

Anaerobic transformation of PCBs is catalyzed bybacteria in aquatic sediment from a wide range of contaminated and uncontaminatedsites. Higher activities in contaminated sites suggest that the dechlorination reactionsprovide a selective advantage to the microbial population, indicating the potentialfor significant natural attenuation. A number of studies have clearly demonstratedthat natural attenuation of PCB is taking place in anaerobic sediments at significantrates. Methanogenic conditions in freshwater sediments seem to provide the highestrates of reductive dechlorination.



Fig

ure

3.1

4

Pat

hway

s of

pen

tach

loro

phen

ol (

PC

P)

degr

adat

ion.

OC

H3

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

Cl

OH

Cl

Cl

Cl

Cl

OH

OH

Cl

Cl

Cl

Cl

OH

Cl

Cl

Cl

Cl

OH

Cl

OH

Cl

Cl

Cl

Cl

Cl

Cl

OH

OH

Cl

Cl

Cl

Cl

Cl

Cl

OH

Cl

Cl

Cl

OH

Cl

Cl

Cl

HO

OC C

l

CO

OH

OH

Cl

Cl

Rin

g C

leav

age

Rin

g C

leav

age

OH

Cl

Cl

Cl

Cl

3O

CH

OC

H3

Cl

Cl

Cl

Cl

3O

CH



Dechlorination converts the more highly chlorinated congeners to less chlori-nated products containing one to four chlorines. Complete dechlorination does notoccur, but the depletion of the more highly chlorinated congeners dramaticallyreduces not only the toxic and carcinogenic potential, but also the bioaccumulationpotential. A variety of dechlorination patterns have been identified as a function ofthe microbial community involved. The patterns are constant within a given micro-bial community or enrichment, supporting the premise that dechlorination providesa selective advantage to the organisms involved. The electron donors for the dechlo-rination in sediment are unknown. Addition of exogenous carbon sources does notstimulate the reaction. In contrast, “priming” the mixtures with low levels of bro-mobiphenyl or specific isomers of tetrachlorobipheny

l1,46,47

seems to selectivelyenrich a population of PCB-dechlorinating bacteria and dramatically stimulate thedechlorination of other congeners.

The lower chlorinated PCB congeners, whether part of the original Arochlormixture or derived from reductive dehalogenation, are biodegraded by aerobic bac-teria.

25,48

The initial attack is catalyzed by a 2,3- or 3,4-dioxygenase followed by asequence of reactions that leads to ring cleavage and accumulation of chloroben-zoates readily degraded by a variety of bacteria. The enzymes that oxidize PCBsare produced by bacteria growth or biphenyl, and addition of biphenyl to slurry-phase reactors stimulates the growth and activity of PCB degraders. Such stimulationhas been shown to be effective in the field. There is also good evidence that aerobicPCB degradation is taking place in contaminated river sediments.

48

It seems clear that reductive dechlorination is ongoing at a wide range of PCB-contaminated sites. The strategy of anaerobic dechlorination followed by aerobicdegradation seems to be particularly effective with PCB whether in an engineeredsystem or in natural systems occurring during resuspension of anaerobic sediments.To date, the complete biodegradation of PCB is slow and difficult to predict orcontrol in the field. Several new strategies, including construction of novel strains,may increase the potential for effective PCB biodegradation.

Nitroaromatic Compounds:

The literature on biodegradation of nitroaromaticcompounds has been reviewed recently.

25,49,50

These compounds are subject to reduc-tion of the nitro groups in the environment under either aerobic or anaerobic con-ditions. Reduction does not lead to complete degradation in most instances and could

Figure 3.15

Structure of PCB.

CIm

m = 1 5 n = 1 5

CIn



be considered nonproductive for purposes of natural attenuation. In contrast, aerobicbacteria able to grow in nitrobenzene, nitrotoluenes, dinitrotoluenes, dinitrobenzene,nitrobenzoates, picric acid, and other nitrophenols have been isolated from a varietyof contaminated sites, suggesting that natural attenuation is taking place. Mineral-ization of dinitrotoluenes in aquifer material from a dinitrotoluene-contaminated sitewas measured recently.

51

It was concluded that the indigenous microorganismsprovide a significant degradative capacity for the contaminant.

The simple nitroaromatic compounds can be considered excellent candidates fornatural attenuation as long as the degradation process yields a selective advantage.Some of the compounds, including 3-nitrophenol, nitrobenzene, 4-nitrotoluene, and4-nitrobenzoate, are degraded via catabolic pathways that minimize the use ofmolecular oxygen and are particularly well suited for operation in the subsurfacewhere oxygen is limiting. The pathways all involve a partial reduction of the mol-ecule prior to oxygenative ring fission. For example, the first three steps in thepathway for degradation of nitrobenzene can take place in the absence of oxygen,

49

which is required only for ring fission and subsequent metabolism.Mixtures of the isomeric nitro compounds can be problematic for microbial

degradation. For example, the industrial synthesis of polyurethane produces largeamounts of 2,4- and 2,6-dinitrotoluene in a ratio of 4:1. Bacteria able to grow on2,4-dinitrotoluene have been studied extensively. Unfortunately, 2,6-dinitrotolueneinhibits the degradation of 2,4-dinitrotoluene and may prevent natural attenuation.Bacteria able to grow on 2,6-dinitrotoluene have been isolated recently, and insightabout the metabolic pathway might allow better prediction of degradation of themixture.

25

Nitroaromatic organic contaminants are associated uniquely with military activ-ities and include the explosives trinitrotoluene (TNT), royal Dutch explosive (RDXor hexahydro-1,3,5-trinitro-1,3,5-triazine), and octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocene (HMX).

25

Manufacturing, loading, storage, and decommissioning oper-ations have generated large quantities of explosive wastes, some of which weredeposited in soils and unlined lagoons and subsequently leached to groundwater.

Despite the number of sites contaminated with explosives, only a few rigorousfield studies have been conducted to determine the transport, fate, and influence ofmicrobial activity on explosives. Furthermore, the field studies carried out to dateare inconclusive in establishing the role of biodegradation in the fate of nitroaro-matics.

25,51

Laboratory studies clearly show the potential for microorganisms tometabolize nitroaromatic compounds.

2,49,51,52

However, microbes apparently cannotreadily use TNT, RDX, or HMX as primary substrates for sources of the carbon andenergy needed for their growth. Instead, cometabolic reactions generally prevail.

1,49

Under aerobic and anaerobic conditions, microorganisms routinely reduce the nitrogroups on nitroaromatics to amino nitro groups. These changes can increase toxicityof the molecules and cause them to form polymers, and/or strongly sorb onto soils.

1,52

Recent reports have shown that aerobically and anaerobically grown bacteria canuse TNT and RDX as nutritional nitrogen sources,

2,53,54

but metabolic byproductaccumulation is common. The possibility of natural attenuation of nitroaromaticscannot be precluded, but the kinds of conditions needed are not clearly understood.



Nitrate Esters:

A variety of nitrate esters including glycerol trinitrate, pen-taerythritol tetranitrate, and nitrocellulose have been used extensively as explosives.Recent studies have indicated that the nitrate esters can be degraded by bacteriafrom a variety of sources.

25,55,56

Bacterial metabolism releases nitrite which can serveas a nitrogen source and yield a selective advantage for the organisms. The biodeg-radation of nitrate esters has only recently been studied extensively and little isknown about degradation in the environment. The recent laboratory results showconsiderable promise that natural attenuation is possible, but more information isneeded on the bioavailability, toxicity, and kinetics of the process.

Pesticides:

Most pesticides used in the past 20 years in the U.S. have beenformulated to degrade in the environment, and a considerable amount of informationis available on degradation kinetics in soil and water. The U.S. EnvironmentalProtection Agency Risk Reduction Engineering Laboratory in Cincinnati, OH hasdeveloped an extensive Pesticide Treatment Database that contains information ona variety of compounds.

25

Many pesticides hydrolyze and yield compounds thatserve as growth substrates or sources of nitrogen or phosphorus for bacteria.Enhanced degradation of pesticides has been studied extensively

25,57

and is closelyrelated to natural attenuation. For example, carbamates,

25,57

chlrophenoxyacetates,

25

dinitrocresol, atrazines,

25

and some organophosphates serve as growth substrates forbacteria and would be good candidates for natural attenuation. A variety of otherpesticides are hydrolyzed by extracellular enzymes derived from soil bacteria butprovide no advantage to the organisms that produce the enzymes. Similarly, someof the organohalogen insecticides can be reductively dehalogenated but provide noadvantage to specific organisms. Their biodegradation rates are proportional to thebiomass and activity in the soil. Other organohalogens, such as lindane, can serveas growth substrates for specific bacteria,

25,58,59

but such bacteria seem not to bewidely distributed (Figure 3.16).

Microbial Transformation of Inorganic Contaminants:

Many research reportshave documented that microorganisms can transform inorganic contaminants.

1

How-ever, unlike organic compounds, which microbes can destroy completely to CO

2

,H

2

O, and other innocuous products, most inorganic contaminants can be changedonly to forms with different solubilities and mobilities. Microbial reactions can leadto precipitation, volatilization, sorption, or solubilization of inorganic compounds.These outcomes can be the direct result of enzymes produced by the microbes, orthey can be the indirect result of microbiological production of materials that alterthe biogeochemical environment.

One nearly universal means by which microorganisms lower concentrations ofinorganic contaminants in water is adsorption to the microorganisms themselves.Adsorption can be caused by electrostatic attraction between the metals and themicrobes or by highly specific scavenging systems that accumulate metals to highconcentration within the cells.

1,60

Although sorption to microbial biomass probablycannot be harvested from the subsurface, which would be required to prevent laterrelease of contaminants, it is not likely to be a major factor in natural attenuation.

Metals:

Microbial effects on metals vary substantially depending on the metalinvolved and the geochemistry of the particular site. The behavior of many toxicmetals depends on the microbially mediated cycling of naturally occurring elements,



especially iron and manganese. The possible fates of chromium and mercury illus-trate the variable effects of microbially mediated reactions on metals.

Chromium:

As with many metals, the effects of microbial transformation onchromium vary with its chemical form (technically, its oxidation state). In ground-water, the predominant form of chromium is the oxidized form, Cr(VI), present aschromate (CrO

42–

) and dichromate (Cr

2

O

72–

) ions. Cr(VI) (known as hexavalentchromium) is toxic and mobile. Reduced chromium, Cr(III), is less toxic and lessmobile because it precipitates as Cr(OH)

3

at groundwater pH values of 4.5 to 10.5.A variety of aerobic and anaerobic microorganisms enzymatically reduce Cr(VI) toCr(III), but the physiological reason for this ability has not been adequately inves-tigated. Among the hypotheses explaining these reduction reactions are detoxifica-tion (to move Cr away from the cells), cometabolism (fortuitous enzymatic

Figure 3.16

Pathways of lindane degradation.

Cl

Cl

Cl

ClCl

Cl

Lindane

Cl Cl

Cl

Cl

γ - 3,4,5,6,- tetrachlorocyclohexane

Cl

ClCl

Cl

Cl

Metabolites



reactions), and the use of Cr(VI) as a respirator electron acceptor. Microbes alsomay cause indirect reduction of Cr(VI) by producing sulfide, Fe(II), and reducedorganic compounds because Cr(VI) reduction occurs spontaneously in the presenceof these substances. Regardless of the mechanism involved, natural attenuation thatrelies on chromium reduction requires environmental conditions that strongly favorthe reduced form of chromium.

Mercury:

Mercury is sometimes present in soils and sediments at contaminatedsites in the form of mercuric ion, Hg(II), elemental mercury, Hg(0), and the biom-agnification-prone organic mercury compounds monomethyl- and dimethylmercury(both of which can accumulate at hazardous levels in the food chain). All microbialtransformations of mercury are detoxification reactions that microbes use to mobilizemercury away from themselves.

1

Most reactions are enzymatic, carried out byaerobes and anaerobes, and involve uptake of Hg(II) followed by reduction of Hg(II)to volatile forms (elemental Hg(0) and methyl- and dimethylmercury) or the forma-tion of highly insoluble precipitates with sulfide. In general, natural attenuation basedon microbial mercury reduction and volatilization seems implausible because thevolatile forms remain mobile, although immobilization as Hg(II) sulfides may bepossible if the electron donors needed to sustain the microbial production of enzymesand the sulfate needed for precipitation are present together.

Nonmetals:

Arsenic is a relatively common toxic groundwater contaminant, dueboth to its use in industry and agriculture and to its natural weathering from rocks.Arsenic can exist in five different valence states: As(-III), As(0), As(II), As(III), andAs(V), where the roman numerals indicate the charge on the arsenic atom. Depend-ing on its valence state and the environment in which it exists, arsenic can be presentas sulfide minerals (e.g., As

2

S

3

), elemental As, arsenite (AsO

2–

), arsenate (AsO

43–

),or various organic forms that include methylated arsenates and trimethyl arsine. Thetwo most common forms of arsenic in natural systems are arsentate As(V) andarsenite As(III). As(V) is less soluble and less toxic than the more soluble As(III)form. The more oxidized arsenate would be expected as the dominant form in aerobicsurface waters, and arsenite may be the dominant form in reduced groundwatersystems. As(V) (arsenate), like phosphate, exists mainly in its deprotonated formsat natural pH levels, and so is readily adsorbed onto the positively charged surfacesof minerals such as Fe(III) oxides. The more toxic aresenite exists primarily as aneutral dissolved species at pHs typical of natural systems, and its transport istherefore not as much retarded by sorption onto oxide surfaces. Half times foroxidation of an arsenite in the presence of Mn(IV) oxides in laboratory experimentshave been measured as 10–20 min, compared to 17 h in natural systems and 8760h for solution of arsenite and dissolved oxygen without Mn oxides. Arsenic speci-ation in natural systems is not consistent with thermodynamic equilibrium and thekinetics of redox conversions of arsenic are relevant to its fate and transport.

Microorganisms can transform arsenic for one of several physiological reasons.Under anaerobic conditions, microbes can use As(V) as a terminal electron acceptor.Under aerobic conditions, oxidation of reduced As (e.g., arsenite) generates energyfor microbes. Under anaerobic and aerobic conditions, microbes transform arsenicby methylation, oxidation, or reduction mechanisms that mobilize it away from



microbial cells. However, microbial transformation of arsenic is not promisingbecause this element can exist in many mobile forms.

Selenium, another nonmetal, is used in a number of commercial and industrialprocesses (including photocopying, steel manufacturing, glass making, and semi-conductor manufacturing) and is sometimes present at contaminated sites. Seleniumcontamination has also resulted from irrigation practices that led to the accumulationof selenium dissolved from soils. Although selenium is an important micronutrientfor plants, animals, humans, and some microorganisms (largely because of its rolein some key amino acids) when present at very low concentrations, it is toxic athigher concentrations. In natural environments, selenium has four inorganic species:Se(VI) (selenate, SeO

42–

), Se(IV) (selenite, SeO

32–

), Se(0) (elemental selenium), andSe(-II) (selenide) and exists primarily as the two soluble species, Se(VI) andSe(IV).

1,61

Like arsenic, selenium also has many volatile organic forms. Reducedinorganic selenium compounds can be oxidized under aerobic conditions, althoughthe oxidation does not support microbial growth. Oxidized selenium (selenate) canserve as a final electron acceptor for anaerobic microorganisms, resulting in produc-tion of selenide and/or elemental Se. Methylation of the various selenium compoundsis a detoxification mechanism that mobilizes Se away from microbial cells, butmethylselenium is mobile and highly toxic to mammals. Anaerobic microbial reduc-tion of selenate and selenite to insoluble elemental selenium can immobilize andremove Se from aqueous solution. Nonetheless, given the complex chemical andbiological processes that influence the fate of selenium and its many mobile forms,microbial reactions are not a promising means for controlling Se contamination.

The speciation of Se in natural systems is dependent on the redox potential, pH,microbial interactions, solubility, complexing ability of soluble and solid ligands,and reaction kinetics. Se(VI) (selenate), the predominant water soluble Se species,mainly occurs in well aerated alkaline soils of higher redox potential, while Se(IV)selenite occurs mostly in natural systems of moderate or reduced redox potential.Although both ions are highly water soluble, the higher adsorption properties ofSe(IV) make it less mobile in the subsurface than (Se(VI). Overall the redox statusappears to be the most predominant controlling factor over Se speciation.

61

Oxyanions:

Oxyanions are water-soluble, negatively charged chemicals in whicha central atom is surrounded by oxygen. Nitrate (NO

3–

) is one such oxyanion. It cancome from natural sources or human sources including nitrogen fertilizers. AlthoughNO

3–

can occur naturally, it is a serious health concern at high concentrations becauseit can cause the respiratory stress disease methemoglobinemia in infants and becauseit can produce cancer-forming nitrosamines.

The major microbial process that destroys nitrate is reduction to nitrogen gas(N

2

) via a process called “denitrification.” Microbes can use nitrate as a terminalelectron acceptor when oxygen is not available. The denitrification process has beenongoing for millions of years and is widespread among microorganisms; it occursreliably in every anaerobic habitat with abundant carbon and electron sources.Natural attenuation by denitrification is possible, as long as the supply rate of anelectron donor is sufficient to sustain the reaction. Many organic compounds, aswell as H

2

and H

2

S, can serve as the electron donor.

1



The oxyanions chlorate (C1O

3–

) and perchlorate (C1O

4–

) or their precursors(chlorine dioxide, hypochlorite, and chlorite) are produced by a variety of papermanufacturing, fertilizer, water disinfection, aerospace, and defense industries.Although not naturally occurring, these highly oxidized forms of chlorine are ener-getically favorable electron acceptors for microorganisms. Knowledge of chlorateand perchlorate biodegradation reactions is quite limited compared to understandingof denitrification.

61

However, laboratory studies using bacterial cultures and envi-ronmental samples (soil, freshwater sediments, and sewage) have shown that micro-organisms can reduce perchlorate and chlorate when supplied with common electrondonors (such as carbohydrates, carboxylic acids, amino acids, H

2

, or H

2

S). Reducingperchlorate and chlorate generates nontoxic chloride iron.

1,61

Microbial transforma-tion of perchlorate or chlorate is plausible if the supply rate of electron donors isadequate.

62

3.5 MONITORING AND SAMPLING FOR NATURAL ATTENUATION

At long last, natural attenuation has come into its own. Over the past five years,great strides have been made in conceptualizing natural attenuation and developingprotocols, field methodologies, guidance documents, and strategies for implemen-tation. However, the most important aspect of a monitored natural attenuation (MNA)evaluation at a site is the need to collect biogeochemical and groundwater qualitydata of the highest quality to predict the natural attenuation capacity of the system.As typically practiced, natural attenuation studies place heavy emphasis on quanti-fying aqueous-phase electron acceptors, contaminants, and byproducts by samplinggroundwater in monitoring wells. In response to this need, a number of companiesoffer multiparameter,

in situ

and down-hole groundwater quality field monitoringdevices that can facilitate the collection of the biogeochemical information.

It can be easily concluded that groundwater samples from zones in which con-taminants are being naturally biodegraded are often in dramatic nonequilibrium withambient conditions. Furthermore, contact of these samples with the atmosphere cancause significant shifts in aqueous biogeochemistry. The key to minimizing oravoiding shifts in the biogeochemistry of reduced samples, in particular, is minimiz-ing contact with atmospheric air. Associated sampling considerations to avoidinclude the following:

• Purging wells at a high rate may lower the water level in the monitoring well.During recharge, there is significant contact between the groundwater and theatmospheric air as the groundwater trickles into the well.

• Use of a bailer for sample collection results in exposure of the sample to theatmospheric air as the sample is poured into the sample bottle.

• Sample holding times, typical with many commercial laboratories, offer the oppor-tunity for changes in the biogeochemistry of the sample.

• Other than samples for volatile organic compounds (VOCs) analysis, groundwatersamples are often collected in such a way that there is headspace in the samplebottle. Agitation of the sample bottle during handling and shipping may result inmixing and thus altering of biogeochemistry.



Since many of the commonly employed groundwater sample collection tech-niques in the past presented varying degrees of contact between the sample and theatmosphere, an inherent concern always exists regarding the reliability and repre-sentativeness of results obtained through these sampling methods for the bio-geochemical parameters of interest. During the past few years, a low-flow, minimalaeration method has evolved which produces the most representative samples forparameters particularly sensitive to artificial aeration and resulting changes in bio-geochemistry. This method involves slow purge rates, a down-hole pump, a flowcell for probe measurements, and a sample-bottle filling procedure that minimizessample aeration (Figures 3.17 and 3.18). Various studies have reported that the low-flow, minimal aeration method using the Grundfos pump produces the most repre-sentative results for most parameters. Minor biases in the results of methane, dis-solved iron (Fe

2+

), and sulfide are possible.Many devices have been developed to collect data

in situ

(down-hole); conse-quently, data errors related to sampling artifacts associated with above-ground datacollection and sequential parameter measurements can be avoided. Furthermore,because many of these units can be coupled to automatic data recorders that providefor immediate data collection and storage in the field and subsequent data transferto personal computers in the laboratory, errors related to data transcriptions can alsobe eliminated.

Figure 3.17

Schematic of minimal aeration, low-flow groundwater sampling technique.

Slow Purge Rateto Minimize WaterTable Drawdown

Water Table

Submersible Pump

Monitoring Well(2"Ø or Greater)

Beaker (for Probe Measurement)or Sample Bottle

Fill to OverflowingWith DischargeEnd of TubeFully Submerged

3-WayValve

Meter

Probe

Flow Cell ProbeMeasurement Device

Valve for AdditionalRegulation of Pump

Discharge Rate

Flexible Tube



Field portable meters capable of measuring (DO) concentrations are availablefrom a variety of manufacturers. These instruments can record DO levels in freshwater or saltwater and most are equipped to make temperature and salinity correc-tions. Oxidation/reduction potential (ORP, Eh, or REDOX) can be difficult to mea-sure even with the best available instrumentation. The sensing device (most often aplatinum electrode in a circuit with a standard reference electrode) may be unstablein fresh waters with low ionic strength. The time required to obtain a stable readingmay be quite long in some cases. Although it is possible to measure REDOX in thefield, considerable operator skill and experience are necessary to obtain accurateresults.

Two types of field measurements for DO and REDOX are possible with thecurrent generation of water quality instrumentation: on-site and in situ. On-site refersto measurements in which a water sample is removed from the aquifer or body ofwater and a sensor immediately placed in it for measurement. Great care is takento isolate the sample from the atmosphere. In situ or “down-hole” sensors refer tomeasurements made by lowering the probe directly into the well or surface waterat the desired depth. After a suitable equilibration time, continuous monitoring ofwater quality can be performed.

Two types of on-site measurements are available: discrete sampling and flow-through sampling. Discrete samples are collected in the appropriate sample container(e.g., 300-mL biological-oxygen-demand (BOD) bottles or other suitable glass-stoppered bottles capable of preventing entrainment of atmospheric oxygen). TheDO or REDOX sensor is then placed in the sample for measurement. Flow-throughcells incorporate the sensor in a cell in line with a pump. DO and/or REDOX andother primary water quality parameters are continuously monitored as the waterflows through the cell. The flow-through technique provides immediate results andminimizes problems resulting from the collection and transport of samples to an on-site laboratory or measurement station.

Figures 3.18 Geochemical consequences due to atmospheric interferences during sampling.

(Eh ) CO2

O2

CH4

Fe2+

Fe(OH)3Precipitate

Atmosphere

Anaerobic, ReducingGroundwater

O2CO2CH4

DOFe2+

CH4Alkalinity

===

0.510 - 50 mg/L 2 - 20 mg/L± 500 mg/L

===

21% 0.03% 0%



Two types of in situ measurements are available: short-term continuous moni-toring and long-term continuous monitoring. Once calibrated, positioned at thedesired depth, and equilibrated to the sample conditions, most probes can sendcontinuous readings to surface instrumentation. Meters may display or log theseresults for a short period of time. A probe designed for long-term monitoringincorporates features to allow it to be anchored in place and operated unattendedfor long periods of time. Long-term monitoring can be useful in evaluating ground-water quality before and during corrective action.

The membrane electrode has been used to monitor DO levels for a long timeduring site characterizations. Dual DO/REDOX measurements can be complemen-tary. If REDOX measurements indicate a negative or reducing environment, thecorresponding DO reading should be low (e.g., <1 mg/L). In situ DO/REDOXmeasurements can also be used to evaluate stratification in an aquifer.

Despite their data-collection virtues, however, it is important to realize that suchunits, especially when they are leased, can collect inaccurate data if certain precau-tions and procedures are not followed prior to and during their use in the field.Therefore, the following sections provide considerations and recommendationsregarding the use of such multiparameter, groundwater quality field equipmentdesigned to maximize the accuracy of biogeochemical data collected. It is veryimportant that the field technician has considerable experience with general fieldmonitoring equipment and procedures to avoid human bias and that he or she hasbeen afforded the opportunity to become familiar with the multiparameter monitor-ing system to be used.

Equipment consideration: Many firms that lease multiparameter monitoring unitswill calibrate the sensing electrodes in the laboratory prior to shipment to the enduser. As a consequence, the remediation engineer is typically told that field calibra-tions of the sensors in the unit are unnecessary. However, there are two considerationswith respect to the as-received accuracy and utility of the monitoring system. First,the previous user may have subjected the equipment to severe groundwater environ-ments, improper handling, and inadequate cleanup. If those conditions are notcompletely corrected by the vendor prior to shipping, erroneous readings can becollected. Second, the monitoring system may have been subjected to harsh treatmentduring shipping such that the calibrated function of one or more of the parametershas been seriously altered or perhaps completely eliminated. As a consequence, itis a good practice to follow the instructions specified to ensure proper systemfunction and to maximize the probability that the key electrode sensors are indeedproviding accurate field measurements.

Initial system checkout: Upon receipt of the monitoring equipment, it should becarefully unpacked and inspected for signs of damage or fouling. Inspect both themeter housing for cracks or blemishes and the electrode sensing unit to verify it hasbeen properly cleaned. Expose the electrodes in the sonde according to the manu-facturer’s directions and verify that each electrode is intact and has been properlypackaged for shipment. Also inspect the cable between the meter and the sonde fordamage or contamination. If any evidence of equipment damage is identified, imme-diately contact the vendor and evaluate the need for a replacement unit.



Otherwise, attach the cable to the meter, place the sonde in a beaker of distilledwater or tap water, and turn on the system. Within moments, the unit typicallyperforms an internal system checkout and the readouts for the different parametersshould become activated. If necessary, scroll through the different parameter readoutsto verify proper function and reasonable readings. If any spurious readings areobserved, one or more of the electrode sensing units may be experiencing problems.Check to make sure the suspect electrode is firmly attached to the sonde. If spuriousreadings are still observed, contact the vendor for technical advice and evaluate theneed to replace the unit.

Response time is the most important quality control check to be made in thefield for DO/REDOX systems. It is used to determine the condition of the electrodes,especially those in water with high contaminant concentrations, dissolved inorganicsalts, sediments, and other insoluble material. Fouling of the sensing electrodes isthe most likely cause of errors in measurement of DO/REDOX.

3.5.1 Dissolved Oxygen (DO)

Electrode inspection and membrane replacement: The user should inspectthe membrane covering the DO electrode in the sonde for any bubbles that mayhave formed beneath the membrane during shipment. Although an unlikely devel-opment with the current generation of DO electrodes, if any bubbles are presentbeneath the membrane, spurious DO readings are likely to result. Consequently, ifany bubbles are observed, the membrane must be replaced using the materials andinstructions provided in the field kit supplied with the multiparameter unit. Once amembrane has been properly replaced, the electrode should be immersed in distilledwater for a minimum of five minutes prior to field use to allow the new membraneto equilibrate. After the membrane has been checked, verify the proper function ofthe DO sensing system by turning on the meter and placing the sonde in a beakerof aerated, room temperature (∼20°C) tap water and observing the DO readout onthe meter. After about two minutes, the DO reading at sea level should be about 9mg/L. Refer to the table provided in Appendix B to determine the appropriate DOreading for the altitude of interest and the current barometric pressure.

Probes designed to detect DO consist of reference and sensing electrodesimmersed in supporting electrolyte and separated from the sample solution by aselective membrane (Figure 3.19). The oxygen sensor consists of the membrane anda closely fitted electrode. The sensing electrode is considered the cathode wheremolecular oxygen is reduced, and the reference electrode is considered the anode.Only species that can permeate the membrane and are reduced at the sensingelectrode will produce a signal, resulting in the highly selective and sensitive natureof the DO sensor. The cell current is linearly proportional to the DO concentrationand can be converted to concentration by simple calibration procedures.

Several types of DO sensor designs are available; the most commonly usedelectrode is the polarographic (commonly called the standard) electrode. This probeutilizes an applied potential to reduce molecular oxygen and requires circulation ofthe water being analyzed. If the water is not moving at about one foot per second,an error in DO concentration will result. The error is caused by the buildup of a



Fig

ure

3.1

9P

rinci

ple

of m

easu

rem

ent

of d

isso

lved

oxy

gen.

e -

(App

lied

Vol

tage

)e

- (S

pont

aneo

us R

eact

ion)

Pol

arog

raph

ic E

lect

rode

Ano

de (

Lead

)

KC

I Ele

ctro

lyte

Cat

hode

(G

old)

Per

mea

ble

Mem

bran

eP

erm

eabl

e M

embr

ane

Cat

hode

(S

ilver

)

Alk

alin

e E

lect

roly

te

Ano

de (

Lead

)

DO

Cel

l of G

alva

nic

Typ

e



concentration gradient in the vicinity of the DO membrane as oxygen is consumedby the sensor. Circulation of the water replenishes the sample near the sensor.

The galvanic cell (Figure 3.19) is less commonly used for detecting DO. Thevoltage detected by this type of probe is produced by the spontaneous reduction ofmolecular oxygen at the cathode (analogous to a fuel cell). Because less oxygen isconsumed from the sample, this cell is less sensitive to low water flow. For both typesof sensors, molecular oxygen is reduced by a noble metal cathode fitted closely to thepermeable membrane. Chemical changes that occur in the electrolyte, and on thesurface of the electrode as a result of the chemical reactions in the cell, will eventuallynecessitate cleaning the electrode surfaces and changing the electrolyte. There is athird type of DO sensor which reportedly consumes no oxygen from the sample.

Field Calibrations: On the day of the site visit, obtain the current site specificbarometric pressure and altitude. Upon arrival at the site, the monitoring unit shouldbe activated and allowed to warm up for at least five minutes before any fieldcalibrations are attempted. For the highest level of accuracy in DO measurements,the multiprobe system should be calibrated at the ambient groundwater temperature.This can be accomplished as follows:

Lower the sonde down a monitoring well or a measuring cell that is uncontam-inated and upgradient from the other monitoring wells of interest. Avoid aeratingthe groundwater surface during sonde passage into the groundwater. Monitor thetemperature function on the meter and determine when the sonde has attained thermalequilibrium with the groundwater. This may take over 90 seconds. Remove the sondeand immediately calibrate the DO electrode using the standard procedure providedin the operation manual. Because the sonde will tend to maintain the ambientgroundwater temperature during calibration, DO calibration will have been essen-tially accomplished at the ambient groundwater temperature. Remember to adjustthe calibration procedure to account for the current barometric pressure and the site’saltitude.

For DO calibrations at groundwater sites where low DO conditions are antici-pated, the meter and sensing unit should be calibrated at two end points. The firstend point, the higher-end calibrated value, is obtained by following the proceduresprovided in the operation manual. The method provides the high-end value becauseit calibrates the instrument under oxygen-saturated conditions.

To obtain the low-end calibration value, three different methods are recom-mended:

• For the first method, add one tablespoon of sodium sulfite (Na2SO3) crystals to abeaker containing 300 ml of tap water. After three minutes, lower the sonde inthe beaker (making sure that the DO electrode is immersed in the solution) andobserve the DO reading. It should descend to 0 mg/L within three minutes. If not,adjust the meter to read 0 mg/L. After calibration, rinse the electrode with distilledwater.

• For the second method, add a cube or packet of baker’s yeast to a beaker containing300 ml of tap water and allow the yeast to dissolve completely. Lower the DOelectrode into the solution and wait three minutes. Again, if the DO reading does



not descend to zero after three minutes, adjust the meter to read 0 mg/L. Aftercalibration, rinse the electrode as before.

• For the third method, bring a small bottle of pressurized nitrogen gas to the field.Attach a small flexible hose to the outlet of the pressure regulator on the nitrogenbottle and slightly open the regulator. This will allow a slow flow of nitrogen gasto exit the tubing. Orient the tubing to within two centimeters of the oxygenelectrode to deliver the nitrogen gas directly to the surface of the oxygen electrodemembrane. After three minutes, the DO reading on the meter should decline to 0mg/L. If it does not, adjust the meter to read 0 mg/L.

Regardless of which method is used to obtain the low-end calibration value, thecalibration should be conducted immediately after the high-end value has beenobtained at the ambient groundwater temperature. If no adjustment to the meter wasrequired to obtain the low-end calibration value, the DO function of the meter isnow calibrated. If a meter adjustment was required for the low-end reading, imme-diately repeat the high-end calibration step under ambient groundwater temperatureconditions. The DO function of the meter should now be properly calibrated.

Field Measurements: The sonde should be slowly lowered into the groundwaterof the well until the DO and other parameter sensing devices are immersed in thegroundwater. Again, avoid aerating the groundwater during sonde passage into thegroundwater. The sonde’s electrodes should be placed at either: 1) the midpoint ofthe well screen or, 2) in cases where the water column in the well is shorter thanthe length of the well screen, the midpoint of the available water column in the well.The appropriate depth should be determined by knowing the depths of the upperand lower ends of the well screen and by obtaining the depth to groundwater in thewell on the day of sampling. The DO measurement should be recorded when thereading has stabilized, generally within two minutes.

During the collection of field measurements in multiple groundwater monitoringwells, the DO membrane should be checked after each well monitoring event forbubbles or organic fouling due to the presence of nonaqueous phase liquids (NAPLs)or microbial “slimes” in the groundwater in the monitoring well. If bubbles areencountered or the membrane is severely fouled with an organic film or microbialslime, the membrane should be replaced prior to further use.

If only light organic fouling is observed, however, the membrane may be cleanedby immersing the sonde in a dilute detergent solution and “swishing” the electrodearound in the solution. For electrodes lightly fouled with microbial slimes, theelectrodes can be disinfected using a 10% solution of denatured alcohol in distilledwater. Once it has been verified that the membrane has been cleaned, it should berinsed with distilled water.

Regardless of the presence or absence of organic films, however, for mostgroundwater monitoring circumstances, it is good practice to decontaminate thesonde and any cabling exposed to groundwater. Decontaminate using appropriate,standard decontamination procedures prior to the using the equipment in each sub-sequent wells.



3.5.2 Oxidation–Reduction (REDOX) Potential (ORP)

The oxidation-reduction potential of groundwater is a measure of electron activ-ity and an indicator of the relative tendency of a biogeochemical system to acceptor transfer electrons. In a very oxidizing environment the activity of electrons is lowand in a very reducing environment the activity of electrons is high. The electronactivity is characterized by using the lowercase “p” notation (just as in pH).

pe = – log [e–] (3.19)

The pe of a surface water at pH 7, in equilibrium with atmospheric oxygen, iscalculated to be 13.6; it decreases to approximately 4 in an environment where Fe3+

reduction takes place and drops to approximately –4 where sulfate reduction andmethanogenic conditions exist. pe can be calculated from the measured concentra-tions of reactants and products in a REDOX half-reaction (Figure 3.20). A scaleequivalent to pe is the Eh scale, which is expressed in volts and is based on thedetermination of electron activity using electrochemical methods. At temperaturestypical of the natural environments:

(3.20)

where Eh is in volts.Eh is often confused with the closely related REDOX potential or ORP measure-

ment. ORP or REDOX is measured by placing a REDOX electrode into the watersample; the REDOX electrode is a piece of metallic platinum, which acquires amore negative potential with respect to its reference electrode under reducing con-ditions where electron activities are higher (Figure 3.21). ORP is the voltage mea-sured between this REDOX electrode and the reference electrode placed in the sameenvironment. If the activities of the oxidizing species are greater than those of thereducing species, a voltage greater than the reference electrode voltage will result.Although REDOX is temperature dependent, temperature corrections are rarelyperformed because of the lack of theoretical knowledge involving the exact natureof the active species in the sample.

ORP provides a useful, approximate characterization of REDOX conditions in theaquatic invironment, although it lacks precise theoretical definition. Although ORPand Eh are both measured in volts (millivolts) and do show some rough correlation,they are defined quite differently and should not be treated as synonymous.

Electrode inspection: The surface of the oxidation-reduction potential (ORP)electrode should be inspected and it should be verified that the surface is clean ofany organic or inorganic films. If it appears to be fouled, it can be gently cleanedusing a slightly abrasive cloth dipped in a mild detergent solution as describedpreviously. After cleaning, it should be rinsed with distilled water.

Epe

h =16 7.



Laboratory reference checks: ORP sensing systems tend to be quite robustwith respect to maintaining calibration and are typically calibrated by the vendorprior to shipment to the user. The ORP electrode tends to respond to a linearfashion over a large range of ORP conditions at groundwater sites. However, wherestrongly negative ORP readings are possible, it may be advisable to verify theproper operation of the ORP function by conducting high-end and low-end ORPreference checks. In those cases, the proper operation of ORP electrode systemscan be evaluated upon receipt of the monitoring unit using the following two-pointreference check method:

1. For high-end reference value, obtain the “ZoBell” solution from the vendor orprepare 125 ml of the solution using the method supplied in many chemical textssuch as Standard Methods. Regardless of whether you purchase or prepare thesolution, make sure you obtain a copy of the ORP temperature table from thevendor. Immerse the sonde’s ORP electrode in the solution and observe the ORPreading on the meter. At 25°C, the ORP reading should be +237mV.

Figure 3.20 The REDOX sequence of different electron acceptor reactions catalyzed bymicroorganisms.

Approximate range due in part to variability in reaction/productconcentrations

Aer

obic

Res

pira

tion

Oxidant in Use

-5

2O NO3 MnO 2 Fe(OH)3 SO4 CO22--

0

5

10

15900

600

300

0

-300

hE

(m

V)

ρ ε

Den

itrifi

catio

n

Man

gane

se R

educ

tion

Iron

Red

uctio

n

Sul

fate

Red

uctio

n

Met

hano

gene

sis



2. At low-end reference value evaluation can be obtained by using the sodium sulfitesolution prepared for the low-end DO calibration method described previously.Once the sonde has been immersed in the sodium sulfite solution, the ORP valueshould be negative.

If the value observed at the high-end check is considerably different from theexpected value or if the value obtained during the low-end evaluation is positive,the ORP function may be damaged and the vendor should be contacted for furtherguidance or perhaps for meter and/or sonde replacement.

Field measurement: Once the sonde has been lowered to the appropriate depthin the monitoring well, the ORP measurement can be recorded after it has stabilized,typically after 90 seconds.

3.5.3 pH

The acidity of a given sample is determined by the concentration of hydrogenions present. This concentration is expressed by the pH. The value of pH is usuallyexpressed by the equation:

pH = – log [H+] (3.21)

Figure 3.21 Principle of measurement REDOX.



Although this equation is written in terms of concentration, in fact the quantitiesin the brackets should be activity, which may be thought of as “corrected” concen-trations that take into account nonideal effects in aqueous systems. These nonidealeffects arise from electrostatic forces between ions dissolved in the water, and theyincrease as the total concentration of dissolved ions, measured as ionic strengthincreases.

Electrode Inspection: The pH electrode should be inspected and it should beverified that any organic or inorganic films are absent. If the electrode appears tobe fouled, the sensing region of the electrode can be gently cleaned using a non-abrasive cloth dipped in a mild detergent solution as described previously. Aftercleaning, the electrode should be rinsed with distilled water.

Laboratory Calibrations: There are several standards available for calibratinga pH electrode. A three-point calibration should be performed using standardsanticipated to bracket the range of pH levels expected to be observed in the field.For example, for most groundwater situations, calibrations using standard solutionsof pH = 4, 7, and 10 would be appropriate.

Calibrate the pH function using the pH = 7 standard first. At sites where alkalineconditions are anticipated, calibrate the pH electrode with the pH = 10 standard last.At sites where acidic conditions are anticipated, calibrate the electrode with the pH= 4 standard last. Once the calibrations have been completed, rinse the pH electrodewith distilled water. The proper operation and calibration of the pH function hasnow been verified.

Field Measurements: pH measurements can be recorded once the pH readinghas been stabilized, typically within 60 seconds.

3.5.4 Filtered vs. Unfiltered Samples for Metals

One of the most hotly contested protocols in the last few years is whether tofieldfilter samples that are to be analyzed for dissolved metals. While partisanscommonly make the case either for always filtering or always not filtering, twoessential factors lead inevitably to the conclusion that filtering is technically appro-priate in most cases and not in others:

• Metal concentration data may be used for many purposes, and the data use dictateswhether filtering is appropriate. For example, if direct ingestion from a drinking-water source is involved, data from an unfiltered sample is appropriate. Alterna-tively, for assessment of contaminant removal during a remediation project, filteredsamples may be appropriate.

• Different hydrogeologic conditions may cause groundwater samples to be turbid,no matter what well installation, development, purging, or sampling methods areused. As indicated above, turbid water may give metal concentrations that are notindicative of the concentrations actually moving in the aquifer and would be biasedhigh relative to actual dissolved metals transport.



3.5.4.1 Field Filtration and the Nature of Groundwater Particulates

Monitoring wells are purged and samples are brought to the surface by pumpsof many types or by bailer. The types of sampling devices and their advantagesand disadvantages are discussed in numerous research papers and regulatoryguidance manuals. For a given well sampling event, one device may be used forpurging stagnant water from the well, while another may be used to take thesample.

Filtration is most often conducted on samples to be analyzed for metals content.The distinction is commonly made between a sample to be analyzed for total metalscontent (an unfiltered sample) and a sample to be analyzed for dissolved metalscontent (a filtered sample). As discussed later, the presence of very small particlescalled colloids adds substantial complexity to the discussion of total vs. dissolvedmetals content. Many of these colloidal particles typically will not settle out due togravitational forces.

Many types of filters are available with regard to both the type of filter apparatusand the pore size of the filter. Filtration can occur in the open: a water sample ispoured into a filter funnel and the sample is pulled by vacuum through a membrane.Alternatively, the filtration can occur in a closed system where the sample is pushedby the sampling pump through its discharge tube and finally through an “in-line”filter unit. In either the open or closed system, the filtrate, once through the filter,is directed into the sample bottle for preservation (usually with nitric acid for metals)and shipment for analysis.

A wide range of filter pore sizes is available. Since the late 1970s, the moststandard nominal pore diameter used in the environmental business has been 0.45micrometers (µm). Other common filters that have been used in environmentalresearch and for special applications have pore diameters ranging from 5 µm at thehigh end to 0.1 and 0.03 µm at the low end.

Historically, the groundwater system was viewed as having two parts: the immo-bile rock/soil phase and the mobile water phase. However, in the recent past, empha-sis has been placed on another partially mobile phase — colloids — in evaluatingcontaminant transport. Colloids are very small organic or inorganic particles thatcan range from less than 0.1 to 10 µm in diameter. The truly dissolved phase hasmolecules or polymers that are substantially smaller than 0.1 µm. Colloid compo-sition, charge, and aquifer conditions vary considerably through space and time, butresearch has shown that particles up to about 2 µm in diameter can move withgroundwater.63 Particles at the higher end of the colloid size range may be trappedby small pore spaces in the soil matrix or settle out due to gravity.

Colloidal particles may be present naturally in groundwater or may be releasedfrom soil or rock surfaces during well installation or sampling or may be depositedat the bottom of the well due to precipitation of the dissolved metals in the wellvolume itself. Even the action of a bailer in a well can substantially increase theconcentration of colloids in a groundwater sample. If the objective of an investi-gation is to assess the movement of metals in the subsurface, field filtration will



remove some intermediate-to-large size colloids that are not mobile under naturalconditions, but have been mobilized near a well due to well installation or samplingdisturbance.

Of special importance is that the standard environmental filter has a pore size(0.45 µm) that is in the middle of the size range where colloids can be expected tomove in the subsurface. This filter does not provide a clean cutoff between dissolvedand colloidal species that are mobile and colloids and larger particles that are not.As a matter of fact, no single filter pore size could provide the correct cutoff betweenmobile and immobile species because aquifer, well, and sampling conditions andcircumstances vary widely. Short of making an intensive research effort for eachproject site, until recently there has been no way to accurately distinguish mobilefrom immobile particulates. As discussed later, recent research and experience withlow-flow sampling has provided a means, under some circumstances, of retrievingsamples with only dissolved, naturally occurring, and mobile colloids.64,65

Before the installation of numerous monitoring wells at contaminated sites duringthe past 20 years, the thinking about groundwater samples was shaped largely byour experiences with production wells. These wells were installed in “true” aquifersthat generally, in the case of unconsolidated formations, were composed of relativelycoarse-grained, well-sorted materials with high hydraulic conductivities. These wellscould be readily developed to yield water with low turbidity, so filtration was notusually needed to obtain a water sample considered representative of formationwater. In addition, filtration to remove any particles was not desirable because waterfrom such a well was expected to be directly consumed.

Filtration can be considered undesirable because the extra handling in the fieldcan alter the sample’s chemistry. Aside from the intended removal of particulateslarger than the pore size of the filter, dissolved or colloidal chemicals in the samplecan adsorb onto the filter membrane or apparatus and the filter pore size may bealtered by particle clogging. Alternatively, ions associated with the filtration systemcan leach into the sample, thereby giving false positive results. Finally, the exposureof a sample to air during filtration can cause metals such as iron to oxidize andprecipitate. These iron precipitates may be stopped by the filter, thereby loweringthe iron concentration in the sample filtrate. The precipitation may also entrain othermetals, resulting again in lowered concentrations in the sample.

3.5.4.2 Reasons for Field Filtration

Many monitored sites are located over aquifers with low hydraulic conductivities.By definition, an aquifer is a formation that contains sufficient saturated permeablematerial to yield sufficient quantities of water to wells and springs. Although theconcepts “low hydraulic conductivity” and “aquifer” seem mutually exclusive, as aregulatory matter, the “uppermost aquifer” is frequently the most contaminated andrequires intense monitoring. Low hydraulic conductivity formations screened bymonitoring wells may be uniformly fine grained, such as in a silty clay unit. Alter-natively, a lower permeable unit to be monitored may be characterized by a widearray of grain sizes; a common example of such poorly sorted material is glacial



till. In either case, fine-grained materials of clay or colloid size (particle diameterof less than 2 µm) will be present.

Monitored formations with low hydraulic conductivities present numerous chal-lenges with respect to an excess of colloids being present in a groundwater sample.First, such fine-grained formations intrinsically contain numerous small particles.Second, because of the difficulty of making water flow through the formation,removal of colloidal particles in the zone around the well through well developmentis difficult. The colloid load present in the sample can still be larger than what is inthe formation, even after a substantial well development effort has been made andthe well has been sampled numerous times. In contrast, production wells are devel-oped continuously as large volumes of water are pumped for long periods of time.Finally, in low permeability materials, water levels are substantially drawn down tothe point that a well may have little or no remaining water in it, even at very lowpumping rates. The drawdown causes a high groundwater gradient near the well,with the attendant increase in groundwater flow rate in individual pore spaces. Theaction of purging and sampling may easily mobilize particles that would not normallybe moving with the groundwater. It should be noted that the expected shearing ofcolloidal particles from large soil particle surfaces is proportional to the square ofthe groundwater pore velocity.

In many geologic units, large amounts of iron oxyhydroxides are present. If theREDOX conditions are oxidizing, these oxyhydroxides are in the low solubilityferric form and are present as coatings on the surfaces of large, immobile aquifersolids like sand and grains. At low or near neutral pH values, these ferric oxyhy-droxide coatings are positively charged and can therefore attract and pull colloidalclay particles from the groundwater solution. However, if substantial amounts ofnatural or contaminant organic matter are present, the REDOX conditions can bereducing, and the ferric oxyhydroxides convert to the more soluble ferrous form,which enters the groundwater phase. This combination of soluble ferrous ions andassociated colloidal clay increases groundwater turbidity.

Much of the regulatory literature creates the expectation that, regardless ofgeology, if sufficient care is taken, a well can almost always be designed, installed,and developed so that it produces water with low colloid content, and therefore, lowturbidity. There is also the expectation that well purging will be conducted so thatthe water sample will have low turbidity. The most common criterion for acceptablewell construction and development is a well that yields samples with turbidity ofless than five nephelometric turbidity units (NTUs). However, some geologic matri-ces will yield turbid water to a monitoring well, no matter how proficient the wellinstallation and development. As discussed more fully in the section on low-flow,the only recourse in such formations is either to purge and sample a well at a ratethat does not exceed the natural recharge to the well or to use a bailer, which hasits disadvantages.

As a practical matter in tight formations, however, purging and sampling ratescannot be lowered to the level of natural groundwater flow. Substantial drawdownand the attendant disturbance to the formation will occur even at the lowest pumpingrate achievable with current technology. Therefore, if low-flow rate sampling is not



possible, the next best procedure is to sample at a higher flow rate followed bysample filtration.

Hydrogeologic and biogeochemical conditions can vary through time at a mon-itoring location. In addition, sampling method and operator technique can changeduring the history of a monitoring program. As a result, the turbidity and metalconcentrations in a series of samples can vary above and beyond the real changesof concentration in the aquifer. At wells where it is difficult to control the turbidityand colloid concentration in samples, filtration will help to level out the variations.Frequently, the ability to track changes in concentration through time is an importantobjective of any remediation project. More consistent data will increase the powerof statistical analyses such as trend evaluations.

3.5.5 Low-Flow Sampling as a Paradigm for Filtration

The desire to disturb the aquifer as little as possible has led to research and therapid acceptance of minimum aeration, low-flow sampling of wells. Research in thelate 1980s and 1990s has shown that, in many geologic formations, groundwatermoves horizontally through the screened portion of the well and interacts little withthe water standing in the well above the screened zone.64 If the sample is pumpedfrom the well at a rate that is less than or equal to the natural flow rate through thescreen, no stagnant water from above the screen zone will be present in the sample.In contrast to traditional well purging techniques with pumping rates of 4 liters perminute or more, low-flow purging and sampling occurs typically in the range of 0.1to 1 L/min. Current pump technology allows pumping rates down to about 0.1 L/min;for wells that yield less than this amount, the low-flow methodology is not asapplicable.

The insertion of a bailer or sampling pump into a well causes substantial dis-turbance to the water in and around the well. Bailing is particularly troublesome inthis regard; comparison studies show that bailed samples have much higher turbidityand order of magnitude or more higher-metals concentrations in samples collectedvia low-flow pumping rates. As a result, it is essential that low-flow equipment isused during sampling of MNA parameters.

Low-flow sampling with dedicated pumping equipment has proven to be animportant advance in groundwater sampling. Research and practice indicate thatlow-flow samples have lower turbidity and are more representative of formationwater than samples obtained at higher flow rates with bailers or pumps. The lack ofparticulates has been demonstrated by showing little change in analyte concentrationwhen low-flow samples are filtered with membranes ranging in pore size from 0.03to 5 µm in one study65 and 0.1 to 10µm in another.64

In conclusion, filtered samples, even those taken with a bailer, have similaranalyte concentrations to unfiltered samples taken by the low-flow methodology,which is state of the art with respect to collecting samples considered representativeof aquifer conditions.



3.5.6 A Comparison Study

During the mid-to late 1990s, filtered and unfiltered pairs of samples werecollected at two study sites by ARCADIS Geraghty & Miller, Inc. The wells andtemporary well points were mostly installed in the shallow, unconsolidated zone,characterized chiefly by fill (cinders and a variety of other natural and artificialmaterials), till, and meadow mat (high natural organic matter). Lithologic logs formany locations describe the presence of fine-grained materials. REDOX measure-ments were taken, and because of the presence of dissolved organic matter (bothnatural and contaminants), conditions were reducing at a large number of the wellsin portions of both sites. The presence of fine-grained materials and low REDOXvalues enhanced the formation of colloids in groundwater.

As part of this study, ARCADIS Geraghty & Miller, Inc. reviewed the 1266 pairs(i.e., each location multiplied by the number of analytes) of filtered and unfilteredgroundwater sample results. The ratios of filtered (dissolved) to unfiltered (total)concentrations varied widely from the dissolved concentrations being less than 1%of the total concentration to cases where the dissolved concentrations were greaterthan the total concentration.

Of special note is that dissolved concentrations of only certain metals werefrequently low (10% or less) compared to total concentrations. Aluminum, iron,copper, chromium, lead, vanadium, and zinc fell into this category. This studyshowed that filtered, bailer or high-flow samples give results most similar to theunfiltered low-flow samples, which are presumably most representative of undis-turbed formation water.

A detailed evaluation of the analytical data and geologic conditions revealed thatsamples with lower values of dissolved concentrations had the following characteristics:

• Higher organic matter concentrations• Low REDOX potentials and reducing conditions• Fine-grained particles in the lithology• Very low yield to the point where they can be pumped or bailed dry

As discussed in detail, filtration has both good and bad aspects. The improvedaccuracy and consistency derived from the removal of the large, normally immobileparticles liberated during well installation and sampling outweighs the loss of someof the smaller particles yielding better overall results.

An important argument for filtration is that the sample results from location tolocation and sampling event to sampling event will be more consistent with filtering.Turbidity and colloid load will vary between samples; by reducing the importanceof this variable through filtration, spatial and temporal relationships will emergemore quickly. It is these relationships that help to determine the location of a sourcearea, the effectiveness of a remedial action, or the behavior of dissolved constituentsover time.

Perhaps the most powerful argument for filtration arises from the results of low-flow sampling. When wells are pumped at no more than the rate of natural recharge



to the well, analyte concentrations in samples are considered to be the most repre-sentative of formation water quality. In cases where low-flow pumping is not prac-tical, the closest results are achieved by filtering samples obtained by bailer or high-flow pumps. Unfiltered samples obtained by these methods would be expected tohave a substantially high bias.

In selected cases where there are data collection objectives different from generalsite characterization and remediation planning, there may be reason to collect unfil-tered samples as well as filtered samples. Furthermore, dedicated, low-flow samplingequipment may be considered at specific locations if frequent sampling is indicated,well yield is adequate, and there is no expectation that data from these wells willneed to be compared to data from wells without such equipment.

REFERENCES

1. National Research Council, Natural Attenuation for Groundwater Remediation,National Academy Press, Washington, D.C., 2000.

2. United States Environmental Protection Agency, Use of Monitored Natural Attenu-ation at Superfund, RCRA Corrective Action and Underground Storage Tank Sites,Directive Number 9200.4-17P, Office of Solid Waste and Emergency Response,Washington, D.C., 1999.

3. American Society for Testing and Materials (ASTM), Standard Guide for Remedia-tion of Groundwater by Natural Attenuation at Petroleum Release Sites; Draft, Phil-adelphia, 1997.

4. Wiedermeier, T. et al., Technical Protocol for Implementing Intrinsic Remediationwith Long Term Monitoring for Natural Attenuation of Fuel Contamination Dissolvedin Groundwater, Vols. 1 and 2, Air Force Center for Environmental Excellence,Brooks Air Force Base, San Antonio, TX, 1995.

5. U.S. Army Science Board, Infrastructure and Environment Issue Group, Remediationof Contaminated Army Sites: Utility of Natural Attenuation, Draft Report, Departmentof the Army, Washington, D.C., 1995.

6. Department of Energy (DOE), Final Programmatic Environmental Impact Statementfor the Uranium Mill Tailings Remedial Action Ground Water Project, DOE/E15-0198, NTIS, Springfield, VA, 1998.

7. Chapelle, F. H. and Paul M. Bradley, Selecting remediation goals by assessing thenatural attenuation capacity of groundwater systems, Bioremediation J., 2, 227–238,1998.

8. Potter, S., personal communication, 1999.9. Wiedemeier, T. H. et al., Natural Attenuation of Fuels and Chlorinated Solvents in

the Subsurface, John Wiley & Sons, Inc., New York, 1999.9a. Fetter, C. W., Contaminant Hydrogeology, Macmillan Publishing, New York, 1993.10. Hemond, H. F. and E. J. Fechner, Chemical Fate and Transport in the Environment,

Academic Press, New York, 1994.11. Pankow, J. F. and J. A. Cherry, Dense Chlorinated Solvents and Other DNAPLs in

Groundwater, Waterloo Press, Portland, OR, 1996.12. Kram, M. L. et al., Comparison of DNAPL characterization approaches, in Treating

Dense Non Aqueous Phase Liquids (DNAPLs), Int. Conf. Remed. Chlorinated andRecalcitrant Compds., Monterey, CA, May 22–25, 2000.



13. Perry, J. J., Microbial cooxidations involving hydrocarbons, Microbiol. Rev., March1979, 59–72.

14. Alexander, M., Biodegradation and Bioremediation, 2nd ed., Academic Press, NewYork, 1994.

15. Buscheck, T. E. and G. M. Alcantar, Regression techniques and analytical solutionsto demonstrate intrinsic bioremediation, In Situ and On Site Bioremed. Symp. Proc.,San Diego, CA, 1995, 109–116.

16. Borden, R. C. and Bedient, P. B., Transport of dissolved hydrocarbons influenced byoxygen limited biodegradation: theoretical development, Water Resour. Res., 22, No.13, 1973–1982, 1996.

17. Kennedy, L. G. et al., Iron and sulfur mineral analysis methods for natural attenuationassessments, Bioremed. J., Vol. 2, No. 3, 259–276, 1998.

18. Baedecker, M. J. et al., The fate and effects of crude oil in a shallow aquifer: thedistribution of chemical species and geochemical facios, US Geo. Surv. Water Resour.Div. Rep. 88–4220, USGS, Reston, Virginia, 1989.

19. Baedecker, M. J. et al., Crude oil in a shallow sand and gravel aquifer: biogeochemicalreackons and mass balance modeling in anoxic groundwater, Appl. Geochem., 8,569–586, 1993.

20. Lovley, D. R. et al., Oxidation of aromatic contaminants coupled to microbial ironreduction, Nature, 339, 297–299, 1989.

21. Chappelle, F. H. et al., Measuring rates of biodegradation in a petroleum hydrocarboncontaminated aquifer, Groundwater, 34, 691–698, 1996.

22. Reinhard, M. et al., In situ BTEX biotransformation under enhanced nitrate andsulfate reducing conditions, Environ. Sci. Technol., 31, 28–36, 1997.

23. Kirtland, B. C., C. Marjorie Aelion, and P. A. Stone, Monitoring anaerobic naturalattenuation of petroleum using a novel in situ respiration method in low-permeabilitysediments, Bioremed. J., 4, 181–187, 2000.

24. Weiner, J. M., T. S. Lauck, and D. R. Lovely, Enhanced anaerobic benzene degradationwith the addition of sulfate, Bioremed. J., 4, 159–174, 1998.

25. Spain, J., Organic chemicals with potential for natural attenuation, Bioremed. J., 1,1–10, 1997.

26. McNally, D. L. et al., Biodegradation of three and four ring polycyclic aromatichydrocarbons under aerobic and denitrifying conditions, Environ. Sci. Technol., 32,2633–2639, 1998.

27. Mihelcic, J. R. and R. G. Luthy, Degradation of polycyclic aromatic hydrocarboncompounds under various redox conditions in soil-water systems, J. Appl. Environ.Microbiol., 54, 1182–1187, 1988.

28. Coates, J. D. et al., Anaerobic degradation of polycyclic aromatic hydrocarbons andalkanes in petroleum contaminated marine harbor sediments, J. Appl. Environ. Micro-biol., 63, 3589–3593, 1997.

29. Zhang, X. and L. Y. Young, Carboxylation as an initial reaction in the anaerobicmetabolism of naphthalene and phenanthrene by sulfidogenic bacteria, J. Appl. Envi-ron. Microbiol., 63, 4759–4764, 1997.

30. Wilson, J. T., J. S. Cho, B. H. Wilson, and J. A. Vardy, National Attenuation of MtBEin the Subsurface Under Methnogenic Conditions, USEPA/600/R-00/006, Washing-ton, D.C., 2000.

31. Salanitro, J. P. et al., Field scale demonstration of enhanced MtBE bioremediationthrough aquifer bioaugmentation and oxygenation, Environ. Sci. Technol., 34,4152–4162, 2000.



32. Bradley, P. M. et al., Aerobic mineralization of MtBE and TBA by stream bedsediment microorganisms,

Environ. Sci. Technol.

, 33, 1877–1879, 1999.33. Borden, R. et al., Intrinsic biodegradation of MtBE and BTEX in a gasoline contam-

inated aquifer,

Water Resour. Res.

, 33, 1105–1115, 1997.34. Schirmer, M. and J. Barker, A study of long term MtBE attenuation in the Borden

aquifer, Ontario, Canada,

Groundwater Monitor. Remed.

, 113–122, Spring, 1998.35. Kolhatkar, R., J. T. Wilson, and L. E. Dunlap, Evaluating Natural Biodegradation of

MtBE at Multiple UST Sites, paper presented at the API Petroleum Conf., Anaheim,CA, November, 2000.

36. Bouner, E. J., B. E. Rittmann, and P. L. McCarty, Anaerobic degradation of haloge-nated 1- and 2-carbon organic compounds,


, 15, 596–599, 1981.37. Ramanand, K. et al., Reductive dehalogenation of chlorinated benzenes and toluenes

under methanogenic conditions,

J. Appl. Environ. Microbiol.

, 59, 3266–3272, 1993.38. Reineke, W. and H. J. Knackmuss, Microbial metabolism of haloaromatics: isolation

and properties of a chlorobenzene degrading bacterium,


,47, 395–402, 1984.

39. Nishino, S. F., J. C. Spain, and C. A. Pettigrew, Biodegradation of chlorobenzene byindigenous bacteria,

Environ. Toxicol. Chem.

, 13, 871–877, 1994.40. Cozza, C. L. and S. L. Woods, Reductive dechlorination pathways for substituted

benzenes: A correlation with electronic properties,

Biodegradation

, 2, 265–278, 1992.41. Halden, R. U. and D. F. Dwyer, Biodegradation of dioxin related compounds, a review,

Bioremed. J.

, 1, 11–25, 1997.42. Mohn, W. W. and J. W. Tiedje, Microbial reductive dehalogenation,

Microbiol. Rev.

,56, 482–507, 1992.

43. Haggblom, M. M. and R. J. Valo, Bioremediataion of chlorophenol wastes, in

Micro-bial Transformation and Degradation of Toxic Organic Chemicals

, Young, L. Y. andCerniglia, C. E., Eds., Wiley-Liss, New York, 1995, 435–486.

44. Suflita, J. M. and G. T. Townsend, The microbial ecology and physiology of aryldehalogenation reactions and implications for bioremediation, in

Microbial Transfor-mation and Degradation of Toxic Organic Chemicals

, Young, L.Y. and Cerniglia, C.E., Eds., Wiley-Liss, New York, 1995, 243–268.

45. Luthy, R. G. et al., Aqueous solubility of PCB congeners from an Aroclor and anAroclor – hydraulic oil mixture,

Water Res.

, 31, 561–573, 1997.46. Bedard, D. L. and J. F. Quensen, Microbial reductive dechlorination polychlorinated

biphenyls, in

Microbial Transformation and Degradation of Toxic Organics

, Young,L. Y. and Cerniglia, C. E., Eds., Wiley-Liss, New York, 1995, 127–216.

47. Bedard, D. L. et al., Stimulation of microbial para-dechlorination of polychlorinatedbiphenyls that have persisted in Housatonic River sediment for decades,

Environ. Sci.Technol.

, 30, 687–694, 1996.48. Flanagan, W. P. and R. J. May, Metabolite detection as evidence for naturally occur-

ring aerobic PCB biodegradation in Hudson River sediments,


,27, 2207–2212, 1993.

49. Spain, J. C., Biodegradation of nitroaromatic compounds,

Ann. Rev. Microbiol.

, 49,523–555, 1995.

50. Spain, J. C., Ed.,

Biodegradation of Nitraromatic Compounds

, Plenum PublishingCorp., New York, 1995.

51. Bradley, P. M. et al., Potential for intrinsic bioremediation of a DNT-contaminatedaquifer,

Ground Water

, 35, 12–17, 1997.52. Funk, S. B. et al., Full scale anaerobic bioremediation of trinitrotoluene (TNT)

contaminated soil,

Appl. Biochem. Biotechnol.

, 51–52, 625–633, 1995.

L1282_C03-B_frame Page 128 Tuesday, June 19, 2001 1:01 PM


53. Binks, P. R. et al., Degradation of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) by

Stenotrophomonas,

Int. Conf. Remed. Chlorinated and Recalcitrant Compds.,Monterey, CA, May 22–25, 2000.

54. French, C. E. et al., Aerobic degradation of 2,4,6-trinitrotoluene by

Enterobactercloacae

PB2 and by pentaerythritol tetranitrate reductase,

J. Appl. Environ. Micro-biol.

, 64, 2864–2868, 1998.55. White, G. F. and J. R. Snape, Microbial cleavage of nitrate esters: defusing the

environment,

J. Gen. Microbiol.

, 139, 1947–1957, 1993.56. White, G. F. et al., Biodegradation of glycerol trinitrate and pentaerythritol tetranitrate

by

Agrobacterium radiobacter

,


, 62, 637–642, 1996.57. Topp, E. et al., Isolation and characterization of an N-methylcarbamate insecticide

degrading methylotrophic bacterium,


, 59, 3339–3349,1993.

58. Siddartha, K. S. et al., Degradation of alpha-, beta-, and gamma- hexachlorocyclo-hexane by a soil bacterium under anaerobic conditions,


,56, 3620–3622, 1990.

59. Cookson, J. T.,

Bioremediation Engineering: Design and Application

, McGraw-Hill,New York, 1994.

60. Chen, S. and D. B. Wilson, Genetic engineering of bacteria and their potential forHg

2+

bioremediation,

Biodegradation

, 8, 97–103, 1997.61. Dungan, R. S. and W. T. Frankenberger, Microbial transformations of selenium and

the bioremediation of seleniferous environments,

Bioremed. J.

, 3, 171–188, 1999.62. Logan, B. E., A review of chlorate and perchlorate respiring microorganisms,

Bioremed. J.

, 2, 69–80, 1998.63. Puls, R. W. and R. M. Powell, Transport of inorganic colloids through natural aquifer

material: Implicaitons for contaminant transport,


, 26, 614–621,1992.

64. Puls, R. W. and M. J. Barcelona, Groundwater Sampling for Metals Analyses,EPA/540/4-89/001, VSEPA, 1989.

65. Puls, R. W. and R. M. Powell, Acquisition of representative groundwater qualitysamples for metals,

Groundwater Monitor. Rev.

, 167–176, Summer, 1992.

CREDIT

Materials from pages 100 through 104 contain substantial excerpts from Spain,J., Synthetic chemicals with potential for natural attenuation,

Bioremediation Jour-nal

, 1(1): 1–9, 1997.

L1282_C03-B_frame Page 129 Thursday, March 6, 2003 11:59 AM


131

CHAPTER

4

In Situ

Reactive Zones

CONTENTS

4.1 Introduction ..................................................................................................1324.2 Engineered Anaerobic Systems ...................................................................135

4.2.1 Enhanced Reductive Dechlorination (ERD) Systems .....................1354.2.1.1 Early Evidence..................................................................135

4.2.1.1.1 Biostimulation vs. Bioaugmentation...............1364.2.1.2 Mechanisms of Reductive Dechlorination

.....................138

4.2.1.3 Microbiology of Reductive Dechlorination

...................142

4.2.1.3.1 Cometabolic Dechlorination

.........................142

4.2.1.3.2 Dechlorination by Halorespiring Microorganisms

..........................................144

4.2.1.4 Electron Donors

..........................................................147

4.2.1.4.1 Production of H

2

by Fermentation

................149

4.2.1.4.2 Competition for H

2

......................................152

4.2.1.5 Mixture of Compounds on Kinetics.................................1554.2.1.6 Temperature Effects

....................................................158

4.2.1.7 Anaerobic Oxidation

...................................................158

4.2.1.8 Electron Acceptors and Nutrients

.................................158

4.2.1.9 Field Implementation of IRZ for Enhanced Reductive Dechlorination

............................................................160

4.2.1.10 Lessons Learned

.........................................................163

4.2.1.11 Derivation of a Completely Mixed System for Groundwater Solute Transport of Chlorinated Ethenes

..170

4.2.1.12 IRZ Performance Data......................................................1774.2.2

In Situ

Metals Precipitation .............................................................1834.2.2.1 Principles of Heavy Metals Precipitation.........................1874.2.2.2 Aquifer Parameters and Transport Mechanisms ..............1954.2.2.3 Contaminant Removal Mechanisms.................................196



4.2.3

In Situ

Denitrification.......................................................................1974.2.4 Perchlorate Reduction ......................................................................199

4.3 Engineered Aerobic Systems .......................................................................2004.3.1 Direct Aerobic Oxidation.................................................................200

4.3.1.1 Aerobic Cometabolic Oxidation.......................................2024.3.1.2 MTBE Degradation ..........................................................204

4.4

In Situ

Chemical Oxidation Systems...........................................................2054.4.1 Advantages .......................................................................................2064.4.2 Concerns...........................................................................................2074.4.3 Oxidation Chemistry ........................................................................208

4.4.3.1 Hydrogen Peroxide ...........................................................2114.4.3.2 Potassium Permanganate ..................................................2134.4.3.3 Ozone ................................................................................216

4.4.4 Application .......................................................................................2184.4.4.1 Oxidation of 1,4-Dioxane by Ozone................................2224.4.4.2 Biodegradation Enhanced by Chemical

Oxidation Pretreatment.....................................................2234.5 Nano-Scale Fe (0) Colloid Injection within an IRZ ...................................223

4.5.1 Production of Nano-Scale Iron Particles .........................................2284.5.2 Injection of Nano-Scale Particles in Permeable Sediments............2314.5.3 Organic Contaminants Treatable by Fe (0) .....................................231

References

...................................................................................................233

Oxidation-reduction process plays a major role in the mobility, transport, andfate of inorganic and organic contaminants in natural waters. Hence, the manip-ulation of REDOX conditions to create an in situ reactive zone (IRZ) to meetthe cleanup objectives was a predictable evolution … .

4.1 INTRODUCTION

The concept of

in situ

reactive zones is based on the creation of a subsurfacezone, where migrating contaminants are intercepted and permanently immobilizedor degraded into harmless end products. Figures 4.1a and b pictorially describe theconcept of

in situ

reactive zones (IRZ). The successful design of these reactive zonesrequires the ability to engineer two sets of reactions: 1) between the injected reagentsand the migrating contaminants; and 2) between the injected reagents and thesubsurface environment to manipulate the bio-geo-chemistry to optimize the requiredreactions, in order to effect remediation. These interactions will be different at eachcontaminated site and, in fact, may vary within a given site. Thus, the major challengeis to design an engineered system for the systematic control of these reactions underthe naturally variable or heterogeneous conditions found in the field.

The effectiveness of the reactive zone is determined largely by the relationshipbetween the kinetics of the target reactions and the rate at which the mass flux ofcontaminants passes through it with the moving groundwater. Creation of a spatially


IN SITU

REACTIVE ZONES 133

fixed reactive zone in an aquifer requires not only the proper selection of the reagents,but also the proper mixing of the injected reagents uniformly within the reactivezone. Furthermore, such reagents must cause few side reactions and be relativelynontoxic in both its original and treated forms.

When dealing with dissolved inorganic contaminants such as heavy metals, theprocess sequence in a pump and treat system required to remove the dissolved heavymetals present in the groundwater becomes very complex, operation- and mainte-nance-intensive, and costly. In addition, the disposal of the metallic sludge, in mostcases as a hazardous waste, is also very cost prohibitive. Therefore,

in situ

treatmentmethods capable of achieving the same mass removal reactions for dissolved con-taminants in an

in situ

environment are evolving and gradually gaining prominencein the remediation industry.

The advantages of an

in situ

reactive zone to address the remediation of ground-water contamination are as follows:

• An

in situ

technology enables implementation of most ground treatment processesand eliminates the expensive infrastructure required for a pump and treat systemwith no disposal of water or wastes

Figure 4.1a

Pictorial depiction of an

in situ

reactive zone (IRZ) formation.

Figure 4.1b

Cross sectional view of the creation of an IRZ around an individual injection wellat a selected location.

Plan View

Source Area IRZ Grid

ContaminantPlume

Individual Reactive Zones Created by Individual Injection Points Providing a Collective In Situ Reactive Zone (IRZ) Curtain


ContaminantZone

Reagent



• Inexpensive installation because primary capital expenditure for this technologyis the installation of injection wells at appropriate locations

• Inexpensive operation that allows inexpensive reagents to be injected at fairly lowconcentrations and, hence, should result in insignificant cost; only samplingrequired is for groundwater quality monitoring and performance monitoringparameters are usually done in the field; remediation of large volumes of contam-inated water without any pumping or disposal needs

• Can be used to remediate deep sites because cluster injection wells or in-wellmixing systems can be installed to address deeper sites

• Unobtrusive because once the system is installed, site development and operationscan continue with minimal obstructions

•

In situ

degradation of contaminants because organic contaminants and a fewinorganics such as NH

4+

, NO

3–

, and CIO

4–

can be degraded by implementing theappropriate reactions

• Immobilization of contaminants because once the dissolved heavy metals areprecipitated out, the capacity of the soils and sediments is utilized to adsorb, filterout, and retain inorganic contaminants

Manipulation of the reduction-oxidation (REDOX) potential of an aquifer is aviable approach for

in situ

remediation of REDOX-sensitive groundwater contami-nants. In addition, various microbially induced or chemically induced reactions alsocan be achieved in an

in situ

environment. As noted earlier, creation of spatiallyfixed reactive zones to achieve these reactions is very cost effective in comparisonto treating the entire plume as a reaction zone.

Since the first IRZ for the precipitation and remediation of hexavalent chromium(Cr

6+

), was installed in 1993, this technology has advanced by leaps and bounds.

1

Currently the application of this technology can be classified into three categoriesbased on the creation of specific bio-geo-chemical and REDOX environments: 1)engineered anaerobic systems, 2) engineered aerobic systems, and 3)

in situ

chemicaloxidation.

The engineered anaerobic systems can be further divided into enhanced reductivedechlorination (ERD) systems,

in situ

denitrification,

in situ

perchlorate transforma-tion, and

in situ

heavy metals precipitation. The ERD application has been expandedto many contaminants since the first trichloroethene (TCE) application site. The IRZtechnology has been successfully applied to remediate the following chlorinatedcompounds:

• Chlorinated ethenes: tetrochloroethane (PCE), trichloroethane (TCE), dichloroet-hene (Cis 1,2 DCE, and 1,1 DCE), vinylchloride

• Chlorinated ethanes: 1,1,2,2 tetrachloroethane (1,1,2,2 PCA), 1,1,1 trichloroet-hane, (1,1,1 TCA), 1,1,2 trichloroethane (1,1,2 TCA), 1,1 and 1,2 dichloroethane(DCA), chloroethane (CA)

• Chlorinated phenols: pentachlorophenol (PCP), and tetrachlorophenol• Chlorinated pesticides• Perchlorate

In addition, the IRZ technology has been successfully applied to precipitate thefollowing dissolved metals at contaminated sites: Cr

6+

, Pb

2+

, Cd

2+

, Ni

2+

, Zn

2+

, Hg

2+

.


IN SITU

REACTIVE ZONES 135

4.2 ENGINEERED ANAEROBIC SYSTEMS

4.2.1 Enhanced Reductive Dechlorination (ERD) Systems

4.2.1.1 Early Evidence

The first microbially mediated reductive dechlorination of PCE and TCE wasobserved in the early 1980s, and this study

2,3

reported the degradation of PCE tononchlorinated end products in an acetate-fed, continuous-flow methanogenic glassbead column. It appeared that the first step in the degradation pathway was dechlori-nation to TCE. Further anaerobic oxidation of TCE to carbon dioxide and hydrochloricacid was suggested. In 1984

4

, further evidence of dechlorination of PCE beyond TCEcame in an experiment where sediments from an aquifer recharge basin were incubatedwith PCE and methanol as the electron donor. Significant concentrations of TCE,

cis

-1,2 DCE and VC were observed after three weeks, whereas in sterile controls nodechlorination had occurred. Another study in the 1980s demonstrated that dechlori-nation of PCE to VC in a methanogenic column was achievable.

5

Similar studies using

13

C-TCE, showed that TCE was dechlorinated exclusively to

cis

-DCE in soil.

6

In 1989, the first evidence of complete dechlorination of PCE to ethene undermethanogenic conditions with methanol as electron donor was demonstrated.

7

Another study found PCE reduction via ethene to ethane with lactate as electrondonor in a flow-through column filled with a mixture of polluted sediment andanaerobic granular sludge.

8

Meanwhile, numerous publications showed that micro-organisms capable of reductively dechlorinating chlorinated ethenes are abundantin polluted anaerobic environments. (An overview of the biological reductive dechlo-rination pathway of chlorinated solvents is shown in Figure 4.2.) PCE and TCE aredechlorinated mainly to

cis

-DCE, although sometimes

trans

-DCE and 1,1-DCE havealso been found as products.

9,10

However, the formation of the 1,1-DCE is believedto be a result of abiotic dechlorination in the presence of sulfide.

10

Evidence from the earlier studies indicated that the dechlorination of PCE to

cis

-DCE was found to be a relatively fast process, whereas, subsequent rates ofdechlorination of

cis

-DCE to VC and ethene were significantly slower or evenabsent.

7,11

Dechlorination of 1,1-DCE and

trans

-DCE was less studied.In some of the earlier reports and studies the dechlorination of chlorinated

ethenes was often found to be incomplete, both in the laboratory and in fieldexperiments, resulting in an accumulation of

cis

-DCE and VC. It was not fullyunderstood at that time why dechlorination beyond these compounds was problem-atic, other than raising valid questions regarding the required microbial consortiafor complete dechlorination. During that time (late 1980s and early 1990s) micro-organisms capable of dechlorinating DCE and VC had not been isolated yet, althoughseveral enrichment cultures existed. Little was known about the substrate require-ments of these bacteria. Later studies reported that PCE dechlorination in a contam-inated soil down to ethene was only achieved by adding a complex mixture of organicelectron donors. Significant research was focused, during the early to mid 1990s,on the microbial ecology that could perform complete dechlorination of PCE to



ethene and the biogeochemical conditions under which this biotransformation couldbe achieved.

The choice of a suitable electron donor for the stimulation of

in situ

dechlori-nation is still a matter of discussion and may be dependent on local conditions; thiswill be discussed in detail in a later section. When hydrogen is assumed to be themajor electron donor for dechlorination, its amendment can only be achieved byusing substrates yielding hydrogen after anaerobic degradation.

12

Often, short-chainorganic acids are produced as intermediate products, which may lead to acidificationof the groundwater and soil. Additionally, electron donors that support dechlorinationare generally readily degraded by nondechlorinating microorganisms, leading tocompetition for the substrate and excessive bacterial growth in soil pores near theinjection well. As a result, significantly more electron donor mass will be neededthan theoretically necessary to reduce all chlorinated ethenes present to ethene.

4.2.1.1.1 Biostimulation vs. Bioaugmentation

The first level of the treatment hierarchy for chlorinated ethenes is intrinsicbioremediation, or natural biodegradation, whereby indigenous microflora destroythe contaminant(s) of concern without any stimulation or enhancements. The secondchoice in this hierarchy, biostimulation or enhanced biodegradation, involves stim-ulating the indigenous microbial populations and thus enhancing microbial activity

Figure 4.2

Biological and abiotic degradation pathways of the common chlorinated com-pounds encountered at contaminated sites (adapted from McCarty and Semprini,1994; after Vogel et al., 1987, and Wiedermeier et al., 1999).

PCE CT

TCE CF

cis-1,2 DCE* DCM

VC CM

Ethene

Ethane

1,1- DCE

Acetic Acid

Primary ReactionBiotic Reactions Abiotic Reactions

CO2, H2O,CI-

1,1,1-TCE

1,1- DCA

CA

*


IN SITU

REACTIVE ZONES 137

so that they destroy the target compounds at a rate that meets the cleanup objectivesat the site. At almost every contaminated site a natural population of degradativemicroorganisms exists within the contaminated zone; however, specific nutrients,growth substrates inducers, electron donors, and electron acceptors may be requiredto create optimal microbial activity.

12

Thus, through the introduction of requiredadditional reagents, the native degradative microbial population can be stimulatedto grow, multiply, and destroy the target contaminants. Most environments containmicroorganisms able to grow on and destroy a variety of chlorinated compounds;at some sites, the persistence of these compounds, is not a consequence of theabsence of organisms but rather of the absence of the full set of conditions necessaryfor the indigenous species to function rapidly.

12

In the past there was a significantdebate among remediation experts whether the microorganisms responsible forcometabolic degradation and dehalorespiration are ubiquitous. Current belief is thatthese organisms are nearly ubiquitous.

When intrinsic bioremediation or biostimulation is not feasible at a given sitedue to the absence of an appropriate microbial population, bioaugmentation may beutilized.

Bioaugmentation

involves injection of selected exogenous microorganismswith the desired metabolic capabilities directly into the contaminated zones alongwith any required nutrients to effect the rapid biodegradation of target compounds.Two distinct bioaugmentation approaches have been developed for remediatingchlorinated ethenes. In the first approach, degradative organisms are added to com-plement or replace the native microbial population. The added microorganisms canbe selected for their ability to survive for extended periods or to occupy a specificniche within the contaminated environment. If needed, stimulants or selective cosub-strates can be added to improve survival or enhance the activity of the addedorganism. Thus, the goal of this approach is to achieve prolonged survival and growthof the added organisms and degradation of the target contaminants.

In the second bioaugmentation approach, large numbers of degradative bacteriaare added to a contaminated environment as biocatalysts which will degrade asignificant amount of the target contaminant before becoming inactive or perishing.

12

Additional microbes can be added as needed to complete the remediation process.Attempts can be made to increase the production of the degradative enzymes or tomaximize catalytic efficiency or stability, but long-term survival, growth, and estab-lishment of an active microbial population are not the primary goals of this treatmentapproach.

In the past, bioaugmentation has been implemented frequently and successfullyonly in bioreactors. The conditions in these bioreactors are controlled and quitedifferent from those in nature, and prior to start-up, no microorganisms are presentanyway. Hence, the addition of enriched cultures is essential. Furthermore, bioreac-tors are engineered and controlled systems where conditions can be readily alteredor optimized for a particular process and can be designed to promote the multipli-cation and activity of the inoculated species — in contrast to contaminated field sites.

The record of success of

in situ

bioaugmentation systems for chlorinated com-pounds has been rather spotty. On the one hand, the initiation or enhancement ofdegradation has been reported (far more commonly in samples of the contaminatedenvironments in simulated laboratory experiments) following the addition of



enriched bacterial cultures that can metabolize and grow on chlorinated ethenes. Onthe other hand, a number of failures in the field have been reported.

Such reports of failure of bioaugmentation came as no surprise to microbialecologists. Without question, a species with a substrate uniquely available to it hasa distinct advantage, yet that advantage may not be sufficient to compensate formany other traits also necessary for survival, no less multiplication, in a naturalecosystem. Possessing the requisite enzymes to metabolize a novel compound is anecessary attribute for the organism, but it is not sufficient for the organism tosucceed. Populations of introduced microorganisms are subject to a variety of abioticand biotic stresses, and these must be overcome for these organisms to be able toexpress beneficial traits.

The reasons for the frequent failures of bioaugmentation are many:

12

limitingnutrients and growth factors in the uncontrolled natural environment, suppressionby predators and parasites, inability of the introduced bacteria to penetrate significantspace, metabolism of other nontarget organic compounds present, concentration ofthe target chlorinated compound too low to support multiplication, and other inhib-itory biogeochemical conditions such as pH, temperature, salinity, and toxins.

In summary, the problems usually encountered in scaling up the bioaugmentationsuccesses achieved in laboratory experiments can be summarized as follows:

12

• Contaminant rates established in controlled laboratory studies may differ substan-tially from those in pilot-scale, full-scale, or even other laboratory studies.

• Positive biotransformation results from small systems often are not reproduced indifferent systems.

• Instantaneous biotransformation rates vary widely and in an apparently stochasticmanner, even in well-operated, steady-state systems.

4.2.1.2 Mechanisms of Reductive Dechlorination

Naturally occurring biological processes can degrade organic contaminants

insitu

or during transport in the subsurface under aerobic and/or anaerobic conditions.Microorganisms catalyze degradation reactions to obtain energy for growth, repro-duction, and cell maintenance. Useable energy is recovered through a series ofREDOX reactions where the microorganisms act as “electron transport mediators”(Figure 4.3). Biologically mediated electron transfer couples the oxidation of anelectron donor (organic compound) with the reduction of an electron acceptor (inor-ganic or organic) and results in the production of useable energy for microbialconsortia.

12,13,14

The bulk electron donor acts as a fuel source for the reactions andthe reactions proceed as long as there is a source of bioavailable electrons. Fuelsources can be the target chlorinated compounds, native organic carbon, co-contam-inants such as fuel hydrocarbons, or organic compounds such as carbohydrates. Inaerobic environments, the chlorinated compounds act as electron donors and underanaerobic conditions they act as electron acceptors.

There are two primary mechanisms involved in the biodegradation of chlorinatedorganic contaminants (Table 4.1). First, biodegradation may be growth-linked andprovide carbon and energy to support growth when the compound is used as primary


IN SITU

REACTIVE ZONES 139

substrate and directly utilized by the mediating organisms via the processes includedin Category 1. Some chlorinated solvents are used as electron donors and some areused as electron acceptors when serving as primary growth substrates. When usedas an electron donor (under aerobic and anaerobic conditions) the contaminant isoxidized. Conversely, when used as an electron acceptor, the contaminant is reducedvia the reductive dechlorination process called halorespiration.

17

In addition to their use as a primary growth substrate, chlorinated solvents canalso be degraded via cometabolic pathways. During cometabolism, microorganismsgain carbon and energy for growth from metabolism of a primary substrate, andchlorinated solvents are degraded fortuitously by enzymes present in the metabolic

Figure 4.3

Description of microorganisms acting as electron transport mediators (afterSchwarzenbach et al., 1993).

Table 4.1 Summary of the Categories of Degradation Pathways for Chlorinated Organic

Compounds (Adapted from Wiedemeier et al., 1999)

14

Category 1 Category 2

(used as primary substrate)

(used as cometabolic substrate)

Chlorinated Compound

Halo-respiration

Direct Aerobic

Oxidation

Direct Anaerobic Oxidation

Aerobic Cometabolism (co-oxidation)

Anaerobic Cometabolism

(reductive dechlorination)

Tetrachloroethylene (PCE)

X X

Trichloroethylene (TCE)

X X X

Dichloroethene (DCE)

X X X X X

Vinyl Chloride (VC) X X X X XTrichchloroethane (1,1,1 TCA)

X X X

Dichloroethane (1,2 DCA)

X X X X

Carbontetrachloride (CT)

X X

Methylenechloride (MC)

X X X

(bulk)ox

(bulk)red

(mediator)ox

(mediator)red

+ ne-+ ne-

(Contaminant)ox

(Contaminant)red

+ ne-- ne-



pathways. Cometabolism is a process where the organism receives no direct benefitfrom the degradation of the organic compound.

13,16

There are two types of cometa-bolic reactions: co-oxidation and reductive dechlorination, described as Category 2in Table 4.1. Cometabolic reactions tend to be incomplete and can possibly lead toan accumulation of more toxic daughter products. To date, vinyl chloride (VC) anddichloroethene (cis/trans) are the only chlorinated solvents that can be degraded byall aerobic and anaerobic pathways.

15

The predominant mechanism for the biodegradation of chlorinated solvents inanaerobic environments is reductive dechlorination, whether the organic compoundis a primary electron acceptor (halorespiration) or is cometabolized. Before 1994,reductive dechlorination was thought to be strictly a cometabolic process becausethe organisms that cause these reactions are ubiquitous at most contaminatedsites.

14,15

However, research has shown that combetabolic reductive dechlorinationis “sufficiently slow and frequently incomplete.”

15,18

During reductive dechlorination,the chlorinated solvents act as an electron acceptor and a chlorine atom is replacedwith a hydrogen atom (Figure 4.4).

Cometabolic reduction of the chlorinated solvents is catalyzed by the reductivedehalogenase and reductase enzymes produced by microorganisms.

14,20

Cometabolicdegradation occurs under iron reducing, manganese reducing, sulfate reducing, andmethanogenic environments.

21

The enzymes of these reducing microorganisms areinduced to reduce abiotic forms of Fe (III) to Fe (II), Mn (IV) to Mn (II), sulphateto sulfide or hydrogen sulfide, and carbon dioxide to methane. Electrons are trans-ferred to dissolved contaminants coincidentally during the reducing processes. Thesedegradation reactions are often incomplete, resulting in an accumulation of toxicdaughter products.

Just as aerobic biodegradation systems utilize oxygen as a terminal electronacceptor to stimulate microbial activity, oxidative anaerobic systems require otherterminal electron acceptors, such as nitrate or ferric iron (Fe III), to stimulatebiodegradation. Anaerobic oxidation occurs when anaerobic bacteria use the chlo-rinated contaminant as the electron donor and, in most instances, allow the micro-organism to derive useful amounts of energy from the reaction. It has been shownthat vinyl chloride can be oxidized to carbon dioxide, water, and chloride ion viaFe (III) reduction.

22

Significant anaerobic mineralization of DCE, VC, and methylenechloride also have been reported in the literature.

While in oxidative anaerobic systems the contaminant is used as an electrondonor, in reductive systems highly oxidized contaminants (such as PCE) are usedas electron acceptors. The process begins by supplying excess reduced substrate(electron donor) to a microbial consortium, i.e., a cooperative community of micro-bial species (Figures 4.3 and 4.5). The presence of the substrate expedites theexhaustion of any naturally occurring electron acceptors. As the natural electronacceptors are depleted, microorganisms capable of discharging electrons to otheravailable electron acceptors, such as oxidized contaminants, gain a selective advan-tage. The intricacies of these microbial communities are complex, but recent researchhas provided some insight into methods for enhancing populations of contaminant-degrading microorganisms.


IN SITU

REACTIVE ZONES 141

The reductive dechlorination of PCE to ethene proceeds through a series ofhydrogenolysis reactions (Figure 4.4). Each reaction becomes progressively moredifficult to carry out; subsequently, the DCEs, particularly

cis

-DCE, and vinyl chlo-ride (VC), tend to accumulate in anaerobic environments under natural conditionsdue to the absence of sufficiently reducing conditions.

Figure 4.4

Hydrogenolysis reactions of PCE during reductive dechlorination with H

2

actingas the electron donor and the chlorinated compounds acting as electron accep-tors (adapted from Vogel et al., 1987, and Wiedermeier et al., 1999).

+CI - H ion+

2- H

ElectronFlow

e

Ethane

h

HH

H

CC

Ethene

Vinyl Chloride

C C

H

H H

CI

H

HCI

CI

CC

(Limited Biological Reaction)

1,1-Dichloroethene

CI

HCI

H

CC

(Limited Biological Reaction)

trans-1,2,-Dichloroethene

(Predominant Biological Reaction)

cis-1,2,-Dichloroethene

C C

CI

H H

CI

CI

HCI

CI

CC

Trichloroethene

Perchloroethene

C C

CI

CI CI

CI

H

H

HH

H

H

CC



The oxidation-REDOX potential (ORP) affects the thermodynamics of reductivedechlorination. Microorganisms will facilitate only those oxidation-reductionreactions that have a net yield of energy. For reductive dechlorination to be thermo-dynamically favorable the REDOX potential must be sufficiently low, therebyexcluding the presence of oxygen and nitrate as terminal electron acceptors. Fur-thermore, the presence of nitrate may have an inhibitory effect on PCE dechlorina-tion.

23

The REDOX potential range for reductive dechlorination is shown in Figure4.6. It is important to note that the values of Eh ranges shown in Figure 4.6 and thevalues of ORP measured in the field by remediation engineers are not the same.Both parameters have some correlation and do not represent the same conditions.Figure 4.5 summarizes the mechanisms and the required environmental conditionfor the degradation of chlorinated solvents.

4.2.1.3 Microbiology of Reductive Dechlorination

4.2.1.3.1 Cometabolic Dechlorination

A cometabolic process is defined here as a process in which the compound ofinterest (e.g., PCE) is converted by a biological enzyme system or cofactor in whichthe compound does not serve as a source of carbon or energy.

Pure Microbial Cultures:

Reductive dechlorination is the only biodegradativeconversion known for PCE. This reaction can occur cometabolically or in a metabolicenergy-producing reaction. In both cases, the cofactors of the enzymes involved aremetal-containing porphyrins. Examples

25

of acetogenic and methanogenic bacteria thatdechlorinate PCE cometabolically are listed in Table 4.2. In general, acetogenic bacteriadechlorinate PCE at higher rates than methanogenic bacteria. Metal-containing cofac-tors have been found to catalyze the

in vitro

degradation of chlorinated ethenes.

24,25

In general, reductive dechlorination rates decrease with a declining amount ofchlorine atoms in the molecule.

In vivo

experiments with methanogenic and aceto-genic bacteria indicate that dechlorination rates are low (0.5 to 235 nmol PCE

Figure 4.5 Pictorial description of the conditions which control reductive dechlorination.

Electron Donors

ElectronAcceptor

Processes

Reducing Conditions Environmental Conditions

ORP

HydrogenGeneration

FermentableSubstrates

BTEXDissolved Organic

CompoundsTemperature pH

AnaerobicConditions

ReductiveDechlorination


IN SITU REACTIVE ZONES 143

(mg protein)–1 day–1), compared with those of halorespiring bacteria25,26 (Table 4.3).In vivo usually only one halogen atom is removed. An exception is the reductivedehalogenation of dibromoethene by Methanobacterium and Methanococcus thatyields acetylene as a product.27 However, in many studies the possible formation ofnonchlorinated products during dechlorination reactions was not included in thecarbon balance. Complete dechlorination of PCE to ethene by pure cultures ofacetogenic or methanogenic bacteria has not been observed. This is in contrast to

Figure 4.6 Optimal range for reductive dechlorination.

Table 4.2 Examples of Acetogenic and Methanogenic Bacteria that Dechlorinate PCE

Cometabolic Dechlorination Cosubstrate

Methanosarcina sp. MethanolMethanosarcina mazei MethanolSporomusa ovata MethanolAcetobacterium woodii Fructose

NO3- Reduction

REDOX Potential (Eh)in millivolts (mV) atpH=7 and T=25˚C

1,000mV

500mV

Mn4+ Reduction

Fe3+ Reduction

SO42- Reduction

Methanogenic

0mV

-500mV

Ana

erob

icA

erob

ic

PossibleRange ofReductiveDechlorination

OptimumRange forReductiveDechlorination



findings with mixed anaerobic cultures in which more extensive dechlorination hasbeen observed.7,8,27-29 The latter may be due to the interactions between differentmicroorganisms. Sometimes, however, it is difficult to distinguish between comet-abolic and specific dechlorination in these mixed cultures, and often it is not evenclear which microorganisms are responsible for the dechlorination.

The most often observed degradation pathway of PCE is via reductive dechlo-rination to cis-DCE.10,25,30-32 Several dechlorination rates for chlorinated ethenes havebeen reported in the literature, but it is difficult to compare the data because oftenthe numbers of bacteria involved were not known. Nevertheless, it can be stated thatdechlorination rates in mixed cultures are generally higher than those found forsingle acetogenic or methanogenic strains.

There are a few reports on the degradation of PCE by granular methanogenic sludgefrom upflow anaerobic sludge blanket reactors. This sludge is enriched with acetogenicand methanogenic bacteria and contains high concentrations of cofactors. Such bacterialconsortia, therefore, are suitable as a source of cometabolic dechlorinating activity. Inone of these reactors, fed with a mixture of sucrose, lactic acid, propionic acid, andmethanol as primary substrates, granular sludge showed a fast adaptation to high PCEconcentrations. Influent concentrations of 360 to 420 µM PCE were completely dechlo-rinated to ethene.26,33 Average removal rates of 7.6 µmol (g VSS)–1 day–1 were achieved,with a maximum removal rate of 28.3 µmol (g VSS)–1 day–1. A bacterial consortium ina similar reactor operated in batch mode converted PCE to TCE, cis- and trans-DCEand traces of 1,1-DCE with ethanol as the primary substrate.

4.2.1.3.2 Dechlorination by Halorespiring Microorganisms

Halorespiration is a type of anaerobic respiration in which a chlorinated com-pound is used as a terminal electron acceptor. In this reductive dechlorinationprocess, which enables the conservation of energy via electron transport phospho-rylation, one or more chlorine atoms are removed and replaced by hydrogen. Exam-ples of halorespiring bacteria species are shown in Table 4.3.

Halorespiration, also referred to as dehalorespiration, occurs when the organiccompound acts as an electron acceptor (primary growth substrate) during reductivedechlorination. During halorespiration, the chlorinated organic compounds are useddirectly by microorganisms (termed halorespirators), such as an electron acceptorwhile dissolved hydrogen serves as an electron donor:15,34

H2 + C – Cl ⇒ C – H + H+ + Cl– (4.1)

Table 4.3 Examples of Halorespiring Bacteria

Halorespiration Electron Donor

Dehalobacter restrictus H2

Dehalospirillum multivorans PyruvateDesulfitobacterium sp. strain PCE1 LactateDesulfitobacterium sp. strain TCE1 LactateStrain MS-1 Yeast extract



where C – Cl represents the chlorine bond to the carbon in the chlorinated ethenemolecule. Halorespiration occurs as a two-step process which results in the inter-species hydrogen transfer by two distinct strains of bacteria. In the first step, bacteriaferment organic compounds to produce hydrogen. During primary or secondaryfermentation, the organic compounds are transformed to compounds such as acetate,water, carbon dioxide, and dissolved hydrogen. Fermentation substrates are eitherbiodegradable, nonchlorinated contaminants (i.e., BTEX — benzene, toluene, ethylbenzene, and xylenes — or sugar) or naturally occurring organic carbon. In thesecond step, the nonfermenting microbial consortia utilize the hydrogen producedby fermentation for halorespiration.35,36 Denitrifiers, iron reducers, sulfate reducers,methanogens, and halorespirators can all utilize hydrogen as an electron donor.14

Figure 4.7 shows which reducing environment is favored depending on the hydrogenconcentration. Although compounds produced during fermentation and hydrogenhave been demonstrated to drive halorespiration,32 hydrogen appears to be the mostimportant electron donor for this process.36.37 Halorespiration has been found to belimited if available nutrients are not present. Direct injection of H2 is able to serveas an electron donor for reductive dechlorination of PCE to VC and eventually toethene in cultures provided with the proper nutritional supplements.34,38

Because reductive dechlorination of chlorinated ethenes is a reductive process,microorganisms may exist that can use chlorinated compounds as a terminal electronacceptor and possibly conserve the concomitant energy gain into ATP. This hypoth-esis, developed in the early 1990s, proved to be true.25,26 The first evidence thatbacteria exist that can couple reductive dechlorination of PCE to growth (halores-piration) under strict anaerobic conditions was presented in the early 1990s.25,39 Ahighly purified enrichment culture able to grow by the reduction of PCE to cis-DCEusing hydrogen as the electron donor was described. The dechlorinating organism,later designated Dehalobacter restrictus, uses only hydrogen as the electron donorand can couple growth to the reduction of PCE or TCE to cis-DCE. A recent study25,40

described a new isolate, strain TEA, which is closely related to Dehalobacterrestrictus. Another strict anaerobic bacterium, Dehalospirillum multivorans, capableof coupling dehalogenation of PCE to growth, was identified and describedrecently.25,41 This bacterium is less restricted concerning both electron donors andacceptors.

Several dechlorinating strains belonging to the genus Desulfitobacterium wereisolated from different sources. The strictly anaerobic D. dehalogenans, able to growby the reductive dechlorination of chlorinated phenolic compounds was isolatedrecently. Another Desulfitobacterium, strain PCE1, was isolated from polluted soiland is reported to couple the reduction of both chlorinated phenols and chlorinatedethenes to growth.43 This strain dechlorinates PCE only to TCE, whereas other knownhalorespiring microorganisms dechlorinate PCE further. The same authors alsodescribed a Desulfitobacterium sp. strain TCE1, which is able to use several electrondonors for the reduction of PCE to cis-DCE.43,44 Another researcher has describeda Desulfitobacterium sp. (strain PCE-S) that converts PCE to cis-DCE.45 All isolatedDesulfitobacterium strains are able to use a number of different electron donors andacceptors for growth. The nature and origin of the dechlorinating enzymes in theseorganisms are still unknown.



A recent study described two facultative aerobic bacteria, strain MS-1 and theclosely related Enterobacter agglomerans biogroup 5, which can reductively deha-logenate PCE to cis-DCE under anaerobic conditions.46 It is not clear yet whetherstrain MS-1 and E. agglomerans biogroup 5 are actually halorespiring organisms.

Recently, an anaerobic bacterium, Desulfuromonas chloroethenica (strainTT4B), has been isolated which can not only reductively dechlorinate PCE to cis-DCE with acetate as an electron donor, but also can reduce Fe (III) and polysulfide.These are unique features for PCE-dehalogenating organisms.47

All the above-mentioned organisms are only able to couple growth to the partialreduction of PCE or TCE. An exception is Dehalococcoides ethenogenes strain 195that couples growth to rapid dehalogenation of PCE to VC, followed by a substantiallyslower reduction to ethene.37 Growth of this bacterium is restricted to the presence ofhydrogen, which is the only electron donor supporting the dechlorination reactions.Dechlorination was only sustained by using hydrogen, acetate, vitamin B12, anaerobicdigester sludge supernatant, and cell extracts from mixed cultures in the medium.

Figure 4.7 Range of hydrogen concentrations for the different anaerobic metabolic pathways(after Wiedermeier et al., 1999).

15

10

5

0Denitrifiers Fe(III)

ReducersHalorespirators Sulfate

Reducers

Hyd

roge

n C

once

ntra

tion

(nM

)

Methanogens

Possible Reactions



Chloroethene Reductive Dehalogenases: The biochemistry of PCE dehalores-piration has been studied with enzymes that were purified from a reductively dechlo-rinating pure culture or enrichment, which is able to couple dechlorination to energyconservation (halorespiration). PCE respiration has been studied most extensivelyin Dehalobacter restrictus42,49 and Dehalospirillum multivorans.26,41,45 In general, aPCE respiration chain should contain an electron-donating enzyme, electron carriers,and a reductive dehalogenase as terminal reductase. Studies with D. multivoransand Desulfitobacterium strain PCE-S indicate that a proton gradient or a membranepotential may also be essential for chloroethene respiration because several iono-phores have been found to inhibit dechlorination in whole cell suspensions.45,50 Thenature of the electron-donating enzyme depends on the electron donor. In D. restric-tus, which uses hydrogen as electron donor for PCE respiration, hydrogenase activityhas been localized on the membrane, facing the outside.49 D. multivorans andDesulfitobacterium strain PCE-S are able to use several electron donors for dechlo-rination, and different electron-donating activities have been found.45,50 The electronsthus generated are transported to the dehalogenase via electron carriers such asquinones and cytochromes. It was demonstrated that menaquinone is involved aselectron carrier for PCE respiration in D. restrictus,49 but not in D. multivorans.45,50

Cytochrome b is present in both organisms, but its involvement in PCE respirationhas not been established.

In contrast to the well studied PCE and TCE dechlorination, little is known aboutthe mechanism of DCE and VC dechlorination. It was found that the enzymescatalyzing VC dechlorination in an enrichment culture are membrane bound and, incontrast to the known PCE reductase, cobalamin independent.51 It remains unclearwhether this enrichment is able to use VC as terminal electron acceptor. Recently,an enzyme has been obtained from an enrichment containing D. ethenogenes thatcatalyzes the dechlorination of TCE to cis-DCE, VC, and ethane. This cobalamin-containing TCE-reductive dehalogenase is membrane bound and dechlorinates itssubstrates at similar rates, as have been reported for the PCE dehalogenases. Moreresearch is needed to know what determines the difference in substrate specificityof the cobalt-containing reductive dehalogenases.

4.2.1.4 Electron Donors

The selection of an appropriate electron donor may be the most important designparameter for developing a healthy population of dechlorinating microorganismsduring implementation of an IRZ for enhanced reductive dechlorination. Recentstudies have indicated a prominent role for molecular hydrogen (H2) in the reductivedechlorination of chloroethenes.34,39,48 Most known dechlorinators can use H2 as anelectron donor; some can use only H2. Because more complex electron donors arebroken down into metabolites and residual pools of H2 by other members of themicrobial community, they may also be used to support reductive dechlorination.

From the small but growing pool of knowledge about dechlorinating organisms,it thus appears that H2 may serve an important role in reductive dechlorination ofPCE in many environments. The author recently has observed the quick or directtransformation of PCE or TCE to ethene under very reducing conditions leading to



Figure 4.8 Conceptual diagram of microbial activity to derive energy for growth and multi-plication.

Figure 4.9 Distribution of electrons to generation and cell synthesis during the breakdownof organic electron donors.

PCE

Dechlorinators

Methanogens

Ethene

4 H+ + 4CI-

CO2

H2

Acetic Acid

Complex Organics

H electrons

Second IntermediateElectron Donor

Substrate

H electrons

First IntermediateElectron Donor

Substrate

TargetElectron Donor

Substrate

Carbon SubstrateCell Synthesis Reactions

Electron Acceptor SubstrateEnergy Generation Reactions

fs electrons fe electrons





speculations of the probable effect of high H2 concentrations or reductive dechlori-nation. In natural systems, including contaminated aquifers, most H2 becomes avail-able to hydrogenotrophic microorganisms through the fermentation of more complexsubstrates by other members of the microbial consortium. The dechlorinators mustthen compete with other organisms, such as methanogens and sulfate-reducingbacteria, for the evolved H2 (Figure 4.8). Figure 4.9 also describes the distributionof electrons during the microbial breakdown of organic electron donor substrates.During studies in which ethanol or lactate was used to stimulate dechlorination inmixed anaerobic enrichment culture, both active dechlorination and methanogenesisat high H2 levels was observed; however, when H2 levels fell, dechlorination con-tinued, albeit slowly, while methanogenesis ceased entirely. It was speculated thatthe addition of electron donors fermented only under low H2 partial pressures mightgive selective advantage to dechlorinators over methanogens.

One school of thought in the past was that the rate and quantity of H2 madeavailable to a degrading consortium must be carefully engineered to limit competi-tion for hydrogen from other microbial groups, such as methanogens and sulfate-reducers. Competition for H2 by methanogens is a common cause of dechlorinationfailure in laboratory studies. As the methanogen population increases, the portionof reducing equivalents used for dechlorination quickly drops and methane produc-tion increases.17,18,36 Speculation was that the use of slowly degrading nonmethano-genic substrates would help prevent this. Recent thinking on this issue is evolvingto be different and is discussed later.

Many different compounds may serve as electron donors for the reductive dechlo-rination of chlorinated solvents (Table 4.4). Several researchers suggest that themicrobial reductive dechlorination of chlorinated ethenes depends on the presenceof molecular hydrogen as the actual electron donor, either directly available orproduced from other substrates by fermentation.32,34,55,56 Although this statementapplies to many studies, in several cases it does not hold. Acetate, from which usuallyno hydrogen is produced during anaerobic metabolism, has been shown to supportreductive dechlorination of chlorinated ethenes both in microcosms and environ-mental samples5,7,26,58 and in pure culture.47

Until recently, most research activities concerning the anaerobic degradation ofchlorinated compounds focused on methanogenic systems. Such systems typicallyinvolve the introduction of a fermentable organic compound, such as acetate, lactate,hexoses (present in molasses) or even a co-contaminant such as toluene or phenol,which is fermented to produce hydrogen, among other things. It is now clear thatthese systems probably contained at least two distinct strains of bacteria. One strainfermented the organic carbon to produce hydrogen, and another utilized the hydrogenas an electron donor for dehalorespiration.15 Only in the last two or three years haveresearchers finally recognized the role of hydrogen as the electron donor in thereductive dechlorination process.

4.2.1.4.1 Production of H2 by Fermentation

The production of H2 by different microorganisms is intimately linked with theirrespective energy metabolisms. The production of H2 is one of the specific



mechanisms to dispose excess electrons through the activity of hydrogenase presentin H

2

producing microorganisms.

59

All hydrogen producing microorganisms can becategorized into four groups:

60

• Hetertrophic facultative anaerobes that contain cytochromes and lyse formate toproduce H

2

• Desulfovibrio desulfuricans, which is the only strict anaerobe in this group witha cytochrome system

• Photosynthetic bacteria with light-dependent evolution of H

2

from reduced NADH• Strict anaerobic heterotrophs that do not contain a cytochrome system (clostridia,

micrococci, methanobacteria, etc.)

Production of H

2

by obligate anaerobic microorganisms has optimum stoichi-ometry (1:4, with glucose as substrate) compared with facultative anaerobes (1:2),although the latter process is comparatively simpler than the former.

60

Under natural conditions, fermentation is the process that generates the hydrogenused in reductive dechlorination. In the absence of externally available electronacceptors, many organisms perform internally balanced (different portions of thesame substrate are oxidized and reduced) oxidation-reduction reactions of organiccompounds with the release of energy; this process is called

fermentation

. Sinceonly partial oxidation of the carbon atoms of the organic compound occurs, fermen-tation yields substantially less energy per unit of substrate compared to oxidationreactions. (Oxidation reactions are those in which external electron acceptors par-ticipate in the reaction). For instance, the fermentation of glucose to ethanol andCO

2

has a theoretical energy yield of –57 k cal/mole, enough to produce about

Table 4.4 Electron Donors That Have Been Used to Enhance Reductive Dechlorination and Relative Costs per lb of

PCE

51–53

Electron DonorBulk Price

$/lb$/lb of PCE

Soluble (Fast Release) Donors

Methanol 0.05 0.64Milk 0.05 0.18Ethanol 0.20 – 0.25 NAMolasses 0.20 – 0.35 0.16Sugar (Corn Syrup) 0.25 – 0.30 0.40Sodium Lactate 2.20 NA

Slow Release Donors

Whey 0.05 0.04Edible Oils 0.20 – 0.50 NAFlour (Starch) 0.30 0.85Cellulose 0.40 – 0.80 NAChitin 2.25 – 3.00 NAMethyl Cellulose 4.00 – 5.00 NAHRC

(Regenesis Commercial Material) 12.00 NA

NA – Not Analyzed.



6 ATP. However, only 2 ATPs are produced, which implies that the organism operatesat considerably less than maximum efficiency.59

In any fermentation reaction, there must be a balance between oxidation andreduction. In a number of these reactions, electron balance is maintained by theproduction of molecular hydrogen, H2. In H2 production, protons (H+) of the medium,derived from water, serve as electron acceptor. The energetics of hydrogen produc-tion are actually somewhat unfavorable, so that most fermentative organisms onlyproduce a relatively small amount of hydrogen along with other fermentation prod-ucts. Fermentation reactions that have pyruvate as an intermediate product have thepotential of producing more H2. Conversion of pyruvate to acetyl-CoA is an oxidationprocess and the excess electrons generated must either be used to make a morereduced end product, or can be used in the production of H2.

Fermentation by bacteria can also be important in controlling the biogeochemicalenvironment of anaerobic aquifers. Bacterial fermentation can be divided into twocategories:14,58

Primary fermentation is the fermentation of primary substrates such as sugars, aminoacids, and lipids to yield acetate, formate, CO2, and H2, but also yields ethanol,lactate, succinate, propionate, and butyrate. While primary fermentation often yieldsH2, production of H2 is not required for these reactions to occur.

Secondary fermentation or coupled fermentation is the fermentation of primaryfermentation products such as ethanol, lactate, succinate, propionate, and butyrateto yield acetate, formate, H2, and CO2. Bacteria that carry out these reactions arecalled obligate proton reducers because the reactions must produce hydrogen inorder to balance the oxidation of the carbon substrates. These secondary fermenta-tion reactions are energetically favorable only if hydrogen concentrations are verylow (10–2 to 10–4 atm or 8000 to 80 nM dissolved hydrogen, depending on thefermentation substrate). Thus these fermentation reactions occur only when theproduced hydrogen is utilized by other bacteria, such as methanogens that convertH2 and CO2 into CH4 and H2O. The process by which hydrogen is produced by onestrain of bacteria and utilized by another is called interspecies hydrogen transfer.It should be noted that the terminal products of anaerobic decomposition, CH4, andCO2, respectively, are the most reduced and the most oxidized carbon compounds.

There are a number of compounds besides the ones listed in Table 4.4 that canbe fermented to produce hydrogen (Figure 4.10). While anaerobic degradation ofBTEX compounds has been confirmed for a long time, there is still some controversyas to whether aromatic compounds (without any oxygen in the molecule) such asthe BTEX compounds can be completely mineralized to CO2 without alternateelectron acceptors coupled solely by fermentation with methanogenesis.

Based on a number of field observations of the presence of methane, it is wellknown that fermentation occurs at almost all sites where BTEX compounds are presentin groundwater.14,53 Since methane production requires fermentation products as meth-anogenic substrates, the presence of methane is clear evidence that fermentation isoccurring. Metabolism of BTEX compounds to produce hydrogen probably requiresthe involvement of several strains of bacteria. One possible mechanism is a series ofreactions, in which other electron acceptors are used by nonfermenters to break downthe aromatics to simpler compounds that can be used by the fermenters.



4.2.1.4.2 Competition for H2

In environments where hydrogen is the most important electron donor for dechlo-rination of chlorinated solvents, competition for the uptake of hydrogen betweendifferent types of microorganisms, such as methanogenic, homoacetogenic, sulfi-dogenic, and dechlorinating bacteria, becomes important. In several studies it has beenshown that dechlorinating organisms have a higher affinity for molecular hydrogenthan methanogens.27,35,55 This indicates that the dechlorinating organisms are able tosurvive at lower hydrogen levels, but will possibly be outcompeted by other microor-ganisms when elevated hydrogen levels are present. These studies suggest that a moreeffective dechlorination may be achieved by using an electron donor that generateslow hydrogen concentrations during its fermentation, such as propionate or butyrate.The speculation is that this would then create more favorable conditions for dechlo-rinating bacteria than for hydrogen-consuming methanogens.27,35

Reductive dechlorination of PCE requires the addition of two electrons for eachchlorine removed; for three of the seven recently identified dechlorinating organisms,H2 is one of the substrates (and in some cases, the only one) that can serve as the

Figure 4.10 Steps in the process of biodegradation of PCE by reductive dechlorination. Asshown, biodegradable organic matter is required as an electron donor to initiatethe process. Different types of microbes are involved at each stage. The bottomstep shows that PCE must compete for electrons with sulfate, iron, and carbondioxide, meaning that a large amount of organic electron donors may be neededto supply enough electrons. Note: CDCE = cis-dichloroethene. Source: afterMcCarty, 1997.



direct electron donor. Dehalobacter restrictus is another direct dechlorinator thatuses only H2 as an electron donor, but dechlorinates PCE only to cis-1,2-dichloro-ethene (cisDCE).46,47 Dehalospirillum multivorans also dechlorinates PCE to cis-DCE using H2, but has a much more widely varied biochemical repertoire: it isadditionally able to use various organic substrates such as pyruvate, lactate, ethanol,formate, and glycerol as electron donors.7,37,44 Other PCE-dechlorinating organismshave been isolated that do not use H2.34,61 It was later determined that the half-velocity constant with respect to H2 for this dechlorinator was one-tenth that of themethanogenic organisms in the culture. The threshold H2 level for dechlorinationwas also correspondingly lower than values typically reported for methanogens.Though confirmed thus far with only this one dechlorinator, there are thermodynamicreasons (i.e., the relatively high free energy available from dechlorination) to suspectthat the threshold for H2 use by dechlorinators may generally be lower than that forhydrogenotrophic methanogens.15,27 This suggests a strategy for selective enhance-ment of dechlorination — managing H2 delivery so as to impart a competitiveadvantage to dechlorinators.

Numerous microcosm and site studies have shown successful stimulation ofdechlorination with substrates such as methanol, ethanol, lactate, butyrate, andbenzoate.3,32,36,62,57 However, understanding the fate of the electron donors and thatof the H2 evolved from their degradation, as well as the extent to which their reducingequivalents are channeled to desirable dechlorination or competing H2 sinks, hasimportant implications for determining how best to effectively stimulate latentdechlorinating activity for in situ enhanced reductive dechlorination in an IRZ.

Leading to a new school of thought, recent studies have suggested that the typeof substrates and the rate of fermentation may not have an impact on reductivedechlorination. One study showed the ability of four fermentable substrates to sustainPCE dechlorination long-term (i.e., approximately four months).35 The choice oforganic substrates was based upon their rates of fermentation and the H2 partialpressures that could be developed and maintained. Despite the difference in theresulting H2 partial pressures (ranging approximately 1 × 10–5 to 3 × 10–3 atm), nolong-term effect on dechlorination was observed. This result may indicate either thatlow H2 partial pressures were not required to maintain a competitive dechlorinatingcommunity or that several isolated PCE respiring bacteria do not utilize H2 as anelectron donor.43,46 H2 was not the source of PCE-reducing equivalents in all systemstested. Other laboratory and field studies have also suggested that the steady stateconcentration of hydrogen is controlled by the type of bacteria utilizing the hydrogenand is almost completely independent of the rate of hydrogen production.

As discussed earlier, when hydrogen is produced by fermentative organisms, H2

is rapidly consumed by other bacteria. This utilization of H2 by nonfermenters isknown as interspecies hydrogen transfer and is essential for fermentation reactionsto proceed.59 Note, for example, that a glucose fermenter is unable to utilize glucoseby itself so that both the glucose fermenter and the methanogen benefit from thissymbiotic relationship.

Although H2 is a waste product of fermentation, it is a highly reduced molecule,which in turn makes it an excellent, high-energy electron donor. In this symbioticrelationship, the hydrogen utilizing bacteria gain a high energy electron donor, while,



for the fermenters, the removal of hydrogen allows continuous fermentation to befavorable energetically.

In addition to methanogens, a wide variety of bacteria can utilize hydrogen asan electron donor: denitrifiers, Fe (III) reducers, sulfate reducers and halorespirators.As discussed earlier, for dechlorination to take place, halorespirators must success-fully compete against all these hydrogen utilizers.

It was suggested that the competition is mainly controlled by the Monod half-saturation constant K

s

(H

2

) (the concentration at which a specific bacterial strain canutilize hydrogen at half the maximum utilization rate).

14,27,56

The measured value ofK

s

(H

2

) for halorespirators was 100 nM and for methanogens 1000 nM.

14

This ledto the suggestion that halorespirators would compete successfully for H

2

only at lowconcentrations.

However, a more detailed analysis of halorespiration kinetics and competitionfor hydrogen based on the Monod kinetic model was performed recently.

14,27

Usingthis model, the ability of hydrogen-utilizing bacteria to compete for hydrogen caneasily be predicted from substrate concentration and two properties of the bacteria,

µ

max

(maximum specific growth rate), and K

s

. Table 4.5 lists these parameters forthe various hydrogen-utilizing bacteria.

Table 4.5 illustrates that, from the

µ

max

term, halorespirators will outcompetemethanogens and sulfate reducers at any hydrogen concentration (since at highsubstrate concentration growth rate

µ

≈

µ

max

and at low substrate concentration

µ

≈

(

µ

max

.S)/K

s

). However, denitrifiers will probably outcompete halorespirators undermost conditions as their maximum specific growth rate is approximately three timesfaster than halorespirators’. Based on these detailed analyses and the synthesis ofwide ranging data from field observations, the following probable sequence takesplace at most sites undergoing halorespiration reactions:

14,27

• Aerobic bacteria consume nonchlorinated organic substrates until the oxygen isdepleted

;

to implement enhanced reductive dechlorination, oxygen depletion isforced intentionally in an IRZ.

• Similarly, denitrifying bacteria will consume nonchlorinated organic substratesuntil the nitrate is exhausted

;

nitrate depletion will be forced for enhanced reduc-tive dechlorination.

• Iron reducing bacteria consume nonchlorinated organic substrates until the avail-able Fe (III) is expended.

Table 4.5 Maximum Specific Growth Rate (

µµµµ

max

) and Half Saturation Coefficient (K

s

) for Various H

2

-Utilizing Bacteria (Modified from

Wiedermeier et al., 1999)

Bacterial Strain

µµµµ

max

(hr

–1

) K

s

(mg/L)

Halorespirator 0.019950 0.0002Denitrifier 0.058080 —Sulfate Reducer 0.003936 —Methanogen 0.003792 0.0019



• Fermentation processes consume nonchlorinated organic substrates and generatehydrogen; dechlorinating bacteria consume hydrogen for dechlorination, whilesulfate-reducing bacteria consume nonchlorinated organic substrates and metha-nogens consume hydrogen to generate methane.

Recently researchers have found that steady state H2 concentrations in the fieldare controlled by the type of bacteria utilizing the hydrogen.14 For example, undernitrate reducing concentrations, steady state H2 concentrations were less than 0.05nM, under Fe (III) reducing conditions they were less than 0.2 to 0.8 nM, undersulfate reducing conditions they were 1 to 4 nM, and under methanogenic conditionsthey were 5 to 14 nM (Figure 4.7). Thus it is clear that an increased rate of hydrogenproduction will result in increased halorespiration without affecting the competitionbetween various bacteria for the available hydrogen (Figure 4.11). Attempting tostimulate halorespiration with poor fermentation substrates, as has been suggestedin the past, may unnecessarily limit the amount of dechlorination taking place.

It is clear from this this discussion that, during field scale enhanced reductivedechlorination at contaminated sites, the oxidative poise contributed by dissolvedoxygen, nitrate, Fe (III), Mn (IV) and sulfate has to be depleted as quickly as possibleto achieve efficient steady state reductive dechlorination reactions.53 Thus it is pru-dent to use the cheapest fermentable substrate available (see Table 4.4) to overcomethe oxidative poise (Figure 4.12).

4.2.1.5 Mixture of Compounds on Kinetics

Because few studies have systematically investigated the effect of multiple con-taminants, and not all biochemical mediators of chlorinated aliphatic transformation

Figure 4.11 Effect of oxidative poise on H2 concentrations during field scale implementationof enhanced reductive dechlorination systems, if methanogenic conditions canbe achieved and maintained (adapted from Weidermeier et al., 1999).

Hyd

roge

n C

once

ntra

tion

End

of R

emed

iatio

n

Methanogenic ConditionsSO42- DepletionFe (iii) DepletionNO3 DepletionO2

Depletion



in methanogenic cultures are known, it is difficult to predict how any two chlorinatedaliphatics may influence each other’s transformation. Several possible interactionscan be hypothesized wherein chlorinated aliphatic biotransformation rates mayincrease, decrease, or remain unaffected. Increased rates may be observed throughinduction of transformation pathways or growth on one chlorinated aliphatic, suchas dichloromethane (DCM), which then supports transformation of other aliphaticspresent. Decreased rates may result from competitive inhibition, competition forreducing equivalents, or synergistic toxicity effects. There may be no observableeffect if transformation processes are independent, or concentrations of chlorinatedaliphatics are low.

Mixtures of chlorinated aliphatics often result from reductive dechlorination ofa single parent compound; discerning the effect of mixtures on the transformationof individual compounds is difficult. The most frequently cited example is thesequential reductive dechlorination of polychlorinated ethenes (e.g., perchloroet-hene, trichloroethene) to ethene. Rates of reductive dechlorination in this series tendto decrease as the number of chlorine substituents decrease, so compounds such asvinyl chloride and dichloroethene often accumulate.64 Since all chlorinated ethenesare transformed by the same mechanism, it is conceivable that competitive interac-tions also influence the distribution of products, although it has not been systemat-ically investigated.

The sequential reductive dechlorination of 3,5-dichlorobenzoate to 3-chloroben-zoate to benzoate has been reported.63 In this case, reduction of the 3-chlorobenzoatedid not proceed until the 3,5-dichlorobenzoate was completely transformed. Thiscould not be explained by competing reaction rates, but was successfully described

Figure 4.12 Pictorial depiction of oxidative poise to be overcome during implementation ofan IRZ for engineered anaerobic systems.

SO42-

Fe3+, Mn4+

NO3-

O2

FermentationSubstrateDemand forProduction of H2



with a competitive inhibition model. Similar behavior for the reduction of 4-amino-3,5-dichlorobenzoate to 4-amino-3-chlorobenzoate also was observed.

It is important to note that competition may be significant during degradation ofchlorinated methanes and ethanes as they can be transformed by reductive dechlo-rination and other mechanisms.65,66 Methylene chloride (DCM) can be oxidized toCO2 or converted to acetate while serving as a growth substrate.7 Chloroform (CF)and carton tetrachloride (CT) can be hydrolyzed.66 1,1,1 Trichloroethane (TCA) canbe hydrolyzed or undergo dechlorination.19 Hydrolysis processes for these com-pounds are of particular interest since the products are reactive in some casesdecomposing to harmless end products or reacting with cell material — whereashydrolysis of CF and TCA yields strong nucleophiles (phosgene and acid halide),which may be toxic to the cell. Subsequent hydrolysis of these intermediates yieldsCO2 and acetate; both are subject to complete metabolism in anaerobic culture.

The interaction of dechlorination amongst CM, CF, and TCA in methanogenicacetate-enrichment cultures was investigated in another study.67 Complex interac-tions occurred when mixtures of these chlorinated aliphatics were present: TCAtransformation rates were reduced by the presence of DCM or CF, DCM transfor-mation was enhanced by CF and TCA, and CF transformation rates increased ordecreased depending on the mixture. Acetate utilization varied depending on themixture fed, complicating the interpretation of results. Where acetate utilization wasinhibited, biomass concentrations decreased and steady-state conditions were notachieved during the study.

In all cases where CF and TCA were fed together, the rate of their transformationswas lower than when they were fed individually. The decrease in the rate of trans-formation increases with the concentration of either compound, CF being moreinhibitory than TCA on mg/L basis. Results were described by a competitive inhi-bition model, which was more predictive for the effect of TCA on CF transformationthan the effect of CF on TCA transformation.67

Competitive interactions between substrates can introduce significant limitationsto bioremediation processes. Studies investigating the cometabolic transformationof chlorinated ethenes by methanotrophs,68,69 nitrifiers,70 and phenol-inducedaerobes71 have identified many problems at full scale that result from similar inter-actions. The decrease of biodegradation rates that results from competitive inhibitionmay limit the applicability of bioremediation processes. This is particularly true withCF and TCA, whose growth and biotransformation rates decrease rapidly as theirconcentrations increase.

The effect of multiple substrates on the kinetics of biotransformation reactionshas not been extensively studied. The results reported in the literature demonstratethat a range of effects may be observed even with a mixture of compounds that arestructurally similar. No unifying model can be constructed on the basis of theavailable information. These studies also demonstrate the difficulty of assessing theinteractions that occur when the compounds of interest may have dissimilar degra-dation pathways.

Unraveling the complex interactions of compounds such as those often encoun-tered in contaminated groundwaters will require further research. It is clear fromthese studies that certain combinations of compounds will lead to decreased rates



of biotransformation. This is certainly true for the combination of CF and TCA. Forengineers hoping to implement anaerobic bioremediation, this is important informa-tion in the decision making and design processes. If these interactions are notconsidered, any existing model will significantly underestimate the time requiredfor remediation. Additional studies should attempt to enhance our fundamentalunderstanding of these interactions, and identify other mixtures commonly foundthat significantly affect biodegradation rates. Such knowledge will minimize theapplication of bioremediation to sites where its efficacy would be limited.

4.2.1.6 Temperature Effects

Many studies have reported anaerobic reductive dechlorination of chlorinatedsolvents to occur within a mesophilic temperature range (20 to 37°C). However,several authors have shown dechlorination of these compounds at more ambientgroundwater temperatures between 10 and 20°C,8,26,55 indicating the applicability ofreductive dechlorination in many groundwater environments.

Microbial dechlorination has also been demonstrated under thermophilic condi-tions.26 An enrichment culture, obtained from polluted harbor sediment, rapidly dechlo-rinated PCE to cis-DCE at an optimum temperature of 65°C. Fumarate appeared tobe the best electron donor. A large number of samples from high-temperature anaerobicenvironments has been investigated for the presence of dechlorinating microorganismsas well, but no dechlorinating activity has been found.

4.2.1.7 Anaerobic Oxidation

Microorganisms can anaerobically mineralize VC and DCE in the presence ofa complex, bioavailable electron acceptor such as Fe (III) − EDTA.72,73 Studies havefocused on the possibility of oxidation of VC and DCE when they are used as aprimary growth substrate under anaerobic environments.72,73,74 These results showVC and DCE mineralization under methanogenic and iron reducing conditions inanaerobic streambed sediments without the accumulation of ethene or ethane andbuildup of carbon dioxide.14 Decreases of VC and DCE concentrations correspondedquantitatively to the production of carbon dioxide.

4.2.1.8 Electron Acceptors and Nutrients

Nutrients: In addition to proper electron donor selection, nutrient availabilitymay be a critical factor in maintaining a healthy dechlorinating consortium. In oneinstance, attempts to isolate a microbial species responsible for dechlorination ledto the discovery that nutritional factors probably had been supplied by other con-sortium members. Highly enriched dechlorinating cultures required the addition ofvitamin B12 and sludge supernatant to sustain dechlorination.38 Speculation existsthat acetogens may supply the unknown nutritional factors required by the dechlo-rinating organism(s).34 Fortunately, in situ applications support a variety of microbialspecies. This microbial diversity, combined with the addition of nutritional supple-ments, should support a healthy dechlorinating microbial community.



Alternative Electron Acceptors: Microbial dechlorination of chlorinated ali-phatic hydrocarbons has been found to occur at low REDOX potentials, mainlyunder methanogenic conditions, although dechlorination under sulfate-reducing con-ditions has also been reported.26

A recent study found that, in reactions involving polychlorinated methanes andorganic reductants exhibiting mercapto groups, an alternative initial reaction stepmay be a halophilic dissociative two-electron transfer.75 The proposed reactionmechanisms(s) involving R-S-H or R-S-S-R groups in the complete dechlorinationof polychlorinated methanes may be helpful in the (re)interpretation of microbiallymediated dechlorination reactions of such compounds.

Indirect microbial reductive dechlorination of PCE has also been observed underiron-reducing conditions due to magnetite formation by iron-reducing bacteria. Mag-netite can chemically reduce PCE to lower chlorinated ethenes.76 Besides reductionof chlorinated solvents under iron-reducing conditions, oxidation of cis-DCE andVC has been reported to occur under conditions where Fe (III) is the final electronacceptor.73 The same authors also reported the oxidation of DCE and VC in meth-anogenic, organic compound-rich bed sediment, indicating that the oxidation of thesecompounds is coupled to the reduction of humic acid compounds.74

As discussed in the previous section, the successful application of enhancedreductive dechlorination depends upon the depletion of electron-accepting chemicalspecies. The most environmentally relevant species include O2, NO3

– , Mn (IV), Fe(III), and SO4

–2 . When evaluating a site for enhanced reductive dechlorination appli-cability, one must investigate the relative abundance of these compounds in both thegroundwater and the aquifer solids. Although aqueous-phase acceptors such as O2

and NO3– take primary consideration, it is imperative that aquifer solids be charac-

terized because they can serve as a reservoir of relatively insoluble electron-acceptingspecies such as Fe (OH)3 or CaSO4. Once the electron-accepting species have beenquantified, the amount of electron donor required to deplete them can be estimatedby evaluating the stoichiometric relationship between the selected electron donorand each electron acceptor present on site (Figure 4.12). Higher levels of electronacceptor increase the oxidative poise and thus require more electron donor, thereforeraising treatment costs. A series of generic reactions is given in Table 4.6 to illustratesome of the possible reactants and products.

Once an electron donor has been selected and electron acceptors have beencharacterized, the stoichiometric relationship between them can be determined. Anequation for each electron acceptor present at the site must be balanced using the

Table 4.6 Possible Reactants and Products of Specific Terminal Electron-Accepting Processes

Predicted Reaction Process

Electron donor + O2 → CO2 + H2O Aerobic respirationElectron donor + NO3

– → CO2 + H2O + N2 DenitrificationElectron donor + Mn4+ → Mn2+ + CO2 + H2O Manganese reductionElectron donor + Fe3+ → Fe2+ + CO2 + H2O Iron reductionElectron donor + SO4

–2 → H2S + CO2 + H2O Sulfate reduction



selected electron donor. Once balanced, the molar ratio of donor to acceptor can bedetermined from these equations.

These molar ratios represent an ideal case where the entire electron donor dosageis used to reduce the electron acceptor present in the treatment zone. When calculatingthe actual electron donor dosage, a safety factor must be incorporated to account foruncharacterized electron sinks and the advective transport of electron acceptors intothe treatment zone. Site-specific conditions such as groundwater flow rate, surroundingelectron acceptor concentrations, depth to the water table, rainfall frequency, and levelof site characterization will influence the selection of the safety factor.

Because treatment alternatives and budgetary constraints are different for each site,no rule of thumb exists for screening sites based on electron acceptor concentrations.The required mass of electron donor should be estimated so its cost can be calculated.Afterwards, a site-specific, cost-benefit analysis must be undertaken to determine ifthe site is a good candidate for enhanced reductive dechlorination (ERD) application.

4.2.1.9 Field Implementation of IRZ for Enhanced Reductive Dechlorination

The author’s success and significant experience in creating an IRZ for enhancedreductive dechlorination is based purely on biostimulation of the indigenous capacityof microorganisms present at a contaminated site for dechlorination rather thanbioaugmentation. This experience is based on successful implementation of thistechnology at more than 100 sites. Creation of such an IRZ involves the addition ofan electron donor and supplemental nutrients to the contaminated groundwater zonein order to provide the optimum biogeochemical environment conducive for reduc-tive dechlorination.1

The author’s wide experience in this technology is mostly based on the solubleelectron donor, such as molasses that must be added semicontinuously or in batchinjections1 in order to sustain the microbial activity for the fermentation reactions.Recently, cheese whey has been used as a slow release substrate in less permeablegeologic environments. Preference of molasses and cheese whey is based purely oneconomics as illustrated in Figure 4.12 and Table 4.4.1

The geologic and hydrogeologic setting in which an IRZ system is installedgoverns its successful application. IRZ systems rely on the delivery of dissolvedreagents, such as dilute molasses, throughout a contaminant plume; administeringdelivery of these amendments through both the vertical and horizontal extent ofcontaminant plumes sounds deceptively easy, but requires careful engineering anda knowledge of the geologic parameters affecting groundwater flow and transport.Different configurations, in plan view and cross sections, used for IRZ systemdesigns are shown in Figures 4.13a and b, and 4.14a and b. Creative engineeringconsiderations have to be taken into account to accommodate the requirements of asmaller plume vs. a larger plume and a shallower plume vs. a deeper plume. Theultimate objective of the IRZ system design engineer should be to deliver the electrondonor as fast as possible and to create a uniformly mixed reactive zone in thesubsurface, as well as to maintain the optimum biogeochemical conditions forenhanced reductive dechlorination to occur.



The total treatment time for an IRZ will encompass the time it takes to overcomethe oxidative poise (deplete available electron acceptors), acclimatize and stimulatea healthy population of dechlorinating microorganisms, and allow the dechlorinationreactions to proceed to conclusion. Site-specific conditions will obviously influencethe total time required for treatment; for instance, anaerobic and particularly meth-anogenic sites exhibiting a significant level of natural dechlorination will requireconsiderably less time than sites with aerobic groundwater and no evidence ofdechlorination. Other factors incluencing the time required to treat a site includeaquifers with low hydraulic conductivities, which will require more time for deliveryof substrate throughout the subsurface, and the presence of DNAPL, which wouldrequire considerably longer treatment times due to the rate limitation imposed bydissolution of the contaminant.

The time required for oxidative poise depletion depends on the electron donorsupply and utilization rate, on initial electron acceptor concentrations, and the rateat which they are replenished by groundwater flow and recharge events. The largenumber of variables affecting electron acceptor depletion makes it difficult to predictthe time lag; field data indicate that this lag could be anywhere from ten days toabout three months.

When considering the time required to implement an enhanced reductive dechlo-rination IRZ system, one should include a minimum of six months to perform a

Figure 4.13a Staggered plume-wide injection grid for an IRZ system for remediation of a smallplume.

Figure 4.13b Source area staggered grid and containment curtains at mid-plume and down-gradient locations for a large-size plume.

Injection Points


Source Zone Grid

Containment Curtains




field pilot test to obtain the design parameters to design and scale up a full scalesystem. This six month testing time frame assumes one to two months for creationof a reducing environment after the depletion of the oxidative poise and three tofour months of evaluating treatment data. The actual time required for a pilot testmay exceed six months and will depend on hydrogeologic conditions and whetherthe site was already reduced with partial declorination initially.

The assessment of a particular site for IRZ application should include thedevelopment of a contaminant profile, a hydrogeological profile, and a bio-geochemical profile.

An inventory of contaminants, their concentrations, and distribution throughoutthe plume will be the first step in assessing the feasibility of an IRZ. The presence,relative concentration, and distribution of daughter products is particularly importantwhen assessing sites for enhanced dechlorination potential. Co-contaminant impactsmay be either beneficial or detrimental, so it is important to assess before the onsetof a pilot study.

The success of an IRZ application primarily depends upon the effective deliveryand distribution of the electron donor and nutrients throughout the contaminatedsubsurface. Hence, the ability to control the movement of the injected reagents isimperative at sites with a hydraulic conductivity less than or equal to 10–5 cm/sec.These sites require the addition of a slow releasing electron donor. Faster geologic

Figure 4.14a Injection well clusters for depths between 40 feet and 100 feet of saturated zonecontainment.

ReactiveZones

Dilute Molasses



settings require the addition of a soluble and fast releasing electron donor such asmolasses. Figure 4.15 illustrates the need to inject a soluble electron donor at higherconcentrations to achieve a reasonable size reactive zone from each injection pointto achieve the scale up of the full scale system cost effectively.

Biogeochemistry influences the potential for stimulating and maintaining micro-bially catalyzed reductive dechlorination. These microorganisms require highlyreducing conditions reflected by low REDOX potential measurements and the pro-duction of hydrogen sulfide and methane gas. Additional biogeochemical parameterssuch as pH, alkalinity, temperature, and dissolved organic carbon can also affect thehealth and stability of dechlorinating microorganisms.

4.2.1.10 Lessons Learned

Typically a pilot test or smaller-scale field test follows the initial screeningprocess. The two primary issues to be addressed during the field testing phase arethe provision of properly placed observation wells and allowing for sufficient timeto demonstrate the success of ERD. The amount of time it takes to see positive

Figure 4.14b In well mixing systems with submersible pumps for creating deeper IRZ systems.

SubmersiblePump

Submersible Pump

Dilute Molasses

100'

200'




results of the IRZ implementation is related to many factors, including groundwatervelocity, the time required for introduced reagents to overcome the ambient REDOXconditions in groundwater, and the locations of observation wells in relation to theinjection area. It is prudent to evaluate geochemistry and achievable degradationrates from data collected from wells located at least several months of travel timeapart. Consequently, the minimum duration of a typical pilot study is six monthswith the flexibility to extend the testing based on data collection and site-specificcosts. Field tests that are shorter in duration, or are applied in too small an area,often do not provide information that is applicable to a successful or economicallarge-scale implementation. This minimum period of time should be sufficient toovercome the initial aquifer REDOX conditions and allow for the degradation ofconstituents to a degree that will be observed in the pilot test. Some observationwells should also be placed within one to two months’ groundwater travel time fromthe injection area. This timing/placement should allow for early observations of theIRZ development, and allow for modification of the reagent injection program(strength and frequency) early enough in the planned test duration.

Reagent Delivery: The successful application of an IRZ to remediate chlorinatedsolvents in groundwater first and foremost relies on the timely and consistent deliveryof the organic carbon reagent to the treatment zone.

The author’s experience is primarily based on injecting a dissolvable sucrosesolution (molasses) as a reducing reagent. This serves as a cost effective reagent(0.20– $0.30/pound) that can aggressively alter the REDOX state of groundwater(oxidative poise) in a short time period. Other reagents, or electron donor substrates,such as edible oils and semisolid forms of lactate (such as Regenesis’ HRC®) willrely more on dissolution and diffusion for delivery. On a unit cost basis, these donors

Figure 4.15 Effects of a soluble electron donor and a slow release electron donor in an IRZwhere the hydraulic conductivity is greater than 10–5 cm/sec.

A - Distance of the Reactive Zone for the Slow Release Electron DonorB - Distance of the Reactive Zone for the Soluble Electron Donor

Slow Releasing andSlower Degrading Electron DonorSuch as Soybean Oil

Minimum ConcentrationRequired for AdequateProduction of H2

Distance From Injection Point

Con

cent

ratio

n at

Inje

ctio

n P

oint

Soluble and Fast DegradingElectron DonorSuch as Molasses

A B




are more expensive. However, the application of a slow diffusing reagent may bemore efficient in a highly reducing, slower groundwater velocity environment. Theproper reagents should be selected based on the site hydrogeology and desiredtreatment time frame. Cheese whey has been used by the author as the slow releasesubstrate in less permeable, slow moving groundwater environments.1

Based on the implementation of IRZs for the application of ERD to date, reagentdelivery becomes most complicated in low permeability geologic environments (10–5

centimeters per second (cm/sec) or less hydraulic conductivity) or those with lowgroundwater flow velocities (less than 50 ft/yr). These settings can limit the area ofinfluence of individual reagent injection points due to the absence of sufficientreagent dispersion. Poor donor delivery can also result in other potential complica-tions. These complications can include:

• Uneven application of reagent and resulting treatment; not achieving treatmentgoals

• Lack of sufficient or timely demonstration of the technology during pilot phase• Requirement of too many injection points for a full-scale application

In low permeability and/or low groundwater velocity environments, the reagentcan also accumulate in the vicinity of the injection point. Careful monitoring of thepH, ORP, and total organic carbon (TOC) levels in the groundwater near the injectionwell is necessary to avoid deleterious side effects. These effects are related tofermentation and byproduct formation and are discussed later in this section.

Natural Surfactant Effect: The injection of an abundant source of easilydegradable organic carbon during the application of ERD typically results in a rapidand large increase in the population of microorganisms in the treatment zone. As inany microbiological system, this large population increase will also result in anincrease in production of natural biosurfactants and bioemulsifiers by the microor-ganisms. Natural biosurfactants result in desorption of the chlorinated contaminantsadsorbed to the aquifer media.

To assimilate less soluble substrates, such as chlorinated solvents, microorgan-isms require a large contact area between themselves and the contaminant. Theyachieve this by emulsifying the adsorbed contaminants into an aqueous phase. Mostmicrobes frequently synthesize and excrete chemicals that promote such emulsifi-cation. These excreted chemicals fall into two main groups: biosurfactants andbioemulsifiers (Table 4.7).

Table 4.7 Microbial Surfactants

Structural Type Producing Microorganism

Carbohydrates — lipidsTrehalose — lipids

Nocardia, Mycobacterium, Coryne bacterium, Arthrobacter

Amino acids — lipidsLipopeptides

Bacillus, Streptomyces, Corynebacterium, Mycobacterium

Fatty acids — neutral lipids Pseudomonas, Acinetobacter, Mycococcus, Micrococcus, Candida

Ornithine − lipids Pseudomonas, Thiobacillus, Gluconobacter



Biosurfactants reduce the interfacial tension between water and the chlorinatedcontaminant so that the chlorinated contaminant (or less water soluble compoundssuch as PAHs) is easily micro-emulsified into the water phase. These micro-emulsiondroplets are known to be smaller than microbial cells. Some bacterial glycolipidsare extremely effective surfactants. In addition to enhancing the “mobilization” ofthe contaminants by microemulsions, biosurfactants can also increase apparent sol-ubilities by the partitioning the contaminants into surfactant micelles.

This desorption, or natural surfactant effect, is observed in many biologicaltreatment processes as an increase in the constituent levels in the treatment zoneand, in some cases, downgradient of the treatment zone. In some cases, the constit-uent concentrations in the treatment zone may remain unchanged, due to increasedsolubilization of the contaminants, for a short period even when biodegradation end-product data support the conclusion that sufficient mass is being degraded by theERD processes.

The production of surfactants that facilitate the partitioning of contaminants fromthe DNAPL to the dissolved phase (thus resulting in enhanced biodegradation) hasreceived considerable attention recently.12,13 The success of this approach dependson enhancing and maintaining biodegradation rates faster than the rate of masstransfer from NAPL to the dissolved phase.

The magnitude and composition of soil organic carbon content combined withthe distinct differences in partitioning among the chlorinated alkenes have the poten-tial to develop additional mechanisms that cause temporary increases in constituentconcentrations during enhanced reductive dechlorination applications (Figure4.16a).76 A high TOC gradient present between the groundwater and the aquifer soilmatrix, resulting from the injection of molasses, will also result in desorption ofhydrophobic contaminants for the following reasons:76

Figure 4.16a Effect of Koc on Masssorbed/Masstotal.

0.000 0.002 0.004 0.006 0.008 0.010

20

60

80

100

40

K = 36 (cis-1,2-DCE)oc

ocK = 94 (TCE)

ocK = 265 (PCE)

Per

cent

Sor

bed

Organic Carbon Fraction (f )oc


IN SITU

REACTIVE ZONES 167

• Successive dechlorination of alkene compounds is accompanied by successivedecreases in organic carbon-water partition coefficients, K

oc

(the solubility ofaliphatic compounds rapidly increases with decreasing molecular weight). If thedegradation rate for daughter compounds is equal to or lower than the rate ofparent compound degradation, temporary increases in aqueous-phase concentra-tions will be observed.

• Electron donors such as molasses, typically applied to enhance reductive dechlo-rination, comprise soluble and colloidal carbon compounds, creating an aqueous-phase organic carbon pool that was essentially nonexistent prior to the creationof the IRZ. Creation of an aqueous-phase carbon pool is expected to result in apartial migration of chlorinated alkenes to aqueous-phase carbon sorption, result-ing in increased “apparent” concentrations. Since the aqueous-phase carbon ismobile, chlorinated alkenes may also be transported from their point of origination.As the soluble organic carbon is consumed by the microbial community, a portionof the chlorinated alkenes may remain in dissolved phase for a while until eventualdegradation within the IRZ.

Intuitively, the increased desorption of target constituents within the IRZ allowsfor greater access to the typically “untreatable” adsorbed and separate phase con-taminant mass present at source areas and DNAPL locations. However, this microbialsurfactant effect must be anticipated and pilot or full-scale treatment should incor-porate provisions to evaluate and account for it. For example, the potential for initialincreases of stable parent constituent trends can be of concern to both responsibleparties and regulatory bodies as the data would tend to indicate the technology isnot working and, in some cases, could be considered as actually making conditionsworse. Therefore, during the full scale or pilot test planning stages the possibilityof this desorption effect must be evaluated in detail and anticipated in advance. Also,the possibility for an increase of dissolved chlorinated solvent concentrations tooccur in areas downgradient of the treatment zone must be addressed. Typically, an“outside-in” approach is applied, whereby a steady state IRZ is established in adowngradient portion of the plume before applying ERD to the source area. Desorbedconstituents would then move into an area already undergoing treatment and capableof treating the increased level of mass flux.

Fermentation and By-Product Formation:

During application of ERD a highlyreducing biogeochemical environment is generally created throughout the treatmentzone. This zone will also contain a large excess of organic carbon in the vicinity ofthe injection points, particularly if the geology is less permeable. During the imple-mentation of an IRZ, at (10

–5

cm/sec or less), these conditions can result in theformation of organic acids and alcohols in the groundwater as part of the degradationprocess. If the formed acids and alcohols are not consumed quickly the zone aroundthe injection zone will mimic a fermenter where additional by-products can beformed.

The formation of undesirable byproducts (including acetone and thiol com-pounds and 2-butanone) has been observed at sites where injection was initiatedwithout careful monitoring of altered groundwater conditions near the injectionwells. The occurrences of these byproducts are generally limited in extent and oftensporadic in nature. It is expected that these oxidized by-products will also be utilized



by microbes within the IRZ. Therefore, the lessons learned regarding these potentialoccurrences are as follows:

• Careful and regular monitoring of groundwater within the treatment zone shouldbe provided in order to ensure that pH levels are not depressed below pH = 4.0–4.5,and TOC levels are not excessive (site specific, but generally within 2 to 3000mg/L).

• The remediation plan for application of ERD should be flexible enough to allowfor modification of both frequency of delivery and mass of organic carbon deliv-ered to prevent the build-up of organic carbon and creation of conditions amenableto formation of these byproducts. Modifications in reagent delivery should be tiedto the pH, ORP, and TOC monitoring in the treatment zone.

Overcoming Oxidizing Conditions/High Groundwater Velocity:

As dis-cussed, the implementation of an IRZ for the application of ERD relies on thecreation of a highly reducing biogeochemical environment through provision ofexcess organic carbon to the groundwater. Achieving these conditions can be prob-lematic in groundwater flow systems in which the ambient conditions are veryoxidizing (due to shallow groundwater with abundant recharge) or the groundwatervelocity is very high (

>

1,000 ft/yr). In both situations, the amount of reagent neededto “overcome” the oxidative poise of the naturally oxidizing conditions will be costprohibitive. In addition, the scale-up cost for the full-scale system will be uneco-nomical due to extremely narrow (cigar-shaped) zones of influence from each injec-tion point.

In high groundwater velocity settings the limited transverse dispersion in ground-water can limit the extent of the reactive zone created by an individual injectionpoint. This is of particular importance in settings where drilling costs may be high(i.e., deep settings or complex geology). In such cases, these site-specific consider-ations need to be weighed against other treatment alternatives.

Biofilm Developments:

When injecting an electron donor such as molasses (andelectron acceptors) into an aquifer via injection wells, biofilm development aroundthe injection wells should be anticipated. Biofilms are large aggregations of bacteriaand other microorganisms bound together in a sticky mass of tangled polysaccharidefibers that connect cells together and tie them to a surface. Aerobic and anaerobicbacteria not only can thrive side by side within biofilms when biogeochemicalconditions permit, but also actually seem to collaborate to make themselves morepowerful. The polysaccharide coating acts like armor, giving the microorganismsprotection beyond their usual defense mechanisms.

While the typical average diameter of a bacterium in established biofilms isabout 0.5–1

µ

m, biofilm bacteria rarely adhere directly to solid surfaces. Instead, atdistances shorter than 1 nm, short-range forces such as hydrogen bonding and dipoleformation tend to be the dominant adhesion effects. As bacteria are held in placeand fed by the organic and inorganic molecules trapped by these short-range forces,they form slime that anchors them to solid surfaces. This slime becomes a home foradditional bacterial growth. If the biofilm becomes too thick to permit adequateoxygen penetration, under aerobic conditions any additional biofilm growth mayactually decrease biofilm adherence due to shearing. The thickness of the biofilm


IN SITU

REACTIVE ZONES 169

under anaerobic conditions is significantly smaller due to the above-mentionedshearing effects and the fact that the rate of biomass growth is substantially lowerunder anaerobic conditions.

Under unaerobic conditions, typical of an IRZ, reduction in porosity within thesaturated zone due to biofilm growth will not be significant enough to impact thehydrogeologic conditions for reagent transport. However, well clogging around theinjection wells is an issue to be taken into consideration. Electron donor solutions,such as dilute molasses, are injected at reasonably high concentrations of TOC beforeit gets diluted by mixing with the groundwater within the IRZ. As a result of thehigher concentrations of TOC present around the injection wells, the amount ofbiomass and biofilm growth will be significant.

Since the electron donor solutions are injected in a batch mode at most of theIRZ applications, resistance to injection due to clogging may be an operational issueonly during the injection events. In all the sites in the author’s experience, therewere only two sites where injection under pressure was difficult due to significanthead buildup. Manual cleaning of the well screens will be required under thoseconditions.

Application in Areas of Low Constituent Concentration:

The application ofERD to portions of an aquifer where the constituent concentrations are low (i.e.,less than 100

µ

g/L) can pose additional challenges. A low concentration plume willimpart less microbial conditioning and, therefore, will be more difficult to stimulatethe microbial community. In these environments, a longer lag time for microbialgrowth and conditioning should be expected. It is also difficult to observe directevidence of degradation through the monitoring program in a low concentrationplume.

Application in Areas of High Constituent Concentration/DNAPL:

Given theinherent problems with the use of conventional remediation techniques in areaswhere the constituent concentrations are very high and/or where free phase contam-inant (DNAPL) may be present, ERD has been an attractive potential alternative.The benefit of applying ERD in high concentration regimes (>50 to 200 mg/L ofchlorinated VOCs) is related to the microbial surfactant effect that usually accom-panies this technique.

The surfactant molecule is typically composed of a strongly

hydrophilic

(waterloving) group (or moiety) and a strongly

hydrophobic

(water fearing) group; in fact,the entire surfactant monomer is often referred to as amphiphillic because of its dualnature. The hydrophobic portion of the surfactant monomer is typically a longhydrocarbon chain, referred to as the “tail” of the molecule. The hydrophilic “head”group often includes anions or cations. The hydrophilic group of most surfactantsprovides a high solubility in water; however, the hydrophobic group prefers to residein a hydrophobic phase such as a DNAPL. These compelling effects result in theaccumulation of surfactant monomers at DNAPL-water interfaces (Figure 4.16b).

Physical mobilization of the residual or adsorbed DNAPL by the surfactants isundesirable and will not happen during the IRZ application due to low levels ofsurfactant production (compared to a surfactant flood). Enhanced solubilization ofthe DNAPL will take place and has to be controlled by the enhanced rate ofbiodegradation.



When the groundwater equilibrium is altered, the transfer of more constituentmass from the free or adsorbed phase into the dissolved phase should be expected.An increase in the levels of dissolved constituents in groundwater results in a moretreatable portion of the total contaminant mass. This effect can be used by itself orin conjunction with other ongoing technologies (such as pump and treat) to reducetreatment life span and costs. Care needs to be taken that increased dissolution anddesorption does not result in the vertical or horizontal migration of elevated dissolvedconcentrations from the treatment zone.

The possibility of enhancing downgradient migration is more pronounced whenapplying ERD in a potential DNAPL environment. Therefore, prior to ERD appli-cation in these settings a clear plan to address these possibilities must be developed.This could include application of the technology in an outside-in approach in whichthe downgradient areas are treated initially to develop a steady state “containmentIRZ” and encroach to the source area gradually.

However, if properly accounted for, the possibility of concentration increases,and the impacts to overcome, an ERD can be successfully applied in these settings.The application of ERD will increase the levels of mass reduction within the IRZand once the initial disruption in phase equilibrium is overcome the IRZ technologywill provide greater control of constituent migration from the source area.

4.2.1.11 Derivation of a Completely Mixed System for Groundwater Solute Transport of Chlorinated Ethenes

78

Assumptions and Definitions:

The measured concentrations in a well (C

1

)represent conditions in a unit volume of groundwater, the volume of which is definedby the saturated aquifer thickness times the effective porosity,

η

e

, as follows:

Figure 4.16b

Surfactant monomer accumulation at the DNAPL–water interface.

DNAPL Phase

Water Phase


IN SITU

REACTIVE ZONES 171

V = h

aq

×

1

×

1

×

η

e

(4.2)

This volume of water moves through the soil at a velocity computed by using Darcy’sLaw (

υ

gw

), while dissolved constituents migrate at a reduced velocity proportionalto the retardation factor:

υ

con

=

υ

gw

÷ R

f

(4.3)

Similarly, the measured concentrations (s

1

) in the volume characterize the dissolvedmass, while the total mass of the constituent can be as follows:

s

10

= s

0

×

R

f1

(4.4)

The dissolved constituent in the groundwater is assumed to decay through first orderprocess that can be represented in terms of half-life as follows:

(4.5)

As a constituent degrades, a daughter product is formed at a rate proportional to ayield factor equal to the ratio of the molecular weights of the daughter compoundsto the parent compounds:

(4.6)

Solution for a single constituent:The differential equation representing describing the concentration within this

volume written as the change in mass is equal to the mass in minus the mass outminus the rate of decay; or mathematically as:

(4.7)

where:

V = Volume [L

3

]Q = Flow through the system [L

3

/T]s = Concentration [M/L

3

]W = Mass loading term [M/T]

λ

= First order reaction coefficient [1/T]

λ1

2

12

= lnt

β122

1

=MW

MW

Vdsdt

W t Qs Vs= ( ) − − λ



Equation 4.7 can be written more concisely as follows:

(4.8)

where:

λ′

= Q +

λ

V

Equation 4.8 is a nonhomogeneous ordinary differential equation. The general solu-tion to this classification of equations can be expressed as the sum of the “comple-mentary” or general solution when W(t) equals zero, and a particular solution whenW(t) has a specific form.

s = s

c

+ s

p

Consider first, the solution to Equation 4.8 for a single constituent, with initialconditions s = s

0

, at t = 0. Dividing Equation 4.8 by V, the equation describing thecomplementary function is written as:

Separation of variables,

(4.9)

C

0

= Integration Constant

If W(t) = 0, s

p

= 0 and the above equation is also the specific solution. The integrationconstants in Equation 4.9 can be solved by exponentiation and applying the initialconditions.

Vdsdt

Qs Vs W t

Vdsdt

s Q V W t

Vdsdt

s W t

+ + = ( )

+ +( ) = ( )

+ ′ = ( )

λ

λ

λ

, or

dsdt V

s+ ′ =λ0

dss V

dt

dss V

dt

sV

t C

= − ′

= − ′

= − ′ +

∫ ∫

λ

λ

λln 0


IN SITU

REACTIVE ZONES 173

At t = 0,

s

0

=

C

0

′

, the initial concentration in the control volume. Therefore, theequation describing the change in concentration (mass) in the control volume is asfollows:

(4.10)

Equation 4.10 can be applied to simple steady-state groundwater transport problems(no dispersion) by recognizing analogous processes. Consider a well in a contami-nant plume, and measured concentrations have been stable through multiple ground-water sampling rounds, implying an equilibrium has been reached between thecontinued release of the constituent from residual source materials, and the signif-icant transport process (advection, adsorption, and degradation). The question to beanswered is how far and at what level these constituents will migrate. The controlvolume, V, is equivalent to the volume of active groundwater beneath the water table,i.e., V equals the saturated aquifer thickness, T, times the effective porosity,

η

e

. Time,t, is equal to the constituent transport time, i.e., t equals distance (x) divided by thegroundwater velocity (

υ

gw

) times the retardation factor. The unit flow through thecontrol volume, Q, is equivalent to groundwater recharge (or percolation), N. Theinitial dissolved concentration can also be expressed as ratio of the total mass to theretardation factor (

s

10

/R

f1

).

or

(4.11)

where

sV

t C

s CV

t

= − ′ +

= ′ − ′

exp

exp

λ

λ

0

0

s sV

t= − ′

0 exp

λ

s t sT

t sx

Rgw

f1 0 0 1 1( ) = − ′

= − ′

exp expλ λ

υ

s ts

RR

x

ff

gw1

101

1

1( ) = − ′

exp λυ

′ =+

λλ

11N T

T



Evaluation of Daughter Products

Now consider the presence of a second constituent, which could exist in theenvironment due to a release or be a degradation product, for example, TCE. Thefate of constituent 2 can be represented mathematically as the sum of 2 separateexpressions.

TCE(t) = f(t) + g(t)

where f(t) describes the fate of the portion of the mass that was released to theenvironment, and g(t) describes the fate of the TCE generated from the degradationof PCE. By inspection, f(t) can be written from Equation 4.11 as:

Similiarly, g(t), the mass of TCE generated by the degradation of PCE, can also bewritten from Equation 4.11 as:

or

The above expression describes the change in the total mass over time of the TCEgenerated through the degradation of PCE. Therefore, the equation describing thechange in the dissolved concentration (consistent with Equation 4.11 and f(t) above,and implicitly assuming that only the dissolved PCE degrades) is as follows:

(4.12)

The mass of TCE generated from the degradaion of PCE also degrades consistentwith f(t). Therefore g(t) is written as follows:

(4.13)

f ts

RR

x

ff

gw

( ) = − ′

202

2

2exp λ

υ

αβ

λυ

tR

s s Rx

ff

gw

( ) = − − ′

1210 10 1

1

1exp

αβ

λυ

ts

RR

x

ff

gw

( ) = − − ′

10 121

1

11 exp

ββ

λυ

ts

R RR

x

f ff

gw

( ) = − − ′

10 121

1 2

11 exp

g ts

R RR

xR

x

f ff

gwf

gw

( ) = − − ′

− ′

10 121 2

1 2

1 21

βλ

υλ

υexp exp


IN SITU

REACTIVE ZONES 175

The total change in concentration of TCE over time is therefore expressed as:

(4.14)

Now, consider

cis

-1,2 DCE, the degradation byproduct of TCE. There are threepossible fate and transport pathways for

cis

-1,2 DCE:

a)

existing

cis

-1,2 DCEb)

cis

-1,2 DCE formed by degradation of an existing source of TCE (existing)c)

cis

-1,2 DCE formed by degradation of TCE which originated from the degradationof PCE

c) The generated total

cis

-1,2 DCE from the decay of dissolved TCE from dissolvedPCE is

The equivalent dissolved

cis

-1,2 DCE is:

which changes in concentration with time:

s ts

RR

x s

R RR

xR

x

ff

gw f ff

gwf

gw2

202

10 121 2

2

2

1 2

1 21( ) = − ′

+ − − ′

− ′

exp exp expλυ

βλ

υλ

υ

a ts

RR

x

ff

gw

) expα λυ

( ) = − ′

303

3

3

b ts

R RR

xR

x

f ff

gwf

gw

) exp expββ

λυ

λυ

( ) = − − ′

− ′

20 232 3

2 3

2 31

s

R RR

xR

x

f ff

gwf

gw

10 12 231 2

1 2

1 21 1

β βλ

υλ

υ− − ′

− − ′

exp exp

s

R R RR

xR

x

f f ff

gwf

gw

10 12 231 2

1 2 3

1 21 1

β βλ

υλ

υ− − ′

− − ′

exp exp

γβ β

λυ

λυ

λυ

ts

R R RR

xR

xR

x

f f ff

gwf

gwf

gw

( ) = − − ′

− − ′

− ′

10 12 231 2 3

1 2 3

1 2 31 1exp exp exp



The equation describing the maximum down gradient concentrations of

cis

-1,2 DCEis

DCE = s

3

(t) =

α

(t) +

β

(t) +

γ

(t)

or in expanded form,

(4.15)

From inspection, the maximum concentrations of vinyl chloride, s

4

(t), are expressedas:

(4.16)

Again, from inspection, the equation for the transformation of vinyl chloride toethane can be written. Figure 4.17 describes the transformation of PCE to the finaldesired end product ethene; the shapes of the individual curves will depend on thedegradation rates, retardation factor and other biogeochemical parameters.

s ts

RR

x s

R RR

xR

x

s

R R RR

x

ff

gw f ff

gwf

gw

f f ff

gw

330

320 23

2 3

10 23 121

3

3

2 3

2 3

1 2 3

1

1

1

( ) = − ′

+ − − ′

− ′

+ − − ′

exp exp exp

exp

λυ

βλ

υλ

υ

β βλ

υ

− − ′

− ′

− − ′

1 2 3

3

2 3

3

exp exp

*exp

λυ

λυ

λυ

Rx

Rx

Rx

fgw

fgw

fgw

s ts

RR

x s

R RR

xR

x

s

R R RR

x

ff

gw f ff

gwf

gw

f f ff

gw

440

430 34

3 4

20 23 342

4

4

3 4

3 4

2 3 4

2

1

1

( ) = − ′

+ − − ′

− ′

+ − − ′

exp exp exp

exp

λυ

βλ

υλ

υ

β βλ

υ

− − ′

− ′

+ − − ′

− − ′

1

1 1

3 4

10 12 23 341 2

3 4

1 2 3 4

1 2

exp exp

exp exp

λυ

λυ

β β βλ

υλ

Rx

Rx

s

R R R RR

xR

x

fgw

fgw

f f f ff

gwf υυ

λυ

λυ

gw

fgw

fgw

Rx

Rx

− − ′

− ′

* exp exp1 3 43 4


IN SITU

REACTIVE ZONES 177

4.2.1.12 IRZ Performance Data

The author and his colleagues have implemented about 100 IRZ applications,beginning in 1993.

1

During the technology evolution a lot of lessons were learnedand have been described earlier. The performance data presented in the next fewfigures describe only the transformation or degradation of the contaminants duringIRZ applications. It should be noted that the information is not presented as site-specific case studies, due to shortage of space.

Site in California

This site was a former metal plating facility and was contaminated with TCEand Cr (VI). Very few daughter products were present prior to injection of molasses.A grid-ike IRZ injection system was installed throughout the entire plume (two acresin size) and injection of molasses began during the first quarter of 1996. Figure 4.18presents the degradation and remediation of TCE and the formed daughter productsin terms of average concentrations throughout the entire plume. Figure 4.19adescribes the highest concentration of TCE found at the site and its decline duringthe implementation of the IRZ. The increased concentration of TCE at this well (17ppm) is believed to be a result of the microbial surfactant effect. The REDOXpotential within the plume was maintained at less than –250 mV via batch injectionsof molasses. The TOC concentrations were always maintained above 200 ppm.Figures 4.19b and 4.19c describe the reduction of Cr (VI) concentrations at thesame site.

Site in Northeastern U.S.

At a site in northeastern U.S., PCE and its daughter products were found in afractured bedrock environment. The plume was very long and a pump and treatsystem was already in place at the site. Pilot studies for IRZ implementation wereperformed and the primary objective was to implement a containment IRZ curtain

Figure 4.17

Aqueous-phase concentrations of chlorinated alkene compounds resulting fromthe successive dechlorination of PCE.

Time

Aqu

eous

-pha

se c

once

ntra

tion

PCE

TCE

DCE

Vinyl Chloride

Total VOC



Fig

ure

4.1

8

Per

form

ance

dat

a on

enh

ance

d re

duct

ive

chlo

rinat

ion

at a

site

in C

alifo

rnia

.

TC

E

VC

cis-

1,2-

DC

E

Feb

.19

99O

ct.

1998

June

1998

Mar

.19

98D

ec.

1997

Oct

.19

97Ju

ne19

97A

pr.

1997

Dec

.19

96S

ept.

1996

Jun.

1996

Mar

.19

96D

ec.

1995

5000

4500

4000

3500

3000

2500

2000

1500

1000

Concentration

500

ER

D A

pplic

atio

n -

Obs

erva

tion

Wel

l CO

C C

once

ntra

tions

Apr

.19

95

0S

ept.

1995

Initi

al

Inje

ctio

n


IN SITU

REACTIVE ZONES 179

to bifurcate the plume. Figure 4.20 describes the performance of the IRZ at amonitoring well with the decline of PCE and the formation and degradation of thedaughter products. This figure also shows the formation of ethene as the final endproduct. Figure 4.21a and b is important to note because of the transformation of

cis

-1,2 DCE to the final end product ethene, and also the observed mass balance ofthe conversion.

Figure 4.22 describes the installation locations of the containment IRZ curtain,and the bifurcation of the plume in a short time frame (nine months). At this point,

Figure 4.19a

The reduction of TCE from about 17 ppm at the California site during an IRZimplementation.

Figure 4.19b

Hexavalent chromium reduction at abandoned manufacturing facility inCalifornia.

April '97

Con

cent

ratio

n

Date

TCE

DCE

Vinyl Chloride

5

10

15

20

June '97 Sept. '97 Dec. '97 Mar. '98 June '98 Sept. '98 Dec. '98

Dec. '96

Hex

aval

ent C

hrom

ium

(m

g/L)

Date

20

40

60

80

100

120

140

160

180

April '97 June '97 Oct. '97 Dec. '97

Injection ofmolasses begins



ERD applications are taking place within the source area at this site after theestablishment of the downgradient IRZ curtain.

Site in Wisconsin

This was a former dry cleaning location within a strip mall in Wisconsin. Figure4.23 describes the performance of the implemented IRZ at a monitoring well locatedwithin the plume. All monitoring wells within the plume exhibited similar perfor-mance. Due to development activities at this site, some of the monitoring wells hadto be replaced at identical locations after these activities were finished. That is thereason the pre-injection concentrations are shown as estimated instead of actualconcentrations. In all probability, the concentrations shown as estimated are theactual pre-injection concentrations at those locations.

It is important to note the decline of PCE and the formation and degradation ofthe daughter products. Ethylene concentrations were increasing until all the chlori-nated compounds were degraded.

Site in Ohio

Another IRZ for ERD is being implemented at an industrial facility in Ohio.The performance of the IRZ is described by a monitoring well located about 50 feetfrom the injection locations (Figure 4.24). The disproportionate increase in DCEconcentrations shortly after injection is believed to be due to a combination of themicrobial surfactant effects and the enhanced degradation rates of the desorbedcontaminants.

Site in North Carolina

At this site a pilot study was initiated to address contamination at very highconcentrations of TCE (more than 100 ppm). The performance of the ongoing

Figure 4.19c

Aerobic reduction of concentrations at all the wells.

Feb

. '96

Con

cent

ratio

n (m

g/L)

Date

Remediation InjectionEvents

Total Chromium

10,00

20,00

30,000

40,000

50,000

60,000

70,000

May

'96

Aug

. '96

Nov

. '96

Jan.

'97

Apr

. '97

July

'97

Oct

. '97

Jan.

'98

Apr

. '98

July

'98

Oct

. '98

Jan.

'99

Jan.

'99

Jan.

'99

Jan.

'99

Hexavalent Chromium



Fig

ure

4.2

0C

once

ntra

tion

chan

ges

in a

per

form

ance

mon

itorin

g w

ell w

ithin

an

IRZ

for

ER

D,

70 f

eet

dow

ngra

dien

t fr

om t

he in

ject

ion

wel

l. N

ote:

The

reis

a g

radu

al in

crea

se o

f th

e fin

al t

rans

form

atio

n pr

oduc

t et

hene

and

a r

easo

nabl

y st

eady

leve

l unt

il al

l the

chl

orin

ated

eth

enes

hav

e be

entr

ansf

orm

ed.



pilot study is shown in Figure 4.25. This is the first site where the author hasimplemented an IRZ when the initial chlorinated contaminant concentrations weremore than 100 ppm.

Site in PennsylvaniaAn ongoing pump and treat system had reached asymplotic concentrations at

this site after 13 years of operation and the desire was to accelerate the time requiredfor closure. Once the IRZ was established, the site was closed after reaching cleanuplevels in less than 12 months. (Figure 4.26).

Figure 4.21 Concentration changes in mg/L and µM at a monitoring well 50 feet from theinjection zone.



4.2.2 In Situ Metals Precipitation

The presence of metals in the subsurface environment can be in many forms:elemental, ionic, and/or organometallic. Distribution of the commonly encounteredmetals in the subsurface can be categorized as follows:

• Elemental form• Mercury• Lead• Gold, silver and the other noble metals• Metal alloys: brass (copper and zinc); bronze (copper, tin, and zinc); nickel-

cadmium• Ionic form

• Arsenic: As (III) arsenite, AsO3–2 ; As (V) arsenate, AsO4

–3

• Chromium: trivalent Cr (III); hexavalent Cr (VI), Cr2O7–2 and CrO4

–2

• Iron: Ferrous Fe (III); Ferric Fe (III)• Copper, lead, zinc, cadmium: Cu+1, Cu2+, Pb2+, Zn2+, Cd2+

• Mercury: Hg+1, Hg+2

• Organometallic form• Dimethyl mercury: Hg (CH3)2

• Dimethyl arsenic: AS2 (CH3)4

• Tetraethyl lead: Pb (C2H5)4

• Metal cyanide complexes: Hg (CN)2; Zn (CN)4–2 ; Cu (CN)2

–1 ; Fe (CN)6–4

The common range of concentrations of naturally encountered metals in thesubsurface environment is shown in Table 4.8.

In order to understand the fate of heavy metals in the soil-water system, it isimportant to understand the general characteristics of soil and the chemistry of heavymetals in an aqueous solution. In the aquatic environment, heavy metals may beclassified into at least two different categories: 1) in true solution as free or com-plexed ions, and 2) in particulates from adsorption onto other particles, or incorpo-ration into biomass of living organisms and inorganic precipitates such as hydrox-ides, carbonates, sulfides, and sulfates.

Figure 4.22 The effect of containment IRZs on a long plume of PCE (contamination shownare total VOCs).



Fig

ure

4.2

3C

once

ntra

tion

decl

ines

in a

mon

itorin

g w

ell w

ithin

an

IRZ

for

enh

ance

d re

duct

ive

dech

lorin

atio

n at

a s

ite in

Wis

cons

in.

Est

imat

ed P

re-R

emed

iatio

nG

roun

dwat

er C

ondi

tions

Initi

al

Inje

ctio

n

VC

cis

-1,2

-DC

E

TC

E

PC

E

Eth

ylen

e

Gro

undw

ater

Con

tam

inat

ion

Con

cent

ratio

ns v

s. T

ime

Ethylene

450

400

350

300

250

200

150

100

50 0

3000

2500

2000

1500

1000 50

0

Aug

. 20

00Ja

n.20

00Ju

ly19

99D

ec.

1998

June

1998

Nov

.19

97A

pr.

1997

Oct

.19

96

0

Concentration



Figure 4.24 Performance data from a monitoring well 50 feet downgradient of the injectionzone. Substantial increase in mass was observed due to the microbial surfactanteffects. Note the increasing concentration of the daughter products with time andthe excellent correlation of mass balance on a milli molar basis. The last datapoint shows that the transformation is complete to ethene.

Figure 4.25 Concentrations of VOCs in the pilot observation well vs. time, in situ reactivezone pilot test.

Aug. '99

Con

cent

ratio

n (u

g/L)

Date

TCE

20000

40000

60000

80000

100000

120000

140000

160000

Oct. '99 Dec. '99 Jan. '00 Mar. '00 May '00 June '00 Aug. '00 Oct. '00

cis-1,2-DCE

PCE

Vinyl Chloride



Figure 4.26

TCE and chromium concentrations vs. time.

Table 4.8 Common Concentration Range of Metals in Soils

(mg/Kg)

77

Element Range Average

Antimony (Sb) 2–10 ---Arsenic (As) 1–50 5Barium (Ba) 100–3,000 430Beryllium (Be) 0.1–40 6Cadmium (Cd) 0.01–0.7 0.06Chromium (Cr) 1–1,000 100Cobalt (Co) 1–40 8Copper (Cu) 2–100 3Lead (Pb) 2200 10Mercury (Hg) 0.02–0.3 0.03Nickel (Ni) 5–500 40Selenium (Se) 0.1–2 0.3Silver (Ag) 0.01–5 0.05Tin (Sn) 2–200 1Vanadium (V) 20–500 100Zinc (Zn) 10–300 50

Date

Con

cent

ratio

n (u

g/L)

May '90

10

100

1000

Mar. '94 Aug. '95 Mar. '96 Mar. '97

June '97

Sept. '97

Dec. '97

TCE

Chromium

Pump and TreatRemediation

In-Situ RemediationUsing Molasses

Injection



Many metals are found as very insoluble sulfide (Zn, Ag, Hg, Cu, Cd, Pb, Ni,Co) carbonate and hydroxide (Cr, Fe) forms. Biogeochemical impacts on the ground-water concentrations of species such as sulfides (from sulfate reduction) and car-bonates (via CO2 formation) enable many dissolved metallic ions to be precipitatedand immobilized.

In a soil-water system, the fate of heavy metals is directly related to their statesof identity and the existing biogeochemical conditions. The free and complexedmetal ions may be removed from solution by adsorption and precipitation mecha-nisms, while the particulate heavy metals may be transformed by their own disso-lution and filtration mechanism of soils. In principle, the concentration of heavymetals in an aqueous system is controlled by the congruent and incongruent solubilityof various oxides, carbonates, sulfates, and sulfides.

Metal precipitates in soil systems represent a selective accumulation of at leasttwo or more constituent ions into an organized solid matrix often crystalline innature. The process by which this selective accumulation occurs to form a distinctsolid phase is termed precipitation. A precipitate can be considered a particulatephase which separates from a continuous medium. The fact that solid phases formin soil-water systems means that the overall free energy of formation is negative forthe combined physical-chemical processes operating during the period of formation.The actual steps leading to the formation of a separate solid phase, however, mustoccur at the microscale level: the joining together of the constituent ions or moleculesthat will eventually be recognized as a distinct separate phase.80 Under classicalnucleation theory, three steps are generally considered necessary for those microscaleprocesses to result in the formation of crystals that will persist and survive overrelatively long periods of time: nuclei formation, crystallite formation, and crystal(precipitate) formation.80

Complexation reactions are also important in determining the saturation state ofgroundwater. A complex is an ion that forms by combining simpler cations, anions,and sometimes molecules. The cation or central atom is typically one of the metals,and the anions, often called ligands, include many of the common inorganic speciesfound in groundwater, such as S2–, CO3

2– , SO42– , PO4

3– , NO3–, Cl–. The ligand mightalso be an organic molecule such as amino acid.

4.2.2.1 Principles of Heavy Metals Precipitation

The mechanisms that can be used to reduce the concentrations of heavy metalsdissolved in groundwater are transformation and immobilization. These mechanismscan be induced by both abiotic and biotic pathways. Abiotic pathways includeoxidation, reduction, sorption, and precipitation. Examples of biotically mediatedprocesses include: reduction, oxidation, precipitation, biosorption, bioaccumulation,organo-metal complexation and phytoremediation. In this chapter, immobilizationmechanisms induced only by precipitation will be discussed.

Dissolved heavy metals can be precipitated out of solution through variousprecipitation reactions shown below. A divalent metallic cation is used as an examplein these reactions.



Hydroxide precipitation: Me++ + 2OH– → Me (OH)2 ↓ (4.17)

Sulfide precipitation: Me++ + S2– → MeS ↓ (4.18)

Carbonate precipitation: Me++ + CO3– – → MeCO3 ↓ (4.19)

Theoretical behavior of solubility of these precipitation mechanisms is shownin Figure 4.27.

Hydroxide and sulfide precipitation of heavy metals have been used successfullyin conventional industrial waste water systems. Lime (Ca(OH)2) or other alkalinesolutions such as potash (KOH) are used as reagents for hydroxide precipitation.Sodium sulfide (Na2S) is normally used as the reagent to form extremely insolublemetallic sulfide precipitates. Injection of these chemical reagents into the contami-nated aquifers to create a reactive zone will precipitate the heavy metals out ofsolution. However, injection of a reactive, pH altering chemical reagent into thegroundwater may be objectionable from a regulatory point of view. Obtaining therequired permits to implement chemical precipitation may be difficult. Furthermore,the metallic cations precipitated out as hydroxide could be resolubilized slightly asa result of any significant shift in groundwater pH.

Figure 4.27 Theoretical pathways of solubility of metals.

log

[Me+

+]

pH

Carbonate Precipitate

SulfidePrecipitate

HydroxidePrecipitate



Under reducing conditions, heavy metal cations can be removed from solutionas sulfide precipitates if sufficient sulfur is available. In systems containing a suffi-cient supply of sulfur, neutral to mildly alkaline pH and low REDOX conditions aremost favorable for the precipitation of many heavy metals. Chromium is insolubleunder reducing conditions, as Cr (III) hydroxide, but only at neutral to mildly acidicand alkaline pH values.

Precipitation as sulfides is considered the dominant mechanism limiting thesolubility of many heavy metals. Sulfide precipitation is particularly strong for“chalcophilic” metals exhibiting so-called “B-character,” such as Cu (I), Ag, Hg,Cd, Pb, and Zn; it also is an important mechanism for transition elements such asCu (II), Ni (I), Co (II), Fe (II) and Mn (II).81 Two situations can be distinguished innatural systems during sulfide precipitation conditions: the existence of a certainsulfide precipitation capacity (SPC), or (when exceeding the SPC) the accumulationof free sulfide (as H2S or HS–) in the aqueous phase. At excess sulfide concentrations,solubility of some metals can be increased by the formation of thio complexes.However, the stability of these complexes is still questionable. Possible pathwaysof metal precipitate interactions are shown in Figure 4.28. Figure 4.29 describes thefields of dominance of the different sulfur species in groundwater.

The sulfide ions necessary to mediate sulfide precipitation can be directly injectedinto a reactive zone in the form of sodium sulfide (Na2S). However, the sulfide ion(S2–) is one of the most reduced ion and its stability within the reactive zone is shortlived. It will be converted to sulfate (SO4

– –) very quickly in the presence of oxidizing

Figure 4.28 Heavy metal interactions in an aquifer matrix.

Mineral Surface Organic Surface

Precipitate Occlusion Living Biomass

InorganicComplex

FreeIon

OrganicComplex



conditions within the contaminated plume. Addition of a very easily biodegradableorganic substrate, such as carbohydrates, will enhance the formation of reduced,anaerobic conditions by depleting the available oxidation potential. The presence ofcarbohydrates serves two purposes: microorganisms use it as their growth substrateby depleting the available oxygen, and they use it as an energy source for thereduction of sulfate to sulfide.

Indirect microbial transformation of metals can occur as a result of sulfatereduction when anaerobic bacteria oxidize simple carbon substrates with sulfateserving as the electron acceptor. The net result of the process is the production ofhydrogen sulfide (H2S) and alkalinity (HCO3

–). Sulfate reduction is strictly an anaer-obic process and proceeds only in the absence of oxygen. The process requires asource of carbon to support microbial growth, a source of sulfate, and a populationof sulfate reducing bacteria. Dilute black strap molasses solution is an ideal feed

Figure 4.29 Fields of dominance of sulfur species at equilibrium at 25˚C and 1 atmosphere(adapted from Hem, 1985).

Eh,

in V

olts

pH

420

1.40

1.20

1.00

0.80

0.60

0.40

0.20

0.00

-0.20

-0.40

-0.60

-0.80

-1.006 8 10 12 14

HS-

H2Saq

S2

S0

Water Reduced

Water Oxidized

SO2-4

HSO-4



substrate for this purpose since typical black strap molasses contains approximately20% sucrose, 20% reducing sugars, 10% sulfated ash, 20% organic non sugars, and30% water.1

Whether formed biotically or abiotically, the metal sulfides result from an inter-action between the metal ion and sulfide ion:

Me2+ + S2– → MS ↓ (4.20)

It is the source of the sulfide that determines whether a biological agent isimplicated in metal sulfide formation. If the sulfide results from bacterial sulfatereduction or from bacterial mineralization of organic compounds, it is obviously ofbiotic origin. If it is derived from volcanic activity, it is generally of abiotic origin.The metal sulfides, because of their relative insolubility, form readily at ordinarytemperatures and pressures by interaction of metal ions and sulfide ions. Table 4.9lists solubility products for some common simple sulfides.83

The following calculation will show that relatively low concentrations are neededto form metal sulfides by reacting with H2S at typical concentrations that can beformed in an anaerobic IRZ.83 Let us examine, for instance, the case of iron. Thedissociation constant for iron sulfide (FeS) is:

[Fe2+][S2–] = 1 x 10–19 (4.21)

The dissociation constant for H2S is:

(4.22)

since,

(4.23)

and,

Table 4.9 Solubility Products for Some Metal Sulfides81

CdS 1.4 × 10–28 FeS 1 × 10–19 NiS 3 × 10–21

Bi2S3 1.6 × 10–72 PbS 1 × 10–29 Ag2S 1 × 10–51

CoS2 7 × 10–23 MnS 5.6 × 10–16 SnS 8 × 10–29

Cu2S 2.5 × 10–50 Hg2S 1 × 10–45 ZnS 4.5 × 10–24

CuS 4 × 10–38 HgS 3 × 10–53 H2S 1.1 × 10–7

HS- 1 × 10–15

S xH S

H

2 22 221 1 10− −

+[ ] = [ ][ ]

.

HS H

H Sx

s

− +[ ][ ][ ] = 1 1 10 7. –



(4.24)

Therefore,

(4.25)

About 5.08 × 10–3 mg of Fe2+ per liter (9.1 × 10–8 M) will be precipitated by 3.4mg of hydrogen sulfide per liter (10–4 M) at pH 7. The unused H2S will ensure reducingconditions, which will keep the iron in the ferrous state. Since ferrous sulfide is oneof the more-soluble sulfides, it can be seen that metals whose sulfides have even smallersolubility products will form even more readily at lower H2S concentrations.

Metal sulfides have been generated in laboratory experiments utilizing H2S frombacterial sulfate reduction. It has been reported that sulfides of Sb, Bi, Co, Cd, Fe, Pb,Ni, and Zn were formed in a lactate broth culture of Desulfovibrio desulfuricans towhich sulfate and salts of selected metals had been added.83 Metal toxicity to Des-ulfovibrio desulfuricans depends in part on the concentration of the metallic ion inquestion. Obviously, for the corresponding metal sulfide to be formed, the metal sulfidemust be even more insoluble than the starting compound of the metal. More metalssuch as Cu, Ag, Cd, Pb, Zn, Ni, and Co, in addition to Fe and Mn, can also beprecipitated as metallic sulfides. Precipitated metallic sulfides will remain in an insol-uble, stable form, unless the subsurface REDOX conditions change dramatically.

The production of alkalinity from sulfate reduction, denitrification, and otherreactions causes an increase in pH, which can result in metal precipitation through theformation of insoluble metal hydroxides or oxides. This process follows the reaction:

Me2+ + 2H2O → Me (OH)2 ↓ + 2H+ (4.26)

Chromium PrecipitationIn situ microbial reduction of dissolved hexavent chromium Cr (VI) to trivalent

chromium Cr (III) yields significant remedial benefits because trivalent chromiumCr (III) is less toxic, water insoluble, and, thus, nonmobile, and precipitates outof solution. In fact, it has been stated that the natural attenuation of Cr (VI) to thereduced Cr (III) form within an aquifer is a viable groundwater remediationtechnique.

In situ microbial reduction of Cr (VI) to Cr (III) can be promoted by injectinga carbohydrate solution, such as dilute molasses solution. The carbohydrates, whichconsist mostly of sucrose, are readily degraded by the heterotrophic microorganismspresent in the aquifer, thus depleting all the available dissolved oxygen present inthe groundwater. Depletion of the available oxygen present causes reducing condi-tions to develop. The mechanisms of Cr (VI) reduction to Cr (III) under inducedreducing conditions can be 1) likely a microbial reduction process involving Cr (VI)

S H

HSx

2151 10

− +

−−[ ][ ]

[ ] =

FeH

H Sx

xx

H

H Sx2

2

2

19

22

2

2

21 101 1 10

9 1 10++ −

−

+

[ ] = [ ][ ] = [ ]

[ ] ( ).

.



as a terminal electron acceptor for the metabolism of carbohydrates by species suchas Bacillus subtilis; 2) an extra cellular reaction with by-products of sulfate reductionsuch as H2S; and 3) abiotic oxidation of the organic compounds including the soilorganic matter such as humic and fulvic acids.84

Cr (VI) is known to be reduced both aerobically and anaerobically in differentbacterial systems. In anaerobic systems, membrane preparations reduce Cr (VI),which has been shown to serve as a terminal electron acceptor. Aerobic reductionof Cr (VI) has been found to be associated with soluble proteins. The enzymaticbasis for aerobic chromate reduction is not known, but it has been proposed thatchromate may be reduced by a soluble reductase enzyme with a completely unrelatedprimary physiological role. Based on the diversity of Cr (VI) reducing microorgan-isms in soil, provision of a suitable electron donor such as molasses may be sufficientand the ORP within the IRZ need not be reduced to –250 to –300 mV as is the caseduring ERD applications.85,86

The primary end product of Cr (VI) to Cr (III) reduction process is chromichydroxide [Cr (OH)3], which readily precipitates out of solution under alkaline tomoderately acidic and alkaline conditions.87 To ensure that this process will provideboth short term and long term effectiveness in meeting groundwater cleanup objec-tives, the chromium precipitates must remain immobilized within the soil matrix ofthe aquifer, and could not be subject to Cr (OH)3 precipitate dissolution or oxidationof Cr (III) back to Cr (VI) once groundwater conditions revert back to naturalconditions. Based on the results of significant research being conducted on the insitu chromium reduction process, it is readily apparent that the Cr (OH)3 precipitateis essentially an insoluble, stable precipitate, immobilized in the soil matrix ofthe aquifer.

Contrary to the numerous natural mechanisms that cause the reduction of Cr(VI) to Cr (III), there appear to be only a few natural mechanisms for the oxidationof Cr (III). Indeed, only two constituents in the subsurface environment (dissolvedoxygen and manganese dioxide) are known to oxidize Cr (III) to Cr (VI).88 Theresults of studies conducted on the potential reaction between dissolved oxygen andCr (III) indicate that dissolved oxygen will not cause the oxidation of Cr (III) undernormal groundwater conditions. However, studies have shown that Cr (III) can beoxidized by manganese dioxides, which may be present in the soil matrix. However,only one phase of manganese dioxides is known to oxidize appreciable amounts ofCr (III) and this process is inversely related to groundwater pH. Hence, the oxidationof Cr (III) back to Cr (VI) in a natural aquifer system is highly unlikely.

The Cr (OH)3 precipitate has an extremely low solubility (solubility product,Ksp = 6.7 × 10–31) and thus, very little of the chromium hydroxide is expected toremain in solution. It has been reported that aqueous concentration of Cr (III), inequilibrium with Cr (OH)3 precipitates, is around 0.05 mg/L within the pH rangeof 5 to 12 (Figure 4.30). The pH range of natural aquifer systems will be within 5to 12 and, hence, the potential for the chromic hydroxide to resolubilize is unlikely.Furthermore, the potential for co-precipitation with Ferric ions will further decreasethe solubility of Cr (OH)3.

Dissolved Cr (VI) can be also precipitated as Cr (OH)3 in a reactive zone by theinjection of ferrous sulfate solution into a reactive zone at appropriate concentrations.



Cr (VI) exists as chromate, CrO42– , under neutral or alkaline conditions and dichro-

mate, Cr2O72– , under acidic conditions. Both species react with ferrous ion:

Acidic conditions: Cr2O72– + 6Fe2+ + 14H+ → 2Cr3+ + 6Fe3+ + 7H2O

(4.27)

Neutral or alkaline condition: CrO42– + 3Fe2+ + 4H2O → Cr3+ + 3Fe3+ + 8OH–

(4.28)

Both Cr (III) and Fe (III) ions are highly insoluble under natural conditions ofgroundwater (neutral pH or slightly acidic or alkaline conditions).

Fe3+ + 3OH– → Fe (OH)3 ↓ (4.29)

Cr3+ + 3OH– → Cr (OH)3 ↓ (4.20)

The addition of ferrous sulfate into the reactive zone may create acidic con-ditions and, hence, the zone downgradient of the ferrous sulfate injection zonemay have to be injected with soda ash or caustic soda to bring the pH back toneutral conditions.

Arsenic PrecipitationSoluble arsenic occurs in natural waters only in the pentavalent, As (V) and

trivalent, As (III), oxidation states. Although both organic and inorganic forms ofarsenic have been detected, organic species (such as methylated arsenic) are rarely

Figure 4.30 Cr (III) concentration in equilibrium with Cr (OH)3.

log

[Cr(

lll)]

pH

4

-2

6 8 10 12 14

Cr(OH)3(am)

MCL

-8

-6

-4



present at concentrations greater than 1 ppb and are generally considered of littleenvironmental significance compared with inorganic arsenic species. Thus, thisdiscussion focuses exclusively on the behavior of inorganic arsenic.

Thermodynamics provides useful insight into the equilibrium chemistry of inor-ganic arsenic species. In oxygenated waters, As (V) is dominant, existing in anionicforms of either H2AsO4

– , HAsO42– , or AsO4

3– over the pH range of 5 to 12, whichcovers the range encountered in natural groundwater. Under anoxic conditions, As(III) is stable, with nonionic (H3AsO3) and anionic (H2AsO3

– ) species dominantbelow and above pH 9.22, respectively. In the presence of sulfides, precipitation ofAsS (realgar) or As2S3 (orpiment) may remove soluble As (III) and exert considerablecontrol over trace arsenic concentrations. The thermodynamic reduction of As (V)to As (III) in the absence of oxygen could be chemically slow and may requirebacterial mediation.13 As noted in the previous section, injection of dilute solutionof blackstrap molasses will create the reducing conditions for As (V) to be reducedto As (III) and also provide the sulfide ions for As (III) to precipitate as As2S3. Thesereactions are described by the following equations:10

Reduction of As (V) to As (III) under anaerobic conditions:

HAsO42– → HAsO2

In the presence of S– – under anaerobic conditions:

HAsO2 + S– – → As2S3

Within oxygenated zones in the aquifer, oxidation of Ferrous ion (Fe (II)) andMn (II) leads to formation of hydroxides that will remove soluble As (V) bycoprecipitation or adsorption reactions. The production of oxidized Fe-Mn speciesand subsequent precipitation of hydroxides are analogous to an in situ coagulationprocess for removing As (V).

4.2.2.2 Aquifer Parameters and Transport Mechanisms

REDOX processes can induce strong acidification or alkalinization of soils andaquifer systems. Oxidized components are more acidic (SO4

2– , NO3–) or less basic

(Fe2O3) than their reduced counterparts (H2S, NH3). As a result, alkalinity and pHtend to increase with reduction and decrease with oxidation. Carbonates are efficientbuffers in natural aquifer systems in the neutral pH range.

Many events can cause changes in REDOX conditions in an aquifer. Infiltrationof water with high dissolved oxygen concentration, fluctuating water table, excessorganic matter, introduction of contaminants that are easily degradable, increasedmicrobial activity, and deterioration of soil structure can impact the REDOXconditions in the subsurface. However, there is an inherent capacity to resist REDOXchanges in natural aquifer systems. This inherent capacity depends on the availabilityof oxidized or reduced species. REDOX buffering is provided by the presence ofvarious electron donors and electron acceptors present in the aquifer.



An engineered in situ reactive zone has to take into consideration how the targetreactions will impact the REDOX conditions within and downgradient of the reactivezone, in addition to degrading the contaminants with the available residence time.Furthermore, careful evaluation should be performed regarding the selectivity of theinjected reagents towards the target contaminants and the potential to react withother compounds or aquifer materials. Careful monitoring, short term and long term,should be performed to determine whether the natural equilibrium conditions canbe restored at the end of the remediation process. In some cases modified bio-geochemical equilibrium conditions may have to be maintained over a long periodof time to prevent the reoccurrence of contaminants.

4.2.2.3 Contaminant Removal Mechanisms

As noted earlier, the mechanisms used to reduce the toxicity of dissolved con-taminants can be grouped into two major categories: transformation and immobili-zation. Examples of some of these mechanisms have been discussed earlier. Con-version of chlorinated organic compounds to innocuous end products such as CO2,H2O, and Cl− by either biotic or abiotic reaction pathways is an example of trans-formation mechanisms. Precipitation of Cr (VI) as Cr (OH)3 by either abiotic orbiotic reaction pathways and subsequent filtration by the soil matrix is an exampleof immobilization mechanisms.

It can be assumed, in most cases, that the end products of transformation mecha-nisms will result in dissolved and gaseous species and that the impact of these endproducts on the natural REDOX equilibrium will be short term. If the impact is expectedto be significant, it can be controlled by limiting the reaction kinetics and transport ofthe end products from the reaction zone. Dilution and escape of dissolved gases willalso help in restoring the natural equilibrium conditions in the reaction zone.

Immobilization mechanisms, which include heavy metals’ precipitation reac-tions, in reality transform the contaminant into a form (precipitate) which is muchless soluble. In addition, transport of dissolved heavy metals in groundwater shouldbe considered a two-phase system in which the dissolved metals partition betweenthe soil matrix and the mobile aqueous phase.

Metal precipitates resulting from an in situ reactive zone may move in associationwith colloidal particles or as particles themselves of colloidal dimensions. The termcolloid is generally applied to particles with a size range of 0.001 to 1 µm. Thetransport of contaminants as colloids may result in unexpected mobility of lowsolubility precipitates. It is important to remember that the transport behavior ofcolloids is determined by the physical/chemical properties of the colloids as well asthe soil matrix.

Generally, when fine particles of colloid dimensions are formed, flocculationnaturally occurs unless steps are taken to prevent it. Even when the primary precip-itates are of colloid dimensions, if they form larger lumps a stable dispersed transportcannot take place. These larger flocs will settle on the soil matrix.

Metal precipitates may be pure solids (e.g., PbS, ZnS, Cr (OH)3) or mixed solids(e.g., (Fex, Cr1–x) (OH)3, Ba(CrO4, SO4)). Mixed solids are formed when variouselements co-precipitate or due to interaction with aquifer materials.



Colloidal precipitates larger than 2 µm in the low flow conditions common inaquifer systems will be removed by sedimentation. Colloidal precipitates are moreoften removed mechanically in the soil matrix. Mechanical removal of particlesoccurs most often by straining, a process in which particles can enter the matrix,but are caught by the smaller pore spaces as they traverse it.

Colloidal particles below 0.1 µm will be subjected more to adsorptive mecha-nisms than mechanical processes. Adsorptive interactions of colloids may be affectedby the ionic strength of the groundwater, ionic composition, quantity, nature, andsize of the suspended colloids, geologic composition of the soil matrix, and flowvelocity of the groundwater. Higher levels of total dissolved solids (TDS) in thegroundwater encourage colloid deposition.

In aquifer systems with high Fe concentrations, the amorphous hydrous ferricoxide can be described as an amphoteric ion exchange media. As pH conditionschange, it has the capacity to offer hydrogen ions (H+) or hydroxyl ions (OH–) forcation or ion exchange, respectively. Adsorption behavior is primarily related to pH(within the typical range of 5.0 to 8.5), and at typical average concentrations in soil,the iron in a cubic yard of soil is capable of adsorbing from 0.5 to 2 pounds ofmetals as cations or metallic complexes. This phenomenon is extremely useful forthe removal of As and Cr.

4.2.3 In Situ Denitrification

Nitrogen can form a variety of compounds due to its different oxidation states.In the natural ecosystem, most changes from one oxidation state to another arebiologically induced. The nitrogen forms in Table 4.10 are of interest in relation tothe subsurface environment.

The unionized, molecular ammonia exists in equilibrium with the ammoniumion, the distribution of which depends upon the pH and temperature of the bio-geochemical system; in fact, very little ammonia exists at pH levels less than neutral.Transformation of nitrogen compounds can occur through several mechanisms,including fixation, ammonification, synthesis, nitrification, and denitrification.

Ammonification refers to the change from organic nitrogen to the ammoniumform. In general, ammonification occurs during decomposition of animal and planttissue and animal fecal matter and can be expressed as follows:

Table 4.10 Nitrogen Forms Present in the Subsurface Environment

Nitrogen Compound Formula Oxidation State

Ammonia NH3 –3Ammonium ion NH4

+ –3Nitrogen gas N2 0Nitrite ion NO2

– +3Nitrate ion NO3

– +5



Organic Nitrogen + Ammonifying Microorganisms → NH3/NH4+ (4.31)

Nitrification refers to the biological oxidation of ammonium ions under aerobicconditions by the chemoautotrophic organisms called nitrifiers. Two specificchemoautotrophic bacterial genera are involved, using inorganic carbon as theirsource of cellular carbon:

(4.32)

The transformation reactions are generally coupled and proceed rapidly to thenitrate form.

In situ denitrification can be accomplished by organisms belonging to the generaMicrococcus, Pseudomonas, Denitrobacillus, Spirillum, Bacillus, Achromobacter,Acinetobacter, Gluconobacter, Alcaligens, and Thiobacillus, which are present in thegroundwater environment. Denitrifying organisms will utilize nitrate or nitrite in theabsence of oxygen as the terminal electron acceptor for their metabolic activity. Ifany oxygen is present in the environment, it will probably be used preferentially.The energy for the denitrifying reactions is released by organic carbon sources thatact as electron donors. The microbial pathways of denitrification include the reduc-tion of nitrate to nitrite and the subsequent reduction of nitrite to nitrogen gas.

NO3– → NO2

– → N2 ↑ (4.33)

In biological wastewater treatment processes employing denitrification, a cheap,external carbon source such as methanol is added as the electron donor. It has longbeen known that NO3

– can be converted to N2 gas in anaerobic groundwater zonesin the presence of a labile carbon source.

In situ microbial denitrification is based on the same principle as conventionalbiological wastewater treatment systems, except that it is carried out in the subsurfaceby injecting the appropriate organic carbon source. Since methanol could be anobjectionable substrate from a regulatory point of view, sucrose or sugar solution isan optimum substrate to be injected.

It should be noted that in the hierarchy of REDOX reactions, NO3– is the most

favored electron acceptor after dissolved oxygen. Hence, considerable attentionshould be focused in maintaining the REDOX potential in the optimum range, sothat Mn (IV), Fe (III), sulfate reduction conditions or methanogenic conditions arenot formed in the subsurface. Furthermore, since denitrification is a reduction reac-tion, alkalinity and pH tend to increase in the aquifer. Since the end product N2 gaswill escape into the vadose zone and, hence, the aquifer system is not a closedsystem, increased alkalinity will be observed in the groundwater. If the NO3

– con-centration is not very high, this concern will be short lived.

NH O NO O NO4 2 2 2 3+ − −+ → + →Nitrosomonas

bacteriaNitrobacter

bacteria


IN SITU

REACTIVE ZONES 199

4.2.4 Perchlorate Reduction

Perchlorate has been widely used as a propellant in solid rocket fuel and hasrecently been identified as a contaminant in both groundwater and surface waters.Perchlorate is recognized by the USEPA as a potential health risk; California hasset a drinking water action level of 18 ppb.

Most of the perchlorate contamination in groundwater appears to have comefrom the legal discharge decades ago of then unregulated waste effluents containinghigh levels of ammonium perchlorate. Although ammonium perchlorate was releasedinitially, the salt is highly soluble and dissociates completely to ammonium andperchlorate ions upon dissolving in water:

NH

4

ClO

4

Æ

NH

4+

+ ClO

4–

(4.34)

It is likely that most of the ammonium has been biodegraded and the cation isnow best viewed as mostly Na

+

or possibly H

+

, especially where perchlorate (ClO

4–

)levels are below 100 ppb. At those sites where contamination dates back decades,very little (if any) ammonium has been found.

89

The persistence of perchlorate in groundwater aquifers results primarily from acombination of aerobic conditions and lack of an electron donor. A number ofbacteria that contain nitrate reductases are capable of dissimilatory reduction ofperchlorate.

89,90

Many mixed cultures have reduced perchlorate, chlorate, chlorite,nitrate, nitrite, and sulfate under the right conditions. Inhibition of perchloratereduction also has been observed in the presence of other substrates, particularlychlorate, chlorite, and sulfate.

90

Chlorate reductase has been isolated from microor-ganisms that also possess nitrate reductase. Although most perchlorate strains maybe denitrifying facultative anaerobes, not all denitrifiers are (per)chlorate reducers.Simultaneous reduction of NO

3–

and ClO

4–

has been demonstrated in laboratorystudies.

90,91

The conversion of chlorine in perchlorate to chloride requires the overall transferof eight electrons. The sequence of intermediates involved in perchlorate reductionis as follows:

(4.35)

In situ

bioremediation, via an IRZ, appears to be the most economically feasible,fastest, and easiest means of dealing with perchlorate-laden groundwater at allconcentrations. Microbial transformation of perchlorate to chlorite occurs in theabsence of oxygen as a result of anaerobic respiration. Anaerobic respiration is anenergy yielding process in which the oxidation of an electron donor, such as aneasily degradable organic substrate, is coupled to the reduction of an electron accep-tor, such as perchlorate and chlorate. Chlorite can be inhibitory to microbial activity,and the transformation of chlorite to chloride and O

2

is believed to be a nonenergy

ClO ClO ClO O Cl4 3 2 2- - - -

( )Æ

( )Æ

( )Æ +

( )perchlorate chlorate chlorite chloride

L1282_C04-C_frame Page 199 Thursday, March 6, 2003 11:42 AM


yielding enzymatic detoxification mechanism that protects the cell and allows thebacterium to use perchlorate and chlorate as electron acceptors.

90,91

Implementation of an IRZ with the introduction of easily biodegradable electrondonors such as hexoses, acetate, or lactate (without the presence of other electronacceptors such as SO

42–

) should be able to reduce the concentrations of ClO

4–

presentin the groundwater. A tenfold reduction of perchorate was achieved in columnexperiments at residence times of less than 48 hours. Laboratory column experimentshave demonstrated that perchlorate degradation can be achieved at influent levelsranging from 0.1 to 1000 mg/L.

91

The effluent levels achieved were in the range of0.005 mg/L which is lower than the state of California drinking water action levelof 0.018 mg/L. The author and his colleagues are currently involved in initiatingfield scale testing to implement

in situ

biodegradation of perchlorate.Since most of the perchlorate plumes are decades old and also due to its high

solubility, there are significantly large sized groundwater plumes of this contaminant.Hence, it is very important to select the cheapest electron donor to create an

in situ

reactive zone (IRZ) to achieve perchlorate degradation

in situ

. The persistence ofperchlorate itself is enhanced by the

oxidative

poise

available within these plumes.Hence, it is equally important to select the cheapest electron donor to overcome theoxidative poise within these large plumes.

4.3 ENGINEERED AEROBIC SYSTEMS

4.3.1 Direct Aerobic Oxidation

The majority of the compounds in petroleum products are biodegradable atsignificantly faster rates under aerobic conditions. The amount of oxygen requiredfor complete aerobic mineralization of one gram of hydrocarbon ranges from threeto three and a half grams. In simplistic volumetric terms, 300,000 kilograms ofoxygen-saturated water must be delivered and mixed in order to mineralize onekilogram of petroleum hydrocarbons. This illustrates the need to select the techni-cally and economically most effective method of delivering O

2

into the groundwaterand also to maximize the efficiency of O

2

utilization by the microorganisms in thesubsurface. The total cost of a pound of

dissolved

O

2

delivered into the subsurfacecould range from 0.80 to $10.00, depending on the method selected and the geologicand hydrogeologic conditions encountered at a site. The cheapest method of deliv-ering dissolved O

2

, if hydrogeologic conditions are conducive, is by injecting dilutehydrogen peroxide (at about 100–1000 ppm concentrations) into the contaminatedzone. Other methods of oxygen delivery include various methods of air injectionand expensive methods such as oxygen release compounds.

In addition to the petroleum hydrocarbons, other compounds more conducivefor aerobic biodegradation are: nonchlorinated phenolic compounds, alcohols,ketones, aldehydes, etc. Among the chlorinated compounds chlorobenzene, methyl-ene chloride, and vinyl chloride are among the commonly encountered contaminantsthat are biodegradable faster under aerobic conditions.

L1282_C04-C_frame Page 200 Monday, June 18, 2001 9:44 AM

IN SITU

REACTIVE ZONES 201

The most significant biological mechanism for the degradation of chlorinatedsolvents is when they are used as a primary substrate. In direct oxidation reactions,the chlorinated compound acts as an electron donor and the microorganism usesmolecular oxygen as an electron acceptor. The microorganism obtains energy andorganic carbon from the degraded chlorinated compound. The more chlorinatedcompounds, PCE, carbon tetrachloride (CT), and hexachloroethane (HCA), areneither susceptible to aerobic oxidation nor degraded under anaerobic oxidizingconditions when used as a primary substrate.

14

TCE undergoes slow aerobic degra-dation to trichloroethanol and then to acetic acid, but the reaction is not thermody-namically favorable. Therefore, discussion of aerobic oxidation and mineralizationhas always been focused on DCE and vinyl chloride (VC).

Rates of aerobic oxidation are more rapid for the less chlorinated organics (DCEand VC) when compared to their reductive dechlorination rates. It has been welldocumented in literature that VC is oxidized directly to carbon dioxide and water.Aerobic oxidation of

cis

-1,2 DCE has been speculated. However, it could not beascertained whether DCE was reduced to VC and then direct oxidation of VC producedcarbon dioxide or direct oxidation of DCE occurred to produce carbon dioxide.

Under aerobic conditions, chlorinated aliphatic compounds with one or twocarbons per molecule can be transformed by three types of microbial enzymes:

13

dehalogenases, hydrolytic dehalogenases, and oxygenases. Dehalogenases, whichrequire reduced glutathione as a cofactor, dehalogenate the substrates by means ofnucleophilic substitution. The first product of this degradation pathway is anS-choloralkyl-gluthathione, which is probably nonenzymatically converted toglutathione and an aldehyde. Hydrolytic dehalogenases hydrolyze their substrates,yielding alcohols. Oxygenases use molecular oxygen as a reactant for the attack onthe halogenated compounds; the products could be alcohols, aldehydes, or epoxides,depending on the structure of the compound. Numerous chlorinated short-chainaliphatic hydrocarbons have been demonstrated to undergo aerobic transformation.However, compounds that have all the available valences on their carbon atomssubstituted by chlorine, such as PCE or carbon tetrachloride, have never been shownto transform through any other but reductive pathways. Generally, as the degree ofchlorination increases, the likelihood of aerobic transformation decreases (Figure4.31); the opposite is true for anaerobic (reductive) transformations.

Among the methane compounds, methylene chloride (MC) and chloromethanehave been found to be amenable to aerobic microbial transformation. Pure culturesof the genera

Pseudomonas

and

Hyphomicrobium

have been isolated that can growon methylene chloride as the sole carbon and energy source.

13,15

Alkylhalides (haloalkanes), such as 1,2-dichloroethane (1,2-DCA), are fre-quently hydrolytically dehalogenated.

Xanthobacter autotrophicus

utilizes 1,2-DCAas sole carbon source. Complex communities consisting of methanotrophs andheterotrophs, which inhabit groundwater aquifers, mineralize 1,2-DCA. A

Pseudomonas fluorescens

strain isolated from water and soil contaminated by chlo-rinated aliphatic hydrocarbons was shown to utilize 1,2-DCA, 1,1,2-trichloroethane(1,2,1-TCA) and TCE, but not PCE or 1,1,1-TCA.

13,15



4.3.1.1 Aerobic Cometabolic Oxidation

Chloroalkanes, such as TCE,

cis

- and

trans

-1,2-DCE, 1,1,-DCE, and VC, arealso transformed by several different physiological groups of aerobes. Methan-otrophic communities consisting of methanotrophs that initiate the oxidative trans-formation, and heterotrophs which utilize the products of oxidation and hydrolysis,are very active in this respect, and can achieve complete degradation of chlorinatedalkenes. The same communities fail to transform PCE, however, because this com-pound is too oxidized. Pure cultures of methanotrophs such as

Methylosinus tricho-sporium

OB3b or

Methylomonas

sp. MM2, have been shown to partially transformTCE,

trans

-1,2-DCE, and

cis

-1,2-DCE.

13,14,15

Other microorganisms capable oftransforming chlorinated alkenes belong to the genera

Pseudomonas, Alcaligenes,Mycobacterium

, and

Nitrosomonas

. All of these microorganisms, except the genus

Nitrosomonas

, are heterotrophs which grow on various organic substrates (e.g.,toluene, cresol, phenols, propane, etc.);

Nitrosomonas

is a chemolitotroph whichderives energy from oxidation of ammonia. All of them cometabolize chlorinatedcompounds such as TCE or 1,2-DCE while growing on their respective growthsubstrates; the haloalkenes are only fortuitously transformed, not utilized for growth.However, vinyl chloride seems to be an exception: it has been demonstrated that a

Mycobacterium

strain isolated from soil contaminated by VC could grow on VC asa sole carbon and energy source.

92

Aerobic cometabolism of chlorinated compounds at low concentrations by meth-ane- and propane-utilizing bacteria is well documented. In comparison,

butane-utilizing bacteria

are less susceptible to the toxic effects of elevated chlorinated

Figure 4.31

Relative rates of oxidation and reduction from a range of C1 and C2 chlorinatedcompounds (adapted from Semprini et al., 1992).

EthenesEthanes

Increasing Extent of Chlorination

HCEPCETCE1,1DCE

TCA1,1,2TCA1,1DCA1,2DCE

1,2DCAVC

CAMethanesCTCFMCCM0

11

00

Rel

ativ

e R

ate

of R

educ

tion

Rel

ativ

e R

ate

of O

xida

tion


IN SITU

REACTIVE ZONES 203

compound concentrations. Butane is approximately four times more soluble ingroundwater than methane. Butane injection results in large radii of influence atinjection wellheads. The difficulty of utilizing alkanotrophic bacteria stems from thelow solubility of alkanes and the difficulty of maintaining homogeneous concentra-tion of the dissolved alkane within the reactive zone.

Methanotrophs grow on C

1

compounds as sole carbon and energy sources. Theircatabolic oxygenases are

methane monooxygenases

(MMO) that incorporate oneatom of oxygen from the oxygen molecule into methane to yield methanol.

12,14,92

This alcohol is further oxidized via a series of dehydrogenation steps, throughformaldehyde and formic acid, to CO

2

that is the final product of catabolism. MMOenzymes utilize molecular oxygen as a reactant, and require a reduced electroncarrier to reduce the remaining oxygen atom to water. MMO enzymes have relaxedsubstrate specificity, and will oxygenate many compounds that are not growth sub-strates for methanotrophs. Such compounds include various alkanes, alkenes, ethers,alicycles, aromatics, nitrogen heterocycles, as well as chlorinated alkanes, alkenes,and aromatics.

13,93

Two types of MMO have been suggested: a particulate (membrane-bound) anda soluble enzyme.

13

The soluble MMO (purified from

Methylosinus trichosporium

OB3b and

Methylococcus capsulatus

[bath]), produced under conditions of copperlimitation and increased oxygen tension, has been considered to have broader sub-strate specificity. It has been stated that only the soluble MMO can transform TCE.However, recent findings indicate that the particulate MMO in some methanotrophsmay be as effective in the transformation of chlorinated solvents as the solubleMMO. Since the soluble MMO is not constitutively expressed, whereas the partic-ulate MMO is, the latter methanotrophs (

Methylomonas

sp.) have a significantpotential for

in situ

bioremediation.Thus TCE can be transformed (upon the induction of the oxygenase enzyme by

its substrate) in the presence of the microorganismal growth substrate (cometabo-lism), or in its absence (resting cells transformation). However, TCE is not utilizedby the bacteria as a carbon, energy, or electron source; this transformation is onlyfortuitous. Based on the findings with methanotrophs, it can be concluded that TCEis most likely oxygenated to TCE-epoxide (Figure 4.32).

13,15,93

The epoxide is unsta-ble and is quickly nonenzymatically rearranged in aqueous solution to yield variousproducts including carbon monoxide, formic acid, glyoxylic acid, and a range ofchlorinated acids. Recent findings with purified MMO from

Methylosinus trichos-porium

OB3b indicate that TCE-epoxide is indeed a product of TCE oxygenation.In nature, where cooperation between the TCE oxidizers and other bacteria (mostprominently heterotrophs) occurs, TCE can be completely mineralized to carbondioxide, water, and chloride.

Toluene, phenol, and cresol oxidizers, such as

Pseudomonas putida

or

P. cepacia

,express the TCE transformation activity upon induction by their aromatic substrates.These bacteria have a great potential for remediation of groundwater aquifers con-taminated by mixtures of gasoline or jet fuel (or other petroleum derivatives), andchlorinated solvents, such as TCE, DCE, or VC. If the aromatic contaminants arenot present, however, bacterial growth substrates need to be injected into the site in



order to stimulate the transformation of chlorinated solvents. In this situation, meth-anotrophs become more attractive agents of bioremediation because methane, theirpreferred substrate, is a nontoxic and inexpensive chemical. Once methane andoxygen are injected into the site, methanotrophs (if present) will start cometabolizingchlorinated solvents, as well as a great number of other contaminants (see below),and the accompanying heterotrophs will mineralize their transformation products.As mentioned earlier it is important to maintain reasonably high and uniform O

2

and CH

4

concentrations to achieve significant methanotrophic degradation.

4.3.1.2 MTBE Degradation

There is a growing body of evidence from laboratory and field studies that MTBEcan be degraded under aerobic conditions either by direct metabolism (when MTBEserves as the carbon and energy source for microbial growth) or cometabolism.Evidence on natural attenuation of MTBE is presented in Chapter 3.

Microbes capable of MTBE degradation under aerobic conditions may be presentat most sites, but perhaps under nonoptimum biogeochemical conditions to signifi-cantly reduce the migration of MTBE. Furthermore, these aerobic processes wouldbe expected to be limited at many sites since the MTBE plume is migrating downa largely anaerobic path. Thus any approach to initiating or enhancing

in situ

aerobicbiodegradation of MTBE must overcome at least two major hurdles: 1) creatingsteady aerobic conditions over the long term and 2) generating enough microbialbiomass to accomplish the treatment at a reasonable rate.

In situ

chemical oxidation of MTBE has not been very successful due to theincomplete oxidation and formation of undesirable byproducts such as tertiary-butylformate, tertiary-butyl alcohol, methyl acetate, acetone, and formic acid. Hence,enhanced biodegradation of MTBE has to be optimized and engineered based onthe positive evidence found recently.

Figure 4.32

The top reaction shows how methanotrophs (“methane eaters”) produce theenzyme methane monooxygenase (MMO) in the process of converting methane(CH

4–

) to CO

2

. The bottom reaction shows how MMo then causes the conversionof TCE to CO

2

and HCl. NADH

2

serves as the carrier of electrons released frommethane and TCE. Note: NAD = nicotinamide adenine dinucleotide; NADH =reduced nicotinamide adenine dinucleotide.

Methane oxidation (normal reaction) with methane monooxygenase

TCE epoxidation (cometabolic dechlorination reaction) with MMO

H

C HH

H

OH

C HH

H

MMO

NADH , O2 NAD, H O23NAD, H O

23NADH2

CO2

2CO , 3Cl , 3H2- +

NADH , O2 2

NAD, H O2

CCl

MMO

2NAD, 3H O2 2NADH2

(other microorganisms)

CCl

Cl

HC

ClC

Cl

Cl

H

O


IN SITU

REACTIVE ZONES 205

A pure culture that degrades MTBE has been isolated;

94

however, the prolifer-ation of this organism in the natural environment may be questionable. Two recentfield tests, at Pt. Hueneme, CA, and Vandenburg AFB, CA, have provided positiveresults to the point that the previously held notion that MTBE is not aerobicallybiodegradable is not true anymore. At Pt. Hueneme, controlled testing was done inplots where pure oxygen only, and pure oxygen with bioaugmentation of enrichedcultures were injected in addition to a control plot where only natural conditionswere allowed to exist. The bioaugmented plot showed significant reductions ofMTBE at ppm range concentrations within a very short time period. The plot whereonly O

2

was injected also showed similar reductions in MTBE concentrations —after a lag period of months, however. The question still to be answered is whetherthe microorganisms responsible for MTBE degradation are obligate aerobes and thusa reasonably high DO concentration has to be maintained in the groundwater.

The Vandenburgh AFB field test also provided positive results and raised thepossibilities of implementing engineered

in situ

aerobic biodegradation of MTBE.In two separate long term field tests, dissolved oxygen was released to the MTBEplume through pressurized tubing via controlled interception trenches acting likepermeable walls. In both field tests, significant MTBE reductions took place in thepresence of increased levels of oxygen. MTBE degradation ceased when O

2

injectionstopped, thus indicating that degradation was conclusively aerobic.

Others have reported reductions in MTBE concentrations during

in situ

airsparging projects.

54

Stripping may be the dominant mechanism of MTBE removalfrom groundwater in these projects; however, the contribution by enhanced biodeg-radation due to increased levels of O

2

in the groundwater cannot be discounted,albeit at low levels compared to the mass removal by stripping.

At the Pt. Hueneme and Vandenburgh AFB field tests the means of O

2

injectionwas achieved by injecting pure O

2

. However, scaling up such a system to implementan engineered aerobic IRZ to address large MTBE plumes will be uneconomical,particularly when many of these plumes have migrated beyond property lines. Testingthe injection of dilute hydrogen peroxide to sustain the reasonably higher levels ofDO, which seems to be a requirement for aerobic MTBE degradation, will occursoon. Another means of providing enhanced levels of DO is through the implemen-tation of in-well sparging. Injecting hot air into the well will also enhance the massremoval by air stripping at reasonable air to water ratios. Recirculated water saturatedwith oxygen will create an

in situ

aerobic zone around the well and thus enhancethe aerobic degradation of MTBE (Figure 4.33).

Based on recent field observations, enhancing MTBE degradation withinengineered anaerobic zones may be a viable option. These zones have to be main-tained under methanogenic conditions.

4.4

IN SITU

CHEMICAL OXIDATION SYSTEMS

Chemical oxidation processes have been widely used for treatment of organiccontaminants in wastewaters. Because they are aggressive and applicable to a widevariety of compounds, using these processes, coupled with delivery technologies for



in situ

remediation of contaminated groundwater or subsurface soils, has receivedincreased attention.

In situ

chemical oxidation is an innovative technology withwidely varying opinions regarding its effectiveness on a range of contaminant types.The oxidants frequently used for this purpose are hydrogen peroxide, permanganate,and ozone.

In situ

chemical oxidation is achieved by delivering potent chemical oxidants tocontaminated media so that the contaminants are almost completely oxidized intoH

2

O, CO

2

, and chloride ions or converted into innocuous compounds commonlyfound in nature.

In situ

chemical oxidation will most likely be selected to addressremediation of what may be considered “difficult sites” having one or more of thefollowing characteristics:

77

low permeability soils, highly stratified soils, low-vola-tility target compounds, target compounds with low

in situ

degradation kineticconstants, and dense nonaqueous phase liquids (DNAPLS).

4.4.1 Advantages

The primary advantages of

in situ

chemical oxidation (ISCO) technologies is theirrelatively high speed destruction of contaminants. The cost of reagents is relativelyhigh compared to biological systems, so application is generally far more costly thanbioremediation systems, but significantly lower than other active source removal tech-nologies, such as

in situ

thermal treatment or flushing using surfactants or cosolvents.Since the reaction is nearly immediate, treatment is far more rapid than biologicaltechniques and can be faster than thermal or vapor recovery technologies.

Figure 4.33

Heated in-well sparging for enhanced stripping and aerobic biodegradation ofMTBE.

Air injectionpipe

Clean

water

Air totreatment

Heater Air compressor

Watertable

Packer

Contaminatedgroundwater

Enhanced biodegradation


IN SITU

REACTIVE ZONES 207

The advantages of

in situ

chemical oxidation can be summarized as follows:

• The ability to oxidize dense nonaqueous phase liquids (DNAPLs) if targetedproperly

• A reduction in overall treatment time, allowing the site to reach closure relativelysooner

• The elimination of capital intensive pump and treat systems• The ability to address contamination

in situ

without disturbing above groundstructures

In situ

chemical oxidation can be used as a stand alone treatment or in conjunc-tion with other technologies such as bioremediation. The nature and location of thecontamination, size of the source zone, type of soil, and hydrogeology play asignificant role in choosing the most effective type of ISCO treatment system. Insituations where contamination covers a vast area, economics will dictate the extentto which ISCO is used, but, in many cases, this is a cost effective pretreatment tobioremediation and natural attenuation.

4.4.2 Concerns

The primary concern is ensuring the health and safety of workers. Chemicaloxidation is an exothermic reaction generating heat that can increase temperatureand pressurize gases depending on loading and reaction rates. Strong oxidants arecorrosive and potentially explosive. The design and operation of any ISCO systemmust take into account the hazards of the chemicals and the potential for vigorous,uncontrolled, exothermic reactions in the subsurface. Site conditions that wouldwarrant particular attention in the planning stage include paved sites for which vaporpressures could build up under the pavement, sites with preferential flow paths, orutility corridors through which vapors could migrate.

A significant performance concern is that the oxidation reaction is not complete,and significant DNAPL accumulations remain in untreated areas in the subsurface.Even a small percentage of the original DNAPL mass can result in a rebound in thegroundwater concentrations after treatment to levels similar to those measured beforetreatment, or at least above levels of regulatory concern. In addition, the migrationof contamination to previously uncontaminated areas due to thermal gradients causedby exothermic reactions and to trapping contaminants in gas bubbles created by thereactions should be taken into account.

Another concern is the possibility of increased volatile emissions of volatileorganic compounds. Oxidation can cause significant heat generation and water vaporproduction. As a result,

in situ

steam stripping is a potential mechanism for contam-inant loss, particularly for highly volatile compounds like chlorinated solvents. Forexample, in cases where the hydrogen peroxide concentration exceeds approximately11%, enough thermal energy can be released to cause water to boil, leading to asignificant concern regarding vaporization losses.

95



4.4.3 Oxidation Chemistry

The oxidation chemistry of chlorinated solvents is relatively well understood.Oxidants attack the C-C bonds in these molecules. The double bonds that charac-terize chlorinated ethenes are far more reactive than the single bonds of chlorinatedethanes, and hence PCE and TCE are far more susceptible to oxidation than TCA,for example. However, the chloroethanes are often claimed to be susceptible tooxidation as well.

93,95,96

Current theory is that the oxidants cause formation of anunstable epoxide that then breaks down to yield ketones and aldehydes. Theseproducts may also be susceptible to further oxidation, eventually yielding carbondioxide, water, and chloride.

Several oxidants have been employed in the recent past for ISCO applications.For DNAPL sites, the most common oxidants used have been hydrogen peroxide(H

2

O

2

) and potassium permanganate (KMnO

4

). Permanganate is more expensivethan hydrogen peroxide, but it is also more stable and effective over a broad pHrange. Ozone (O

3

) is the strongest oxidant available, with an oxidation potential (E

°

)of 2.07 v. However, ozone is a gas and therefore most suitable for treating the vadosezone, or possibly LNAPL accumulations in the capillary fringe. Persulfate (S

2

O

8–2 )

salts are also available, with an E° of 2.01v, but these oxidants are relatively expen-sive and require thermal activation.93,95 The relative reaction kinetics of the differentoxidants are shown in Figures 4.34 and 4.35.

Hydrogen peroxide apparently works through two mechanisms: free radicalgeneration and direct oxidation. The direct oxidation has an E° of 1.76 v, and freeradical formation (H2O2 ⇒ 2OH· + 2H+ + 2e–) has an E° of 2.76 v. The latter relies

Figure 4.34 Relative strength of oxidants and relative resistance of some common contam-inants to chemical oxidation.

Resistance toOxidation

OxidantStrength

● Perchloroethylene

● Trichloroethylene

● Vinyl Chloride

● Phenanthrene

● Benzene

● Hexane

● Hydroxyl Radical

● Permanganate

● Ozone

● Hydrogen Peroxide

● Hypochlorite

● Oxygen

High

Low



on so-called Fenton’s chemistry, in which iron acts as a catalyst; therefore, iron isoften added with the hydrogen peroxide. In addition, pH adjustment is commonbecause oxidation is more rapid under acidic conditions.

Permanganate has an E° of 1.70 v and yields MnO2 as an insoluble precipitateunder most conditions.93,95 Catalysts and pH control are not needed for permanganateoxidation. The stoichiometry of complete oxidation reactions yields the followingweight ratios for permanganate (expressed as KMnO4:contaminant): PCE (1.3:1);TCE (2.4:1); DCE (4.4:1); and VC (8.5:1). Of course, this stoichiometry ignores thesignificant oxidant demand due to other reduced and natural organic compounds inthe subsurface, which can be significant.

Optimal use of the in situ chemical oxidation technology is very much dependenton understanding oxidant demand from contaminant oxidation and matrix oxidantdemand. Matrix oxidant demand refers to oxidant consumption that can be attributedto background soil and groundwater conditions (Figure 4.36). Matrix demand canbe derived from oxidation of natural organic matter (NOM), reduced metals, car-bonates, sulfides, etc. Matrix demand can be highly variable (depending on thereductive poise of the contaminated zone), influenced by background geochemicalconditions, and, since permanganate reaction rates are second order, also will dependon the permanganate solution concentration.

The oxidant demand caused by the nontarget compounds can range from 10 to100 times (or even higher) of the stoichiometric demand caused by the targetcontaminants. Hence, it is more important to look at the chemical oxidation demandof the system than at the total organic carbon (TOC) as an evaluation parameter forchemical oxidation.

It should be noted that destruction of natural organic matter can release additionalcontaminants, adsorbed to the organic matter, into the dissolved phase (Figure 4.37)before being completely destroyed by the oxidant. This phenomenon is the primaryfactor contributing to the rebound effects of the target contaminants during chemicaloxidation.

The most commonly observed mobilization of metals, during ISCO, is oxidationof precipated Cr3+ to the dissolved Cr6+. The amount of Cr6+ mobilized will obviouslydepend on the background chromium concentration in the soils. Literature reportsindicate that this dissolved Cr6+ will reattenuate within a short time frame and distance.

The advantages of peroxide as an oxidant include relatively low regulatoryresistance, more field experience in its use than permanganate, and a sparcity of

Figure 4.35 Relative reaction kinetics of various oxidants.



Figure 4.36 Oxidative poise of natural environment and increased potential demand ofoxidants.

Figure 4.37 Natural organic matter destruction releases additional dissolved phasecontaminants.

Progress of treatment

Chemical oxidationdemand

Natural and non-targetorganic matter

Initial dissolved phasetarget contaminant

NO

Mde

stru

ctio

nef

fect

Residual level of NOM



byproducts of oxidation. Disadvantages include the need for pH control in somecases and difficulties in controlling in situ heat and gas production. Permanganateis more expensive and more stable than peroxide, and is effective over a broad pHrange. Oxidation also produces manganese levels, which will precipitate and poten-tially cause reduced porosity. Increases in dissolved manganese levels are also apotential regulatory concern depending on the geochemistry, as is the purple colorof groundwater containing unreacted permanganate. Ozone has been used mostlyfor vadose zone treatment. It is less costly than permanganate or peroxide, but themost significant factor in choosing ozone is that it must be applied as a gas. Gasesmay disperse further in the unsaturated zone than liquids, but vapor recovery andpossible treatment can add considerable cost.

4.4.3.1 Hydrogen Peroxide

Hydrogen peroxide (H2O2) is typically used together with Fe (II) to form Fenton’sreagent. In Fenton’s reagent, H2O2 is decomposed by Fe (II) to produce highlyreactive hydroxyl radicals as expressed by Equation 4.36:

Fe2+ + H2O2 ⇒ Fe3+ + OH• + OH– (4.36)

The hydroxyl radical can nonselectively attack the C-H bonds of organic mole-cules and is capable of degrading many solvents, chloroalkenes, esters, aromatics,and pesticides. The major advantages over other oxidation processes of using Fen-ton’s reagent to treat hazardous wastes can be summarized as:95 1) there are nochlorinated organic compounds formed during the oxidation process as in chlorinat-ing; 2) both iron and hydrogen peroxide are inexpensive and nontoxic; 3) there areno mass transfer limitations because the reaction is homogeneous; 4) no light isrequired as a catalyst and, therefore, the design is much simpler than ultraviolet lightsystems; and 5) H2O2 can be electrochemically generated in situ, which may furtherincrease the economic feasibility and effectiveness of this process for treating con-taminated sites. During treatment, particulates can be generated and the pore sizeand continuity can, therefore, be modified with fine-grained media. As a result, thepermeability can be impacted.

In Fenton’s mechanism, reactions with H2O2 cycle iron between the +II and +IIIoxidation states, yielding OH• and other byproducts. Because OH• is a powerfulindiscriminate oxidant that reacts with many compounds at near diffusion-controlledrates,97,98 H2O2 and iron have been used to generate OH• and oxidize undesirablecontaminants in soils and aquifers.93,95 A wide range of organic compounds (TCE,BTEX, PCP, naphthalene, and pesticides) that are common contaminants of ground-water and soil have moderate to high reaction rate constants with OH• (108 –1010M–1s–1). The stability of H2O2 increases with decreasing pH in Fenton systems,and oxidation efficiency is optimum under acidic conditions.95,97

Under acidic conditions and with an excess of Fe2+, the hydroxyl radical gener-ated can further react with Fe2+ to produce Fe3+ (Figure 4.38a):76

Fe2+ + OH• ⇒ Fe3+ + OH– (4.37)



By properly controlling experimental conditions, ferric iron can be regeneratedback to ferrous iron by a subsequent reaction with another molecule of H2O2:

Fe3+ + H2O2 ⇒ Fe2+ + HO2• + H+ (4.38)

The HO2• radicals produced (Equation 4.38) have been shown also to participate

in oxidation of some organic compounds, although they are much less reactive thanOH•. Based on Equation 4.38, a low pH range of 2 to 4 is preferred to facilitate thegeneration of hydroxyl radicals, although the reaction is feasible up to neutral pH.99

Almost all organic compounds can be treated in situ by this technology.Limitations to Fenton-based remediation strategies arise from excessive H2O2

decomposition via nonproductive reactions (those that do not result in OH• pro-duction), reaction of OH• with nontarget species (scavenging), insufficient iron orH2O2 for radical production, and slow reaction of OH• with the target com-pound.93,95 For example, REDOX cycling of manganese between the +II and +IVoxidation states consumes H2O2, but does not yield OH•. Common groundwateranions (NO3

– , SO42– , C1–, HPO4

2– , HCO3– , CO3

–2 ) react with OH• and may be asource of treatment inefficiency. Furthermore, because H2O2 is generally presentat high concentrations in Fenton systems and has a moderate rate constant forreaction with OH• (2.7 × 107M–1s–1), peroxide is itself a primary source of ineffi-ciency in Fenton-driven systems.95

Figure 4.38a Fenton’s reagent — idealized reactions.76

OX RED

2H O2

-OH + OH-

Fe3+

+Fe2

OX

RED

OX RED

OH-

RH

2R'H + CO

-H +R+



Idealized reactions of Fenton’s reagent and potential reduction in efficienciesdue to disproportionation are shown in Figures 4.38b and c, respectively.76 Thecontaminants of particular interest include chlorinated solvents (e.g., TCE, PCE),polyaromatic hydrocarbons (e.g., naphthalene), PCP, and petroleum products (e.g.,BTEX). Some of these chemicals are very difficult to biodegrade or may takeexceedingly long time in many subsurface settings.

Major concerns for this technology are related to potential ecological effects andchemical handling. The introduction of acid solution can have potential effects onthe ecosystem. During the reactions, both OH– and H+ can be produced; however,their quantities are relatively small compared with the acid introduced and thuswould have no significant effect on the pH of the media. Because large quantitiesof chemicals are required for the treatment, it could be hazardous to handle them.In addition, special measures may be taken during the delivery process because H2O2

can easily break down into H2O vapor and O2, leading to fugitive emissions of VOCsand pressure buildup. One benefit of decomposition of H2O2 is that the released O2

can stimulate aerobic biological activity.

4.4.3.2 Potassium Permanganate

Potassium permanganate (KMnO4) has been used in treatment of drinking waterand wastewater for decades because it can effectively oxidize many water impurities,including phenol, Fe2+, S2–, and taste and odor-producing compounds. Only withinthe past few years has it been used more frequently as an alternative chemical oxidantfor ISCO. KMnO4 is a dry crystalline material that turns bright purple when dissolved

Figure 4.38b Fenton’s reagent failures — Fe3+ catalyzes disproportionation.76

H O + O2 2

OX

RED

2H O2

-OH + OH-

Fe3+

+Fe2



in water. The purple color acts as a built-in indicator for unreacted chemical. ReactedKMnO4 is black or brown, indicating the presence of the MnO2 precipitate — anatural compound present in soil. Other KMnO4 oxidation byproducts include CO2,H2O and the potassium ion K+. Limitations of KMnO4 include its low solubility (65g/l at 68°F) and its inability to oxidize petroleum compounds effectively.

Sodium permanganate (NaMnO4) is an oxidant that performs very similarly toKMnO4; its attributes and limitations are much the same as KMnO4. However,NaMnO4 has a much higher solubility in water, allowing it to be used for ISCO at amuch higher concentration. NaMnO4 is more expensive than KMnO4 on a pound-per-pound basis and users have to be concerned about safety during handling and storage.

Reaction of KMnO4 with organic compounds produces manganese dioxide(MnO2) and carbon dioxide or intermediate organic compounds. The kinetics ofreaction between permanganate and contaminants are obviously an important factorin the overall treatment success achieved. It has been reported that oxidation of TCEby KMnO4 is second order, with a fast second order constant.

An apparent limitation with the reactive hydroxyl radical (OH•) is that it stronglyreacts with common inorganic species in groundwater such as carbonate and bicar-bonate. However, permanganate, a metal-oxo reagent, does not apparently rely ongenerating a hydroxyl radical to oxidize chlorinated ethenes as the other oxidantsdo. Experience indicates that metal-oxo reagents can attack a double carbon-carbonbond powerfully through direct oxygen transfer.76

[Org] + MnO4– → MnO2 + CO2 or [Org]ox (4.39)

Figure 4.38c Fenton’s reagent failures — Fe3+ is lost to precipitation.76

Precipitate

OX RED

2H O2

-OH + OH-

Fe3+

+Fe2



where [Org]ox is the oxidized intermediate organic compound. Permanganate ionspreferentially attach carbon-carbon double bonds, in a manner similar to the attackof ozone.76 A manganate ester forms in the first stage of the reaction and rapidlydecomposes to form a glycol, as shown in Figures 4.39a and b. Manganese dioxide(MnO2) precipitates from the oxidizing aqueous solution. The glycol is cleaved underhigh permanganate concentration or acidic conditions to form aldehydes or ketones.Aldehydes are likely to be further oxidized to carboxylic acids.76,96

When permanganate is used to oxidize chlorinated ethenes, chlorinated interme-diates such as phosgene or formyl chloride might be produced. However, itwas observed that rapid dechlorination of the manganate ester took place when

Figure 4.39a Permanganate oxidation of an alkene.

Figure 4.39b The oxidation of ethylene in a neutral to weak acidic condition.

REDOX

MnO O

H H

O O

CC

H+ +HMn

OO

OO

C C

-

CC

O O

O O

Mn

CH MnO

CH

H

H

4

EthyleneH

H

HC

HC

OOMn

O O

Cyclic HypomangnateEster

H O

C-C bond

fragmentation

OH O

CH

CH H

Glycol Aldehyds

4MnO

H OCC

OOH

O HGlyoxylic Acid

H O

MnO 4

Oxalic AcidOHO

OH O

C C

O

C

Formaldehyde

H HH O

OHH

Formic Acid

C

O



permanganate ion was used to dechlorinate TCE and other chloroethenes.76,93,95,96

For tests run at pH ranging from 4 to 8, oxidation of the manganate ester to carbondioxide was more rapid than the permanganate attack on the solvents. It was alsonoted that only the permanganate ion, MnO4

– , participated in oxidation, and thatmanganese dioxide (MnO2) was the only manganese-bearing product of the severalreactions.

Several studies have been published describing permanganate oxidation of chlo-rinated ethenes, including reports of both field and laboratory applications.76,93-95 Acommon element of these studies is the focus on oxidation of the contaminant com-pounds, without evaluation of oxidation byproducts that may result from reaction ofpermanganate with naturally occurring compounds and organic species associated withsolvent wastes.

One particular problem in laboratory studies is that permanganate is typicallyapplied as an excess reagent — an approach that simplifies analysis of reactions thatare first-order in both reactants. But the excess permanganate oxidizes many potentialreaction byproducts. In field applications, permanganate cannot be applied as anexcess reagent across the entire aquifer and the appearance of ketones, aldehydes,and other reaction byproducts cannot be ruled out.95

Even when permanganate is applied as an excess reagent, byproducts such asacetone and butanone may accumulate during oxidation of contaminated aquifersoils. In an unpublished bench-scale study, a 3% potassium permanganate solutionwas applied to aquifer soils contaminated by an oil-solvent mixture. The applicationcontinued until permanganate depletion, during passage through the aquifer soils,to negligible levels. Newly formed acetone, 2-butanone and other oxidation productswere measured in aqueous-phase samples throughout the test application.76

The compounds that can be oxidized by permanganate in addition to alkenesinclude aromatics, PAHs, phenols, pesticides, and organic acids. The optimum pHrange is 7 to 8, but they are effective over a wide range.

Because Mn is an abundant element in the Earth’s crust and MnO2 is naturallypresent in soils, introduction of KMnO4 to soils as well as production of MnO2

would not be an environmental concern. KMnO4 is as effective as or more effectivethan H2O2 in oxidizing organic compounds. Furthermore, KMnO4 is more stableand easier to handle. The potential problem is that MnO2 particles will be generatedand permeability loss is possible.

4.4.3.3 Ozone

Like hydrogen peroxide and permanganate, ozone is a strong oxidant that canquickly oxidize organic compounds once in contact. Compared to other technologies,in situ ozonation offers several advantages:76,93

• It is much easier to deliver ozone to the contamination zone than aqueous oxidants.• No volatilization of target chemicals is required and, therefore, mass transfer

limitations associated with soil venting can be overcome.• In situ ozonation would likely be more rapid than biodegradation or soil venting

processes, and thus reduce the remediation time and treatment costs.



• Ozone can be electrically generated from air on site.• In situ ozonation is conceptually similar to soil venting processes.• Both vertical and horizontal wells can be used to inject ozone.• Little degradation of ozone occurs during injection and on-site handling is rela-

tively easy.• Similar to H2O2 and permanganate, ozone can be used to treat a variety of organic

compounds.• Ozone is very reactive and corrosive to materials.

Ozone reacts quickly in the subsurface and does not migrate long distances fromthe point of delivery. Currently, ozone is used to treat chlorinated solvents, polyar-omatic hydrocarbons, and petroleum products in situ.

Ozone is unstable in water. The degradation of ozone involves a complex cyclicprocess that may be promoted or inhibited by various substances. In natural systems,the degradation of ozone may be initiated by various substances including thehydroxide ion OH–, and natural organic matter (NOM). Bicarbonate and carbonateions and other hydroxyl radical scavengers will inhibit the degradation by ozone.Many organic compounds are able to initiate, promote, or inhibit the chain-reactionprocesses of ozone decomposition and degradation.

Zwitterions, also known as dipolar ions, are neutrally charged but stronglypolarized molecules that behave as ions.76 Many molecules exhibit a degree of dipolarbehavior — zwitterions can be sufficiently dipolar to confer substantial reactivity.Another zwitterion behavior is exemplified by glycine, an amino acid:glycine actsas a base when titrated with acid, and acts as an acid when titrated with base.

Ozone is a zwitterion comprised of three oxygen atoms, as shown in Figure 4.40.A resonant double bond concentrates negative charge in the terminal oxygen atombound by the single bond.76 Although the diagram suggests a concentration ofpositive charge in the central portion of the molecule, the central atom exerts a pullon the electrons from the resonant double bond, transferring some of the positivecharge to the double-bonded terminal oxygen.

The double-bonded terminal oxygen atom in ozone can initiate electrophilicattack on carbon-carbon double bonds, as shown in Figure 4.41. As the electron pairfrom the alkene migrates toward the electrophilic oxygen atom, the opposite carbonatom becomes electrophilic, attracting the singly bonded oxygen atom into a molo-zonide bridge. This highly unstable compound breaks and reforms as an ozonide,

Figure 4.40 Zwitterion behavior.

-

+O

O OO O

O+

-



which also decomposes spontaneously. The reaction is completed by the formationof two ketones (or aldehydes) and water. The kinetics of ozone attack on chlorinatedethenes is highly influenced by the steric hindrance caused by chlorine atoms.76

The dramatic increase in reactivity to ozone from tetrachloroethene to trichlo-roethene is due to two factors: the reduction in steric hindrance that follows fromelimination of a chlorine atom, and the reduction of the carbon atom from C (II) toC (I), making the electron pair more available to electrophilic attack (oxidation).Reaction rates of ozone and simple alkenes such as styrene are very high, whilealkanes, alcohols, aldehydes, and ketones are only slightly reactive to ozone(Table 4.11).

The final decomposition products of the ozonation of chlorinated ethenes areformaldehyde (CH2O), and phosgene (CCl2O). Formyl chloride (CHClO) is a theo-retical product which is unreported in the chemical literature and presumable is unsta-ble. Phosgene decomposes rapidly in water and is not expected to be observed; andformaldehyde rapidly biodegrades in the highly aerobic post-ozonation environs.76

4.4.4 Application

In general, more than a single application of oxidant is required to meet mostcleanup standards. Several reinjections at periodic intervals have been used for morethorough treatment. Recently, continuous injection using recirculation of amendedwaters has been used to maximize the utilization efficiency of the oxidant as well

Figure 4.41 Ozonation of an alkene.

OC

O C

O

O C

O

OC

O

OO

C CCC

O O

O



as to augment the distribution rate within the reactive zone. A comparison of theproperties of the three commonly used oxidants is presented in Table 4.12.

For single or multiple injections, permanent or temporary injection points areestablished, and an aqueous solution containing the oxidant and any needed catalystsis injected under pressure. The oxidant (and catalyst) concentration, the target pH,the injection well spacing (i.e., radius of influence), the number of injections, andthe injection pressure are all important design parameters affecting cost and perfor-mance. The oxidation reactions occur in the aqueous phase, and NAPL and sorbedphases must be targeted and treated either by interfacial contact with or mass transferto the aqueous phase (Figure 4.42).

Successful prediction of overall rates of mass removal would require rate expres-sions both for nonequilibrium dissolution and oxidation. In the conceptual model(Figure 4.43a), dissolution mass transfer, driven primarily by aqueous phasechlorinated contaminant concentration gradients, is enhanced by the oxidation reac-tion that increases these gradients (Figure 4.43b). The efficiency of chemical oxi-dation for treatment of NAPLs is based on the conceptual model that attributes an

Table 4.11 Kinetic Constants for Ozonation of Various Organic Compounds76,98

Compound Ko3 (M–1s–1)

Tetrachloroethene 0.1Trichloroethene 171,1-Dichloroethene 1101,1-Dichloroethane 0.12Cis-1,2-Dichloroethene 800Trans-1,2-Dichloroethene 5700Styrene 3 × 105

Formaldehyde 0.1Acetaldehyde 1.5

Table 4.12 Comparison of Oxidants

Fenton’s Reagent Permanganate Ozone

Physical State Liquid Liquid GasMolecular Composition OH• MnO4

– O3

OH• formation Yes Under Very Limited Conditions

Under Certain Conditions

Oxidation Potential 2.76 V 1.70 V 2.07 VReaction Times Very Fast Slow FastContaminant Range Many Organics Few Organics Some OrganicsPotential to Entrance Yes If the pH

Conditions Are Not Low

Unlikely Yes

Metal Mobilization/Potential Yes Cr3+ → Cr6+ Yes



increased rate of DNAPL mass transfer to chemical oxidation within the stagnantfilm boundary layer. As the aqueous solvent gradient is increased, the dissolutionmass flux is increased. Simultaneously, the concentration gradient of the oxidantwould be increased, causing an increase in oxidant mass flux towards theDNAPL/water interface.

Figure 4.42 Chemical oxidation systems to address DNAPLs will have to target the mass ofDNAPL.

Figure 4.43a Conceptual model describing mass removal by in situ oxidation.

Pooled DNAPL

Adsorbed DNAPL

Dissolved Phase

Concentration

DNAPL Phase

CO Without Oxidant

WithOxidant

OxidantConcentrations

Mass Flux



Permanganate oxidation of a DNAPL can yield manganese oxide solids that maydeposit on the interface and could result in a reduced interface mass transfer rateand DNAPL oxidation rate. This is a complex process that is not fully understood.

The use of recirculation, with injection and extraction wells, is intended toincrease subsurface mixing. Many investigators have tried this approach with someapparent success. The costs are likely to be higher than even multiple injectionswithout groundwater extraction and reinjection (with possible treatment and filtrationof MnO2 required). However, the degree of mixing and, therefore, contact betweencontaminants and oxidant, will be greater, leading to more complete treatment,especially in heterogeneous subsurfaces. Also, utilization efficiency of the oxidantwill be enhanced by recirculating the unused portion of the oxidant.

In some cases, mixing has been encouraged by use of injection arrays with thinscreen intervals at different depths to fully saturate the target zone and limit the needfor vertical migration of the oxidant (Figure 4.42). High injection pressures havealso been used to create fractures in tighter subsurface materials, again to encouragemigration and mixing of the reactants. Mixing has also been encouraged throughthe use of air injection, to “push” peroxide solutions out into the aquifer. Finally, insome cases, vapor extraction has been used in conjunction with in situ oxidation inthe vadose zone to relieve off-gas pressures, to encourage oxidant migration, and/orto capture any volatile emissions. Oxygen concentrations in the soil air can reachclose to 100% and thus create explosive concentrations near the points of injection.76

The presence of colloidal materials, precipitation, and gas binding can causereduced permeability of the aquifer near injection points. If the geologic materialshave excessive amounts of CaCO3 in the formation gas, binding during the injectionof Fenton’s reagent could be a significant problem.

There do not seem to be well-developed guidelines for the design operation andcost estimation of ISCO systems particularly when DNAPL is present (Figure 4.44).The data needs for determining well spacing, screen intervals, or oxidant mass tobe injected are not clear. There is a need for guidance to estimate the ROI underdifferent conditions (soil texture, groundwater velocity, injection pressure, etc.). Theefficiency of use of oxidants is not well established, and guidance for determiningthe mass needed at a specific site does not seem to be available. Recommendations

Figure 4.43b Conceptual model of interface mass transfer effects of chemical oxidation.

DNAPL GW / Oxidant solution

Boundary layer



regarding operations and monitoring to prevent undesirable reactions (explosions,volatile emissions, or foaming) are also not clear.

4.4.4.1 Oxidation of 1,4-Dioxane by Ozone

Ozone is a powerful oxidant that can degrade contaminants via two mechanisms.The first, commonly referred to as the direct mechanism, involves the reaction ofmolecular ozone with the contaminant. Secondary oxidants, particularly the hydroxylradical, can also oxidize the contaminants present. The oxidation of compounds byhydroxyl radical or other secondary oxidants is referred to as an indirect oxidation(since ozone is not directly involved in the oxidation). OH• is a nonselective oxidant,that is, it oxidizes many substances; consequently, in natural systems it may not bea very efficient oxidant, as it will react not only with contaminants of interest, butalso with other substances present, e.g., the natural organic matter.

In waters that contain carbonate, hydroxyl radical scavenging is greater at higherpH. Therefore the effectiveness of ozonation systems tends to decrease at elevatedpH. For substances such as 1,4-dioxane that are not very reactive with molecularozone, the optimal pH for removal is typically around 8.76

Ozone could be used to treat groundwater extracted from the aquifer. For ex situtreatment systems a treatment time of approximately 20 minutes would be neededfor 99% removal of 1,4-dioxane.76

The use of ozone for in situ treatment of groundwaters may be particularly usefulfor the treatment of contaminants that are strongly sorbed to the aquifer materials(e.g., PAHs) or where the aquifer materials exert relatively little ozone demand.1,4-dioxane is infinitely soluble in water and is not strongly sorbed to solids. In thiscase, in situ ozonation may be desirable as it avoids the cost of removal of the

Figure 4.44 Conceptual description of increased remediation cost due to the presence ofDNAPL.

Conceptual Remediation End-point

Reduction of aqueous phase concentrationsto below regulated clean-up levels

Complete DNAPL removal

Reduction of aqueous phase mass flux

Stabilization of pooled DNAPL

Partial DNAPL removal

Deg

ree

of M

ass

Rem

oval

0%

100%



groundwater and of the expense and problems associated with construction of anozone contractor to treat the extracted groundwater at the site. In situ ozonation canbe implemented using techniques developed for in-well air sparging.

The relative cost of in situ vs. ex situ treatment will depend very much upon howeffectively these systems can be implemented. Ozone is sometimes used in combina-tion with UV light or hydrogen peroxide to treat groundwaters. Both UV light andhydrogen peroxide catalyze the decomposition of ozone to produce hydroxyl radical.The rapid decomposition of ozone can enhance the rate of degradation of compoundslike 1,4-dioxane, which are not very reactive with molecular ozone.

4.4.4.2 Biodegradation Enhanced by Chemical Oxidation Pretreatment

Many experimental efforts have been carried out to evaluate the enhanced biodeg-radation of many recalcitrant compounds such as PCBs, polychlorinated phenols, andPAHs — with limited success. With increased attention to the cleanup of sites withknown DNAPLs and manufactured gas plant (MGP) sites with coal tars, pretreatmentwith chemical oxidation for certain compounds may be a viable technology.

4.5 NANO-SCALE FE (0) COLLOID INJECTION WITHIN AN IRZ

Considerable research during the past several years has focused on the transfor-mation of chlorinated solvents to harmless end products by exploiting the use ofzero valent elemental metals for reductive dechlorination. In addition, elementalmetals can be used to reduce soluble metals such as Cr (VI) to insoluble Cr (III) ormetalloids such as As (V) and Se (VI) to As (III) and Se (IV), respectively.

The most common metal utilized for this purpose is elemental iron, Fe (0).Although met with initial skepticism, the transformation process is surface-basedand is now widely accepted as abiotic reductive dechlorination, involving corrosionof Fe (0) by chlorinated hydrocarbon. Other metals including tin, zinc, and palladiumhave also been shown to be effective.100 The process can be described best asanaerobic corrosion of the metal by the chlorinated hydrocarbon. During this process,the contaminant is adsorbed directly to the metal surface where the dechlorinationreactions occur.

In waters contaminated with chlorinated solvents, three oxidants are availableto drive corrosion of metals: water, dissolved oxygen, and the chlorinated contam-inant. The corrosion reaction involving water (Equation 4.40) is slow but presumablyubiquitous, whereas corrosion of Fe (0) by reaction with dissolved oxygen (Equation4.41) is very rapid as long as O2 is available. The reaction rates with the chlorinatedcontaminant (Equation 4.42) are assumed to be between the two. Under aerobicconditions, dissolved oxygen is usually the preferred electron acceptor and willcompete with the chlorinated contaminant as the favored oxidant (PCE and carbontetrachloride may be comparable).

Fe (0) + 2H2O → Fe2+ + H2 + 2OH– (4.40)



2Fe (0) + O2 + 2H2O → 2Fe2+ + 4OH– (4.41)

Fe (0) + RX + H+ → Fe2+ + RH + Cl– (4.42)

When sufficient oxygen is present, the Fe2+ generated in Equation 4.40 canprecipitate as ferric hydroxide or (oxy) hydroxides at an elevated pH typical ofcorroding Fe systems. In carbonate rich waters FeCO3 precipitation will alsooccur. These precipitates can exert significant additional chemical and physicaleffects within the surface-based Fe (0) reactive system by coating the reactiveiron metal.

Recent research on Fe (0) systems indicates that other mechanisms also may beinvolved in the reductive process. The reductive processes can be summarized asbelow:100

• Fe (0) can act as a reductant by supplying electrons directly from the metal surfaceto the adsorbed chlorinated contaminant (Figure 4.45).

• Metallic Fe (0) may act as a catalyst for the reaction of H2 with the chlorinatedcontaminant. The hydrogen is produced on the surface of the iron metal as theresult of corrosion with water (Figure 4.45).

Figure 4.45 Abiotic reductive dechlorination mechanisms by Fe (0).

Acetylene CC HH

H ClC CChloroacetylene

TCE

H

Cl Cl

Cl

2e-

-2Cl

B

A

-2e + H +

Cl -H

ClCl

H

cis-1,2-DCE

H

H Cl

H

Vinyl chloride

+2e + H-

-Cl

+2e + 2H-

-2e + 2H +

Cl -

H

HH

H

Ethene

+2e + 2H-

H

H H

H

Ethane

H

H

+2e + H-

Cl -


IN SITU

REACTIVE ZONES 225

The rate of reaction of Fe (0) with chlorinated contaminants is dependent uponthe reactivity of individual chemical compounds and the amount of reactive surfacearea on the Fe (0) particles. For degradation of a contaminant by zero-valent ironmetal, the reaction model can be represented as:

102

(4.43)

where

C = reacting contaminant concentrationK

SA

= specific reaction constant

r

a

= the amount of iron surface area

Based on the reported values of K

SA

in the literature, there seems to be an orderof magnitude variability for an individual chlorinated hydrocarbon.

102

It is importantto note, however, that the variability in K

SA

for individual compounds is modest relativeto the five orders of magnitude variability among the various chlorinated hydrocarbons.

In addition to the primary effects of contaminant reactivity and metal surfacearea, several other factors influence the kinetics (K

SA

) of chlorinated contaminantdegradation. One factor is the saturation of reactive surface area with increasingcontaminant concentration. Another factor is metal “type,” which is the variablemost commonly invoked to rationalize otherwise unexplained variability in degra-dation rates by iron. The

r

a

term in Equation 4.43 characterizes quantity of ironsurface area, but does not address differences in the reactivity of the surface. It isimportant to note that as the size of the metallic iron is reduced, surface area goesup as well as chemical reactivity.

High surface areas can be attained either by fabricating smaller particles orclusters where the surface to volume ratio of each particle is high, or by creatingmaterials where the void surface area (pores) is high compared to the amount ofbulk material. If a metal is continually reduced in size it will eventually reach whatis known as superfine particle or nano-scale particle. Such particles can be distin-guished from their corresponding bulk solid form by the size of their surface areasin relation to their weight.

Initial applications of this technology in the mid 1990s used iron filings. Due tosize limitations (not small enough to be injected directly) of the commerciallyavailable iron filings (Table 4.13), the process had to be implemented in the subsur-face as a permeable reactive barrier (PRB). In a PRB, reactive material is placed inthe subsurface where a plume of contaminated groundwater must move through itas it flows, typically under its natural gradient, and treated water comes out the otherside. The placement of the iron filings into the PRB was usually achieved byhydraulic fracturing (Figure 4.46), or via a funnel and gate system where the gatewas filled with iron filings, or by mixing the iron filings with sand in a permeableinterception trench (Figure 4.47). It is obvious that in all these methods the “periph-eral” geotechnical cost for the “placement” of iron filings in the subsurface can be

- = [ ]dCdt

K CSA ar

L1282_C04-C_frame Page 225 Thursday, March 6, 2003 11:45 AM


up to two orders of magnitude higher than the actual cost of the iron filings. As aresult, more recent applications have used iron colloids in the micron size range tocut down on the peripheral geotechnical cost and directly inject the iron colloidsinto the contaminated zone.

The author and a few others have advanced metallic Fe (0) reduction technologyby incorporating nano-scale particles ranging in size from 1 to 999 nanometers (.001to .999 µm). A particle of this size has several advantages in application for in situgroundwater remediation. These advantages include:

• High surface area result in greater reaction kinetics.• The increase in kinetics allows for a lower mass loading of iron in the treatment

zone or reactor because the residence time required for complete dechlorinationis decreased.

• The small size and greater reactivity of the superfine particle allows for theapplication of the technology through direct in situ injection into the subsurface(Figures 4.48 and 4.49).

Table 4.13 Examples of the Surface Area of Different Metallic Iron [Fe (0)] Products103-105

Iron Type Surface Area in M2/g

Iron turnings 0.019 M2/gElectrolytic iron 0.057 M2/gIron granules 0.287 M2/gCommercial iron filings 0.900 M2/gNano-scale iron particles 33.50 M2/g

Figure 4.46 Placement of Fe (0) filings as a reactive barrier via hydraulic fracturing.



Figure 4.47 Placement of Fe (0) filing as a permeable reactive barrier in an interceptor trench.

Figure 4.48 Direct injection of nano-scale Fe (0) particles into the contaminated zone.

Groundwater flow Treated Groundwater

Interceptor Trench

Sand/Fe

mixture

oContaminatedplume

Injection well

In Situ reactive zonefilled with nano scale iron

Fe / Molasses slurryo

Contaminated zone



• The smaller size allows for advective particle transport.• The greater reactivity due to the small size allows for much lower overall iron

mass requirements.

Conceptually, destruction of the contaminant is an interfacially controlled pro-cess, and thus the efficacy of destruction is dominated by the exposed surface areaof the superfine particle. The exposed surface area is easily determined by BETnitrogen adsorption, for which the surface area can be related to an equivalentspherical diameter (desd):

SSA = 6/ρ.desd (4.44)

where

SSA = specific surface area determined by BETρ = material density

In addition to the beneficial effects of increased surface area of the superfineparticles, coupling of a catalyst such as palladium or platinum will lead to increasedreaction rates which are multiplicative (Figure 4.50).

4.5.1 Production of Nano-Scale Iron Particles

Over the last decade, research, driven primarily by needs in the field of materialsscience (hi-tech electronic chips or component industry products), has contributed

Figure 4.49 Injection of nano-scale Fe (0) particles into the contaminated zones in theunsaturated and shallow saturated zones.

oMixed zone with Fe

In Situ mixing



to general technologies designed to produce nano-scale particles. Generally, theresearch has been in the area of colloids composed of ceramic or other nonmetallicinorganic materials and not metal colloids. A significant part of the developmenteffort for the technology is the adaptation of nonmetallic nano-scale productionmethods to the production of metallic nano-scale particles.

The method for production of metal particles in the nano-scale range may bedivided into two primary approaches: 1) “bottom up,” in which colloids of theappropriate size are produced by being assembled from individual atoms; and 2)“top down,” in which colloids of the appropriate size are produced by attrition oflarger existing particles of the metal.

The bottom up approach has a greater number of potentially applicable methodsincluding100:

• Chemical reduction using sodium borohydride; various soluble metal salts (suchas ferrous or ferric chloride for iron) in suspensions of water, or various organichydrocarbon solvents; this process may or may not be enhanced with sonificationduring reaction processes

• Other chemical precipitation reactions in aqueous or hydrocarbon solutions capa-ble of producing metals from soluble salts that may or may not include sonificationduring reaction processes

• Various methods of metal volatilization and subsequent deposition, typically undervacuum

The top down approach uses two primary variations of milling or mechanicalcomminuation that include:

Figure 4.50 Nano-scale bimetallic clusters.



• Using mechanical agitation of a mixture of the desired colloidal metal, a grindingmedia, and an organic or aqueous suspension fluid; examples include ball millsand rod mills

• Systems similar to the above where the mechanical agitation is provided by highspeed gas jets

There are methods available to produce superfine particles that have distinctmorphology and internal crystal structure to further enhance the surface reactivity.106

In addition, it is important to recognize that, in the nano-scale range, quantum sizeeffects begin to become apparent. For example a colloid of 10 nm diameter hasabout 30% of its atoms in grain boundaries (which are highly reactive and subjectto quantum effects). These features have an effect on the physical/chemical behaviorof the particle in use which falls into one of two broad categories reflecting onproduction by bottom up or top down methods.100

• A colloid produced by chemical precipitation or reduction, or through the variousvapor deposition methods, will be nano-structured. This means that the colloidwill have nano-scale crystal domains with sharp boundaries between crystals. Thegrain boundaries are typically only 1 atom thick and there is low dislocationdensity in the crystal structures.100

• The reactivity of a colloid of this type can be controlled primarily through theselection of an appropriate overall colloid size and resulting surface area. Smallersize means greater surface area and reactivity; larger size means lower surfacearea and reactivity.100

• A colloid produced by mechanical attrition will be nano-crystalline. The crystaldomains in the colloid are small, relative to the overall colloid size. The individualcrystal domains are separated by wide amorphous transition regions that exhibita very high dislocation density. These transition regions may be as large as thecrystal domains, but are still termed grain boundaries.100

• The amorphous transition regions will be highly reactive. The reactivity of thecolloid will be dominated by the size and intensity of dislocation density of theamorphous boundary regions rather than the absolute size of the colloid. A rela-tively large colloid produced by this method could have reactivity the same as orgreater than a much smaller colloid produced by bottom up methods.100

Control of the reactivity of the colloid is a critical feature. The iron undergoesanaerobic corrosion, reacting directly with halogenated solvents as well as withwater to produce hydrogen. As the reactivity of the colloid increases, the hydrogenproduction rate increases as well. By controlling the rate of hydrogen productionusing the methods described above, it will be possible to design reactive metalcolloids with reactivity that will generate hydrogen at the rate required for the desireddehalogenation processes — rather than being consumed at excessively higher rates(with just water) at which the iron colloid would be consumed (by the water) withoutreacting with the halogenated solvents undergoing treatment. Controlling the typeof nano-scale particle produced is particularly important for in situ applications inorder to maximize the rate of hydrogen production needed to achieve the remediationobjectives.



4.5.2 Injection of Nano-Scale Particles in Permeable Sediments

Injection of nano-scale particles into intragranular pore space of the geologicmatric is the preferred mode of application for directly addressing microemulsionsof contaminants in source areas or for treating dissolved phase contaminants. Dif-fusion of contaminants and hydrogen generated by Fe (0) and advection of Fe (0)particles should provide the intimate contact between contaminants and Fe (0). Themobility of colloids is governed by mechanical filtration and adsorptive processeswithin the porous media; it is always preferable to achieve the largest reactive zonefrom each injection point for economic reasons.

Colloids and nano-scale particles can be mechanically removed by the soilmatrix. The key parameter to this process is the pore entrance size, which is afunction of grain size. In fine- to coarse-grained silts, pore entrance (throat) sizesrange from 0.7 to 7 µm, in fine- to coarse-grained sands from 24 to 240 µm, andin fine- to coarse-grained gravels from 720 to 7,200 µm. It is obvious from thisinformation that nano-scale particles will travel further from the point of injectionthan typical colloids, particularly within more permeable formations.

Injection of nano-scale particles with shear thinning fluids also will enhance theinjectability of Fe (0) particles into the porous media. In contrast to Newtonianfluids, whose viscosities are constant with shear rate, certain non-Newtonian fluidsare shear thinning, that is, the viscosity of these fluids decreases with increasingshear rate.108 The primary benefit of using these fluids for this application is thatthey increase the viscosity of the aqueous phase without adversely decreasing thehydraulic conductivity. A suspension formulated with a shear thinning fluid willmaintain a relatively high viscosity in solution near Fe (0) particles (where the shearstress is low) relative to locations near the surfaces of the porous media, where theshear stress is high. The increased viscosity decreases the rate of gravitational settlingof the Fe (0) particles while maintaining a relatively high hydraulic conductivitythat permits injecting the Fe (0) suspension into the porous media at greater flowrates and distances. If an easily biodegradable shear thinning fluid is selected, it willalso provide an additional benefit in the form of scavenging the dissolved oxygenpresent within the reactive zone and ensuring that the reactive iron is consumedprimarily by the degradation of the contaminant mass. Above ground engineeringcontrols to prevent agglomeration of the Fe (0) particles in injection solution willalso enable the injected particles to travel farther in the porous media. This willentail control of the ionic state of the suspension fluid to prevent agglomeration, useof surfactants, and determination of the optimum colloidal concentration for thesuspension.

4.5.3 Organic Contaminants Treatable by Fe (0)

Tables 4.14 and 4.15 present a list of organic contaminants that are treatable andnot treatable by Fe (0) based on the current state of science. Table 4.16 presents alist of compounds with unknown reactivity with Fe (0).



Table 4.14 Contaminants Treatable by Zero Valent Iron, Fe (0)101

Organic Compounds Inorganic Compounds

Methanes tetrachloromethane Dissolved Metals Chromiumtrichloromethane Nickeldichloromethane Lead

UraniumTechnetiumIronManganeseSeleniumCopperCobaltCadmiumZinc

Ethanes hexachloroethane Anion Contaminants Sulphate1,1,1-trichloroethane Nitrate1,1,2-trichloroethane Phosphate1,1-dichloroethane Arsenic

Ethenes tetrachloroethenetrichloroethenecis-1,2-dichloroethenetrans-1,1-dichloroethene1,1-dichloroethenevinyl chloride

Propanes 1,2,3-trichloropropane1,2-dichloropropane

Aromatics benzenetolueneethylbenzene

Other hexachlorobutadiene1,2-dibromoethanefreon 113N-nitrosodimethylamine

Table 4.15 Contaminants Presently Not Treatable by Fe (0)101


dichloromethane chloride1,2-dichloroethane perchloratechloroethanechloromethaneheavier PAHs

Table 4.16 Contaminants with Unknown Treatability


chlorobenzenes mercurychlorophenolscertain pesticidesPCBs



REFERENCES

1. Suthersan, S. S., Engineered In Situ Anaerobic Reactive Zones, U.S. Patent 6,143,177,1998.

2. Bouwer, E. J., Rittmann, B. E., and McCarty, P. L., Anaerobic degradation of halo-genated 1-, and 2- carbon organic compounds, Environ. Sci. Technol., 15, 596–599,1987.

3. Parsons, F., Wood, P. R., and DeMarco, J., Transformations of tetrachloroethene andtrichloroethene in microcosms and groundwater, J. Am. Water Works Assoc., 76,56–59, 1984.

4. Vogel, T. M. and McCarty, P. L., Biotransformation of tetrachloroethylene to trichlo-roethylene, dichloroethylene, vinyl chloride, and carbon dioxide under methanogenicconditions, J. Appl. Environ. Microbiol., 49, 1080–1083, 1985.

5. Kleopfer, R. D. et al., Anaerobic degradation of trichloroethylene in soil, Environ.Sci. Technol., 19, 277–280, 1985.

6. Freedman, D. L. and Gossett, J. M., Biological reductive dehalogenation of tetrachlo-roethylene and trichloroethylene to ethylene under methanogenic conditions, J. Appl.Environ. Microbiol., 55, 1009–1014, 1989.

7. DeBruin, W. P. et al., Complete biological reductive transformation of tetrachloroet-hene to ethane, J. Appl. Environ. Microbiol., 58, 1996–2000, 1992.

8. Christiansen, N. et al., Transformation of tetrachloroethene in an upflow anaerobicsludge blanket reactor, Appl. Microbiol. Biotechnol., 47, 91–94, 1997.

9. Kastner, M., Reductive dechlorination of tichloroethylenes and tetrachloroethylenesdepends on transition from aerobic to anaerobic conditions, J. Appl. Environ. Micro-biol., 57, 2039–2046, 1991.

10. DiStefano, T. D. et al., Reductive dechlorination of high concentrations of tetrachlo-roethene to ethane by an anaerobic enrichment culture in the absence of methano-genesis, J. Appl. Environ. Microbiol., 57, 2287–2292, 1991.

11. DiStefano, T. D., Gossett, J. M., and Zinder, S. H., Hydrogen as an electron donorfor dechlorination of tetrachloroethene by an anaerobic mixed culture, J. Appl. Envi-ron. Microbiol., 58, 3622–3629, 1992.

12. Alexander, M., Biodegradation and Bioremediation, Academic Press, New York,1999.

13. Schwarzenbach, R. P., Gschwenda, P. M., and Imboden, D. M., EnvironmentalOrganic Chemistry, John Wiley & Sons, New York, 1993.

14. Wiedemeier, T. H. et al., Natural Attenuation of Fuels and Chlorinated Solvents inthe Subsurface, John Wiley & Sons, New York, 1999.

15. McCarty, P. L. and Semprini, L., Groundwater treatment for chlorinated solvents, inHandbook of Bioremediation, Norris, R. D. et al., Eds., Lewis Publishers, Boca Raton,FL, 1994.

16. Hollinger, C. and Schumacher, W., Reductive dehalogenation as a respiratory process,Antoine Leeuwenhoek, 66, 239–246, 1994.

17. Gossett, J. M. and Zinder, S. H., Microbiological aspects relevant to natural attenu-ation of chlorinated ethenes, in Proc. Symp. Nat. Attenuat. Chlorinat. Org. Ground-water, Dallas, TX, September 11-13, 1996, EPA/540/R-96/509.

18. Vogel, T. M., Criddle, T., and McCarty, P. L., Transformations of halogenated aliphaticcompounds, Environ. Sci. Technol., 21, 722–736, 1987.

19. Neumann, A. et al., Purification and characterization of tetrachloroethene reductivedehalogenase from Dehalospirillum mutivorans, J. Biochem. Mol. Biol., 271,16515–16519, 1996.



20. USEPA, Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Actionand Underground Storage Tank Sites, Directive No. 9200.4-17P, Office of Solid Wasteand Emerging Response, Washington, D.C., 1999.

21. Bradley, P. M. and Chappelle, F. H., Anaerobic mineralization of vinyl chloride inFe (III) reducing aquifer sediments, Environ. Sci. Technol., 30, 2084–2086, 1996.

22. Luijten, M. L. et al., Effect of alternative electron acceptors on the reduction oftetrachloroethene, in Proc. 2nd Int. Conf. Remed. Chlorinat. Recalcitrant Comp.,Monterey, CA, May 2000.

23. Bouwer, E. J., Bioremediation of chlorinated solvents using alternate electron accep-tors, in Handbook of Bioremediation, Norris, R. D. et al., Eds., Lewis Publishers,Boca Raton, FL, 1994.

24. Gantzer, C. J. and Wackett, L. P., Reductive dechlorination catalyzed by bacterialtransition metal coenzymes, Environ. Sci.Technol., 25, 715–722, 1991.

25. Middeldorp, P. J. M. et al., Anaerobic microbial reductive dehalogenation of chlori-nated ethenes, Bioremed. J., 3, 151–169, 1999.

26. Belay, N. and Daniels, L., Production of ethane, ethylene and acetylene from halo-genated hydrocarbons by methanogenic bacteria, Appl. Environ. Microbiol., 53,1604–1610, 1987.

27. Ballapragada, B. S. et al., Effect of hydrogen on reductive dechlorination of chlori-nated ethenes, Environ. Sci. Technol., 31, 1728–1734, 1997.

28. Carter, S. R. and Jewell, W. J., Biotransformation of tetrachloroethylene by anaerobicattached-films at low temperatures, Wat. Res., 27, 607–615, 1993.

29. Berededsamuel, Y. et al., Effect of Perchloroethylene (PCE) on methane and acetateproduction by a methanogenic consortium, Appl. Biochem. Biotechnol., 57-8,915–922, 1996.

30. Fathepure, B. Z. and Tiedje, J. M., Reductive dechlorination of tetrachloroethyleneby a chlorobenzoate-enriched biofilm reactor, Environ. Sci. Technol., 28, 746–752,1994.

31. Gibson, S. A. and Sewell, G. W., Stimulation of reductive dechlorination of tetra-chloroethene in anaerobic acquifer microcosms by addition of short-chain acids oralcohols, Appl. Environ. Microbiol., 58, 1392–1393, 1992.

32. Guiot, S. R. et al., Anaerobic and aerobic/anaerobic treatment for tetrachloroethylene(PCE), in 3rd Int. In Situ On-Site Bioreclam. Symp., Hinchee, R.E., Leeson, A., andSemprini, L., Eds., Batelle Press, San Diego, 1995, 297–305.

33. DiStefano, T. D. et al., Hydrogen as an electron donor for dechlorination of tetra-chloroethene by an anaerobic mixed culture, Appl. Environ. Microbiol., 58,3622–3629, 1992.

34. Wiedermeier, T. H. et al., Protocol for supporting natural attenuation of chlorinatedsolvents with examples, in In Situ and On-Site Bioremediation 3, Batelle Press,Columbus, OH, 1997, 147.

35. Fennell, D. E. et al., Comparison of butyric acid, ethanol, lactic acid and propionicacid as hydrogen donors for the reductive dechlorination of tetrachloroethene, Envi-ron. Sci. Techol., 31, 918–926, 1997.

36. Maymo-Gatell, X. et al., Isolation of a bacterium that reductively dechlorinatestetrachloroethene to ethane, Science, 276, 1568–1571, 1997.

37. Maymo-Gatell, X. et al., Characterization of an H2-utilizing enrichment culture thatreductively dechlorinates tetrachloroethene to vinyl chloride and ethane in the absenceof methanogenesis and acetogenisis, Appl. Environ. Microbiol., 61, No. 11,3928–3933, 1995.



38. Hollinger, C. et al., A highly purified enrichment culture couples the reductive dechlo-rination of tetrachloroethene to growth, Appl. Environ. Microbiol., 59, 2991–2997,1993.

39. Wild, A. et al., Isolation of an anaerobic bacterium which reductively dechlorinatestetrachloroethene and trichloroethene, Biodegradation, 7, 507–511, 1996.

40. Scholz-Muramatsu, H. et al., Isolation and characterization of Dehalospirillum mul-tivorans general nov., sp. nov., a tetrachloroethene-utilizing, strictly anaerobic bac-terium, Arch. Microbiol, 163, 48–56, 1995.

41. Utkin, I., Woese, C. and Wiegel, J., Isolation and characterization of Desulfitobacte-rium Dehalogenans general nov, sp. nov., an anaerobic bacterium which reductivelydechlorinates chlorophenolic compounds, Int. J. Syst. Bacteriol., 44, 612–619, 1994.

42. Gerritse, J. et al., Desulfitobacterium sp. strain PCEI, an anaerobic bacterium thatcan grow by reductive dechlorination of tetrachloroethene or ortho chlorinated phe-nols, Arch. Microbiol., 165, 132–140, 1996.

43. Gerritse, J. et al., Complete degradation of tetrachloroethene in coupled anoxic andoxic chemostats, Appl. Microbiol. Biotechnol., 48, 553–562, 1997.

44. Miller, E. et al., Comparative studies on tetrachloroethene reductive dechlorinationmediated by Desulfitobacterium sp. strain PCE-S, Arch. Microbiol., 168, 613–519,1998.

45. Sharma, P. K. and McCarty, P. L., Isolation and characterization of a facultativelyaerobic bacterium that reductively dehalogenates tetrachloroethene to cis-1, 2-dichlo-roethene, Appl. Environ. Microbio. 62, 761–765, 1996.

46. Krumholz, L. R., Desulfuromonas chloroethenica sp. nov. uses tetrachloroethylene andtrichloroethylene as electron acceptors, Int. J. Syst. Bacteriol., 47, 1262–1263, 1997.

47. Maymo-Gatell, X. et al., Dehalococcus ethenogenes strain 195: ethane productionfrom halogenated aliphatics, in In Situ and On-Site Bioremediation: 3, Batelle Press,Columbus, OH, 1997, 23.

48. Schumacher, W. and Holliger, C. L., The proton electron ratio of the menaquinonedependent electron transport from dihydrogen to tetrachloroethene in Dehalobacterrestrictus, J. Bacteriol., 178, 2328–2333, 1996.

49. Miller, E. et al., Studies on tetrachloroethene respiration in Dehalospirillum multi-vorans, Arch. Microbiol., 166, 379–387, 1996.

50. Rosner, B. M. et al., In vitro studies on reductive vinyl chloride dehalogenation byan anaerobic mixed culture, Appl. Environ. Microbiol., 63, 4139–4144, 1997.

51. Harkness, M. R., Economic considerations in enhanced anaerobic biodegradation,Proc. 2nd Int. Conf. Remed. Chlorinat. Recalcitrant Compds., Wickramanayake, G.B. et al., Eds., Battelle Press, Columbus, OH, May 22–25, 2000.

52. DiStefano, T. D., PCE dechlorination with complex electron donors, Proc. 2nd Int.Conf. Remed. Chlorinat. Recalcitrant Compds., Wickramanayake, G. B. et al., Eds.,Battelle Press, Columbus, OH, May 22–25, 2000.

53. Suthersan, S. S., ARCADIS G & M, Inc., personal communication, 2000.54. Komatsu, T. et al., Biotransformation of cis-1,2-dichloroethylene to ethylene and

ethane under anaerobic conditions, Water Sci. Technol., 30, 75–84, 1994.55. Smatlak, C. R., Grossett, J. M., and Zinder, S. H., Comparative kinetics of hydrogen

untilization for reductive dechlorination of tetrachloroethene and methanogenesis inan anaerobic enrichment culture, Environ. Sci. Technol., 30, 2850–2858, 1996.

56. Bouwer, E. J. and McCarty, P. L. M., Transformations of 1 and 2-carbon halogenatedaliphatic organic compounds under methanogenic conditions, Appl. Environ.Micro-biol., 45, 1286–1294, 1983.



57. Fathepure, B. Z. and Vogel, T. M., Complete degradation of polychlorinated hydro-carbons by a two stage biofilm reactor, Appl. Environ. Microbiol. 57, 3418–3422,1991.

58. Brock, T. D. and Madigan, M. T., Biology of Microorganisms, 5th ed., Prentice Hall,Englewood Cliffs, NJ.

59. Nandi, R. and Sengupta, S., Microbial production of hydrogen: an overview, CRCRev. Microbiol., 24, 61–84, 1998.

60. Tandoi, V. et al., Environ. Sci. Technol., 28, 973–979, 1994.61. Gossett, J. M., Environ. Sci. Technol., 21, 202–208, 1987.62. Carr, C. S. and Hughes, J. B., Enrichment of high-rate PCE dechlorination and

comparative study of lactate, methanol, and hydrogen as electron donors to sustainactivity, Environ. Sci. Technol., 32, 1817–1824, 1998.

63. McCarty, P. L., An Overview of anaerobic transformation of chlorinated solvents, inSymp. Intrinsic Bioremed. Groundwater, Denver, CO, August 30 – September, 1994,135–142.

64. Vogel, T. M. et al., Abiotic and biotic transformation of 1,1,1-trichloroethane undermethanogenic conditions, Environ. Sci. Technol., 21, No. 12, 1208–1213, 1987.

65. Criddle, C. S. and McCarty, P., Electrolytic model system for reductive dehalogena-tion in agueous environments, Environ. Sci. Technol., 25, 973–978, 1991.

66. Hughes, J. B. and Parken, G. F., Individual biotransformation rates in chlorinatedaliphatic mixtures, J. Environ. Eng., ASCE, 99–106, 1996.

67. Alvarez-Cohen, L. and McCarty, P. L., Product toxicity and cometabolic competitiveinhibition of chloroform and trichloroethylene transformation by methanogenic rest-ing cells, Appl. Environ. Microbiol., 57, 1031–1037, 1991.

68. Alvarez-Cohen, L. and McCarty, P. L., Effects of toxicity, aeration, and reductantsupply on trichloroethylene transformation by a mixed methanogenic culture, Appl.Environ. Mecorbiol., 57, 228–235, 1991.

69. Rasche, E. M., Hyman, M. H., and Arp, D. J., Factors limiting aliphatic chlorocarbondegradation by Nitrosomonas europaea: cometabolic inactivation of ammoniamonooxygenase and substrate specificity, Appl. Environ. Microbiol., 57, 2986–2999,1991.

70. Folsom, B. R. and Chapman, P. J., Performance characterization of a model bioreactorfor the biodegradation of trichloroethylene by Pseudomonas cepacia G4, Appl. Envi-ron. Microbiol., 57, 1602–1608, 1991.

71. Bradley, P. M. and Chapelle, F. H., Anaerobic minerilization of vinyl chloride in Fe(III) reducing acquifer sediments, Environ. Sci. Technol., 30, 2084-2086, 1996.

72. Bradley, P. M. and Chappelle, F. H., Kinetics of DCE and VC mineralization undermethanogenic and Fe (III) reducing conditions, Environ. Sci. Technol., 31, 2692–2696,1992.

73. Bradley, P. M. and Chapelli, F. H., Effect of contaminant concentration on aerobicmicrobial mineralization of DCE and VC in stream-bed sediments, Environ. Sci.Technol., 30, No. 5, 553–557, 1978.

74. Bushman, A. et al., Environ. Sci. Technol., 33, 1015–1020, 1999.75. Adrianes, P. et al., Biogeochemistry and dechlorination potential at the St. Joseph

acquifer Lake Michigan interface, in Ed., In Situ and On-Site Bioremediation 3,Batelle Press, Columbus, OH, 1997, 173–178.

76. Payne, F., ARCADIS G & M, Inc., Personal Communication, 2000.77. Pyrih, R. Z., Recognizing the natural attenuation of metals, in Proc. IBS’s 4th Ann.

Conf. Nat. Attenuation, Pasadena, CA, December, 1998.78. Potter, S., ARCADIS G & M, Inc., personal communication, 2000.



79. Sparks, D. L., Ed., Soil Physical Chemistry, 2nd Ed., CRC/Lewis Publishers, BocaRaton, FL, 1998.

80. Hong, J., Forstner, U., and Calmano, W., Effects of REDOX processes on acidproducing potential and metal mobility in sediments, in Bioavailability: Physical,Chemical and Biological Interactions, Hamlink, J. L. et al., Eds., Lewis Publishers,Boca Raton, FL, 1994.

81. Ehrlich, H. L., Geomicrobiology, Marcel Dekker, New York, 1981.82. Melhorn, R. J., Buchanan, B. B., and Leighton, T., Bacterial chromate reduction and

product characterization, in Emerging Technology for Bioremediation of Metals,Means, J. L. and Hinchee, R. E., Eds., Lewis Publishers, Boca Raton, FL, 1994.

83. Turick, C. E., Graves, C., and Apel, W. A., Bioremediation potential of Cr (VI)-contaminated soil using indigenous microorganisms, Bioremed. J., 2, 1–6, 1998.

84. Bader, J. L. et al., Aerobic reduction of hexavalent chromium in soil by indigenousmicroorganisms, Bioremed. J., 3, 201–212, 1999.

85. Palmer, C. and Puls, R., Natural Attenuation of Hexavalent Chromium in Groundwaterand Soils, USEPA, Office of Research and Development, OSWER, EPA/540/S-94/505, 1994.

86. Gary, L. and Rai, D., Kinetics of chromium (III) oxidation to cromium (VI) byreaction with manganese dioxides, Environ. Sci. Technol., 21, 1187–1193, 1987.

87. Urbansky, E. T., Perchlorate chemistry: implications for analysis and remediation,Bioremed. J., 2, 89–97, 1998.

88. Logan, B. E., A review of chlorate- and perchlorate-respiring microorganisms,Bioremed. J., 2, 69–80, 1998.

89. Giblin, T., et al., Removal of perchlorate in groundwater with a flow-through biore-actor, J. Environ. Qual., 29, 578–583, 2000.

90. Hartsman, S., and de Bont, J. A. M., Aerobic vinyl chloride metabolism in Mycobac-terium aunum li, Appl. Environ. Microbiol., 58, 1220–1226, 1985.

91. McCarty, P. L. and Semprini, L., Groundwater treatment for chlorinated solvents, inHandbook of Bioremediation, Norris, R. D., et al., Eds., Lewis Publishers, BocaRaton, FL, 1994.

92. Stefan, R., Environ. Corporation., personal communication, 2000.93. Yin, Y. and Allen, H. E., In Situ Chemical Treatment, Groundwater Remediation

Technologies Analysis Center, Pittsburgh, PA 1999.94. Yan, Y. E. and Schwartz, F. W., Oxidative degradation and kinetics of chlorinated

ethylenes by potassium permanganate, J. Contaminant Hydrol., 37, 343–365, 1999.95. Haag, W. R. and Yao, C. C. D., Rate constants for reaction of hydroxyl radicals with

several drinking water contaminants, Environ. Sci. Technol., 26, 1005–1013, 1992.96. Walling, J., Fenton’s reagent revisited, Acc. Chem. Res., 8, 125–131, 1975.97. Siegrist, R. L., In Situ chemical oxidation: technology features and applications, in

Proc. Conf. Adv. Innovat. Groundwater Remed. Technol., Atlanta, December, 1998.98. Hoigne, J. and Bader, H., Rate constants of reactions of ozone with organic and

inorganic compounds in water − Nondissociating organic compounds, Water Res.,17, 173–183, 1983.

99. Basel, M.D. and Nelson, C.H., Overview of in situ chemical oxidation: status andlessons learned, paper presented at the 2nd Int. Conf. Remed. Chlorinat. RecalcitrantCompds, Monterey, CA, May, 2000.

100. Vance, D., ARCADIS G & M, Inc., Personal Communication, 2000.101. USEPA, Permeable Reactive Barrier Technologies for Contaminant Remediation,

EPA/600/R-98/125, Washington, DC, September, 1998.



102. Tratnyek, P.G. et al., Remediating groundwater with zero valent metals: chemicalconsiderations in barrier design,

Groundwater Monitor. Rev.

, Fall, 1997.103. Johnson, T.L. et al., Kinetics of halogenated organic compound degradation by iron

metal,

Environ. Sci. Technol

., 30, 2634–2640, 1996.104. Gillham, R.W. and O’Hannesin, S.F., Enhanced degradation halogenated aliphatics

by zero valent iron,

Groundwater

, 32, 958–967, 1994.105. Wang, C.B. and Zhang, W.X., Synthesizing nano-scale iron particles for rapid and

complete dechlorination of TCE and PCBs,

Environ. Sci. Technol

., 31, 2154–2156,1997.

106. Vance, D., Suthersan, S., and Palmer, P., Method of Making and Using Nano-ScaleMetal, U.S. Patent (pending).

107. Ichinose, N., Ozaki, Y., and Kashu, S.,

Superfine Particle Technology

, Springer-Verlag,Tokyo, 1992.

108. Cantrell, K.J., Caplan, D.I., and Gilmore, T.J., Injection of colloidal size particles ofFe (0) in porous media with shear thinning fluids as a method to emplace a permeablereactive zone,

Int. Contain. Technol. Conf.

, St. Petersburg, FL, February, 1997.

CREDIT

Figure 4.9 is adapted from Van Briesen, J.M. and Rittmann, B.E.,

Natural Atten-tuation Consideration and Case Studies: Remediation of Chlorinated and Recalci-trant Compounds

, Wickramanayake, G.B., Gavaskar, A.R., and Kelley, M.E. (Eds),Battelle Press, Columbus, OH. With permission.

L1282_C04-C_frame Page 238 Thursday, March 6, 2003 12:08 PM

239

CHAPTER

5

Phytoremediation

CONTENTS

5.1 Introduction ..................................................................................................2405.2 Chemicals in the Soil–Plant System............................................................241

5.2.1

Metals..................................................................................................2415.2.2

Organics ..............................................................................................2425.3 Types of Phytoremediation ..........................................................................244

5.3.1 Phytoaccumulation ...........................................................................2455.3.2 Phytodegradation..............................................................................2485.3.3 Phytostabilization .............................................................................2505.3.4 Phytovolatilization............................................................................2515.3.5 Rhizodegradation..............................................................................2525.3.6 Rhizofiltration...................................................................................2565.3.7 Phytoremediation for Groundwater Containment ...........................2595.3.8 Phytoremediation of Dredged Sediments ........................................260

5.4 Phytoremediation Design.............................................................................2615.4.1 Contaminant Levels .........................................................................2655.4.2 Plant Selection..................................................................................2655.4.3 Treatability .......................................................................................2665.4.4 Irrigation, Agronomic Inputs, and Maintenance .............................2665.4.5 Groundwater Capture Zone and Transpiration Rate .......................267

References..............................................................................................................267

… many accepted agricultural techniques for cultivating, harvesting, and pro-cessing plants have now been adapted for phytoremediation. Overall, the appli-cation of phytoremediation is being driven by its technical and economic advan-tages over conventional approaches … .phytoremediation’s future is not ascientific issue, but rather a “scientific sociology” issue….



5.1

INTRODUCTION

Phytoremediation is defined as “the engineered use of plants

in situ

and

ex situ

for environmental remediation.” The technology involves removing or degradingorganic and inorganic contaminants and metals from soil and water. The processesinclude all plant-influenced biological, chemical, and physical processes that aid inthe uptake, sequestration, degradation, and metabolism of contaminants, either byplants or by the free living organisms that constitute a plant’s rhizosphere. Phytore-mediation takes advantage of the unique and selective uptake capabilities of plantroot systems, together with the translocation, bioaccumulation, and contaminantstorage and degradation capabilities of the entire plant body.

The concept of using plants to alter the environment has been around since plantswere first used to drain swamps. What is new within the context of this newtechnology called phytoremediation is the systematic, scientific investigation of howplants can be used to decontaminate soil and water.

1

Interest in phytoremediationhas been growing in the U.S. during the past few years with potential applicaton ofthis technology at a wide range of sites contaminated with heavy metals, pesticides,explosives, and solvents.

The potential benefits of phytoremediation seem to be as numerous as theproblems it might address. One reason this technology is gaining attention is becauseit is potentially cheaper than conventional treatment approaches for contaminatedsoils and traditional pump and treat systems for contaminated groundwater, such asincineration or soil washing. Another attraction of this technology is that it mayleave topsoil in usable condition, keeping soil fertility and structure intact whilereducing contamination levels at the same time. Phytoremediation is well suited forapplications in low permeability soils, where most currently used technologies havea low degree of feasibility or success, as well as in combination with more conven-tional remediation technologies.

The main advantages of phytoremediation are the low capital costs, aestheticallypleasing technique, minimization of leaching of contaminants, and soil stabilization.The operational cost of phytoremediation is also substantially less than that of conven-tional treatments and involves mainly fertilization and watering for maintenance of plantgrowth. In the case of heavy metals remediation, additional operational costs includeharvesting, disposal of contaminated plant mass, and repeating the plant growth cycle.

It should be emphasized that there is more to phytoremediation than merelyputting plants in the ground and letting them do the work. Phytoremediation alsohas its drawbacks, which even its ardent champions are quick to acknowledge. Firstof all, it is a time-consuming process that can take several growing seasons to cleana site. Vegetation that absorbs toxic heavy metals will have to be harvested andmanaged as a waste. This vegetation containing high concentrations of toxic metalsand organics may also pose a risk to wildlife. The shutdown of plant activity duringwinter months and the seasonal variation of plant metabolic activity is a drawbackfor application of this technology in colder climates. Other limitations of phytore-mediation are that contaminants present below rooting depth will not be treated orextracted and that the plant or tree may not be able to grow in soils at heavilycontaminated sites due to plant toxicity.


PHYTOREMEDIATION 241

Phytoremediation as a technology is still in its early stages. While many scien-tists, engineers, and regulators are optimistic that it will eventually be used to cleanup organic and metallic contaminants, at least two or three more years of field testsand analyses are necessary to validate the initial, small-scale field tests.

1,2

Issues likesoil characteristics and length of the growing season will need to be taken intoaccount and scientists must also determine what sites are most amenable to phy-toremediation. Other issues such as the potential impact on wildlife remain to befully explored. Simultaneously, researchers working in the lab are trying to betterunderstand the processes behind phytoremediation to possibly improve its perfor-mance during cleanup applications.

This chapter will not do justice to this technology by claiming that it will coverthe rapidly progressing state of the science and also describe how these scientificadvances are being applied in the field for efficient remediation. Instead it will serveas a brief state of the science summary that will allow the reader to understand thecurrent status of the technology and its applications, as well as activities of theresearch community to further enhance this technology.

5.2 CHEMICALS IN THE SOIL–PLANT SYSTEM

5.2.1 Metals

Elements occur in the soil in a variety of forms more or less available for uptakeby plants. Many of the contaminants of concern at waste sites are metals or metal-loids. Availability is determined by characteristics of the elements, such as behaviorof the ion as a Lewis acid (electron acceptor) which determines the predominanttype of strength of bond created (ionic or covalent) and, therefore, the mobility ofthe metal in the soil environment. Soil characteristics (e.g., pH, clay and organicmatter content and type, and moisture content) also determine availability to plantsby controlling speciation of the element, temporary immobilization by particlesurfaces (adsorption-desorption processes), precipitation reactions, and availabilityin soil solution. The most general sinks for metals are iron and manganese oxidesand organic matter. Although particulate soil organic matter serves to immobilizemetals, soluble organic matter may act to keep metals in solution in a form absorbedand translocated by plants.

Metal fractionation or sequential extraction schemes — such as toxicity charac-teristic leaching procedure (TCLP) — sometimes are used to describe metal behaviorin soils. Most metals interact with the inorganic and organic matter that is presentin the root-soil environment. Potential forms of metals include those dissolved inthe soil solution, adsorbed to the vegetation’s root system, adsorbed to insolubleorganic matter, bonded to ion exchange sites on inorganic soil constituents, precip-itated or coprecipitated as solids, and attached to or inside the soil biomass.

The final control on availability of metals and metalloids in soil to plants is theselective absorption from soil solution by the root. Metals may be bound to exteriorexchange sites on the root and not actually taken up. They may enter the rootpassively in organic or inorganic complexes with the mass flow of water or actively



by way of metabolically controlled membrane transport systems often meant to takeup a nutrient which the “contaminant” metal mimics. At different soil solute con-centrations, metals may be absorbed by both processes. Absorption mechanisms andquantity absorbed are influenced by plant species (and cultivar), growth stage,physiological state, and the presence of other elements.

Once in the plant, a metal can be sequestered in the roots in vacuoles or inassociation with cell walls and organelles, or translocated to above ground parts inxylem as organic or inorganic complexes. Location and forms of metals in plants,as well as their toxic effects, depend on plant species, growth stage, physiologicalstate, and presence of other metals.

Mechanisms of toxicity of metals tend to be dependent on the nature of thereactivity of the metal itself and its availability in the soil and soil solution media.They may alter or inhibit enzyme activity, interfere with deoxyribonucleic acid(DNA) synthesis or electron transport, or block uptake of essential elements.

2

Avail-ability in response to toxic levels of metals by different plants is due to a numberof defenses. These include exclusion from the root, translocation in nontoxic form,sequestering in nontoxic form, sequestering in nontoxic form in the root or otherplant parts, and formation of unusable complexes containing metals that may oth-erwise be inserted into biomolecules instead of the proper element (e.g., Asreplacing P).

5.2.2 Organics

Organic compounds of environmental concern include nonionic compounds(such as PAHs, chlorinated benzenes, polychlorinated biphenyls (PCBs), BTEXcompounds, and many pesticides), ionizable compounds (chlorophenols, carboxylicacids, surfactants, and amines), and weakly hydrophobic volatile organic compounds(trichloroethene). For the nonionic compounds, sorption in soil is mainly a functionof degree of hydrophobicity and amount of sorbent hydrophobic phase (i.e., soilorganic matter). Sorption of the compound by soil organic matter is reversible. Theactivities of these compounds in soil can be predicted by the organic matter-watercoefficient, K

om

, as estimated by the octanol-water coefficient, K

ow

.

3

Absorption ontocolloidal organic matter in solution may alter the availability of these nonioniccompounds. Ionizable compounds contain anionic or cationic moieties or both withintheir structure. These charged structures interact with organic and inorganic chargedsurfaces in the soil in a variety of reversible reactions. The extent and nature of theassociations with charged surfaces depends on characteristics of the organic com-pound, solution pH and ionic strength, and mineral composition of the soil partic-ulates. Organic compounds may be degraded by microorganisms in the soil tometabolites with greater or lesser toxicity. Very stable compounds, like highly chlo-rinated PCBs, may persist in essentially unaltered form for many years.

Plant roots are not discriminating in uptake of small organic molecules (molec-ular weight less than 500) except on the basis of polarity.

1-4

More water-solublemolecules pass through the root epidermis and translocate throughout the plant. Theless soluble compounds (like many polycyclic aromatic hydrocarbons) seem to havelimited entry into the plant and minimal translocation once inside. Highly lipophilic



compounds, such as PCBs, move into the plant root via the symplastic route (fromcell to cell, as opposed to between cells) and are translocated within the plant. Withina plant the contaminant may be adsorbed on a cell surface or accumulated in thecell. Many contaminants become bound on the root surface and are not translocated.

Not all organic compounds are equally accessible to plant roots in the soilenvironment. The inherent ability of the roots to take up organic compounds can bedescribed by the hydrophobicity (or lipophilicity) of the target compounds. Thisparameter is often expressed as the log of the octanol-water partioning coefficient,K

ow

. Direct uptake of organics by plants is a surprisingly efficient removal mechanismfor moderately hydrophobic organic compounds. There are some differencesbetween the roots of different plants and under different soil conditions, but, gen-erally, the higher a compound’s log K

ow

, the greater the root uptake.Hydrophobicity also implies an equal propensity to partition into soil organic

matter and onto soil surfaces. Root absorption may become difficult with heavilytextured soils and soils with high native organic matter. There are several reportedvalues available in the literature regarding the optimum log K

ow

value for a compoundto be a good candidate for phytoremediation (as an example, log K

ow

= 0.5–3.0; logK

ow

= 1.5–4.0).

2,13

It has also been reported that compounds that are quite watersoluble (log K

ow

< 0.5) are not sufficiently sorbed to the roots or actively transportedthrough plant membranes.

From an engineering point of view, a tree could be thought of as a shell of livingtissue encasing an elaborate and massive chromatography column of twigs, branches,trunk, and roots. The analogous resin in this system is wood, the vascular tissue ofthe tree, and this “resin” is replenished each year by normal growth. Wood iscomposed of thousands of hollow tubes, like the bed of a hollow fiber chromatog-raphy column, with transpirational water serving as the moving phase. The hollowtubes are actually dead cells, whose death is carefully programmed by the tree toproduce a water conducting tissue, which also functions in mechanical support. Acomplex, cross-linked, polymeric matrix of cellulose, pectins, and proteins embed-ded in lignin forms the walls of the tubes. The cell wall matrix is chemically inert,insoluble in the majority of solvents, and stable across a wide range of pH.

Once an organic chemical is taken up, a plant can store (sequestration) thechemical and its fragments in new plant structures via lignification, or it can vola-tilize, metabolize, or mineralize the chemical all the way to carbon dioxide, water,and chlorides. Detoxification mechanisms may transform the parent chemical tononphytotoxic metabolites, including lignin, that are stored in various places in plantcells. Many of these metabolic capacities tend to be enzymatically and chemicallysimilar to those processes that occur in mammalian livers; one report has equatedplants to” green livers” due to similarities of detoxification processes.

Different plants exhibit different metabolic capacities. This is evident during theapplication of herbicides to weeds and crops alike. The vast majority of herbicidalcompounds have been selected so that the crop species are capable of metabolizingthe pesticide to nontoxic compounds, whereas the weed species either lack thiscapacity or perform it at too slow a rate. The result is the death of the weed specieswithout the metabolic capacity to rid itself of the toxin.



The shear volume and porous structure of a tree’s wood provide an enormoussurface area for exchange or biochemical reactions. Some researchers are attemptingto augment the inherent metabolic capacity of plants by incorporating bacterial,fungal, insect, and even mammalian genes into the plant genome.

5.3 TYPES OF PHYTOREMEDIATION

A review of where pyhtoremediation fits into the scheme of hazardous wasteremediation enables us to differentiate the various types and mechanisms of phy-toremediation (Figure 5.1). The scientific understanding of plant, soil and rhizo-sphere biochemistry, and contaminant fate and transport must be contrasted withfield and pilot studies that represent the current proof of concepts. The technologyis summarized below as those approaches ready for application, promising treatmentsexpected to be tested soon, and concepts of phytoremediation requiring intensivedevelopment. Finally, the intrinsic strengths of phytoremediation as a technologyand the future potential of this technology must be reviewed for regulatory accep-tance in terms of hazardous waste remediation.

1,2

Phytoremediation approaches can be summarized as follows based on currentunderstanding of the technology:

• Phytoaccumulation, phytoextraction, hyperaccumulation• Phytodegradation or phytotransformation• Phytostabilization• Phytovolatilization• Rhizodegradation, phytostimulation, or plant assisted bioremediation• Rhizofiltration or contaminant uptake

Optimal performance of the technology is an important key to phytoremedia-tion’s ability to gain wider acceptance as a presumptive remediation technique. With

Figure 5.1

Potential contaminant fates during phytoremediation in the soil–plant–atmospherecontinuum.

Mechanismsfor Organics

Mechanismsfor Inorganics

AtmosphereContaminant

in the air

PlantContaminantin the plant

SoilContaminant

in the root-zone(Rhizosphere)

Phytovolatilization

Phytodegradation

Rhizodegradation

Rhizofiltration

Phytostabilization

Impacted Media Impacted Media

Phytostabilization

Rhyzofiltration

Phytoaccumulation

Phytovolatilization

Rem

edia

ted

Con

tam

inan

t



the possible exception of some of the above mechanisms that are already widelystudied and understood, all of phytoremediation’s major applications require furtherbasic and applied research in order to optimize field performance. Significantresearch and development should be carried out to 1) obtain a better understandingof mechanisms of uptake, transport, and accumulation of contaminants; 2) improvecollection and genetic evaluation of hyperaccumulating plants; and 3) obtain a betterunderstanding of interactions in the rhizosphere interactions among plant roots,microbes, and other biota.

Short of true regulatory reform, phytoremediation’s ability to make furtherinroads will depend on how quickly federal, state, and local regulators becomeconvinced of the technology’s efficacy. While not involved in every decision makingprocess, the public is sometimes a key constituency as well. One can expect publicinterest groups to be more concerned about efficacy and safety issues than cost orother economic factors. However, phytoremediation seems to be faring well withthe general public and, according to many practitioners, has already proven popularwith neighbors and other interested parties at field remediation sites.

5.3.1 Phytoaccumulation

Remediation of contaminated soils using nonfood crops, called phytoaccumula-tion, has attracted a great deal of interest in recent years. Also called phytoextraction,phytoaccumulation, refers to the uptake and translocation of metal contaminants inthe soil by plant roots into the above ground portions of plants.

2

Certain plants,called hyperaccumulators, absorb unusually large amounts of metals in comparisonto other plants and the ambient metals concentration (Table 5.1).

Phytoaccumulators or phytoextractors must have a high accumulation factor, thatis, a high uptake of metals from the soil. The uptake should be metal specific, whichdiminishes the risk of impoverishing the soil of nutrient elements. The property ofhaving a high specific uptake must be genetically stable. Since the removal of metalsfrom the soil is actually achieved through the harvest, it is necessary that the planthave a high transport of the metal(s) from the roots to the shoots to be effectiveduring remediation applications. In addition, a high biomass production of the

Table 5.1 The Number of Taxonomic Groups of Hyperaccumulators Varies According to Which Metal

is Hyperaccumulated

2

MetalNumber of Taxonomic Groups

of Hyper Accumulators

Ni >300Co 26Cu 24Zn 18Mn 8Pb 5Cd 1



phytoaccumulator is needed for high removal of metals per unit area. It is also anadvantage if biomass production is of economic interest.

Hyperaccumulators have been preferred during phytoaccumulation applicationsbecause they take up very large amounts of a specific metal. They are often endemicand of a specific population (genotypes/clones) of a species.

5

However, these plantsseldom have high biomass production and may also have low competitive ability inless polluted areas, probably because the plant uses its energy to tolerate such highlevels of metals in the tissue instead of growth. Hyperaccumulators can accumulate

≥

0.01% of Cd,

≥

0.1% of Cu, or

≥

1.0% Zn in leaf dry mass and may have the metalevenly distributed throughout the plant.

6

There are also high accumulators that accumulate somewhat lower metal con-centrations than hyperaccumulators but much more than “normal” plants. Theyusually have high biomass production. In these plants, there is no uniform distribu-tion of metal throughout the plant, and thus the plant might have high accumulationeither in the roots or in the shoots. These plants are selected and planted at a sitebased on the type of metals present and other site conditions. After they have beenallowed to grow for several weeks or months, they are harvested.

Landfilling, incineration, and composting are options to dispose of or recyclethe metals, although this depends upon the results of TCLP and cost. Planting andharvesting of plants may be repeated as necessary to bring soil contaminant levelsdown to allowable limits. A plan may be required to deal with the plant biomasswaste. Testing of plant tissue, leaves, roots, etc., will determine if the plant tissueis a hazardous waste. Regulators will play a role in determining the testing methodand requirements for the ultimate disposal of the plant waste.

The state of science in phytoaccumulation is as follows:

7

• Botanical prospecting dating to the 1950s in the former USSR and U.S. is availableto practitioners.

• Over 400 species of hyperaccumulators worldwide have been cataloged.• Field test kits for metal hyperaccumulation have been developed.• Uptake and segregation processes using cation pumps, ion transporters, Ca blocks,

metal chelating exudates and transporters, phytochelatin peptides, and metallothio-neins have been evaluated and continuous research is being performed to developfurther understanding.

The hyperaccumulator plants can contain toxic element levels in the leaf andstalk biomass (LSB) about 100 times more than nonaccumulator plants growing inthe same soil, with some species and metal combinations exceeding conventionalplant levels by a factor of more than 1000.

8

Many hyperaccumulator plants, which are nonwoody (not a tree), have beenidentified as having the capacity to accumulate metals.

Thlaspi caerulascens

wasfound to accumulate Zn up to 2000–4000 mg/kg.

9

The Indian mustard plant

Brassicajuncea

, grown throughout the world for its oil seed, was found to accumulatesignificant amounts of lead.

10

One planting of mustard in a hectare of contaminatedland was found to soak up two metric tons of lead. If three plantings could besqueezed in per year, six tons of lead per hectare can be extracted. Both hempdogbane (

Apocynum

sp.) and common ragweed also have been observed to



accumulate significant levels of lead.

Aeollanthus subcaulis

var

lineris

and

Papsalumnotatus

are other hyperaccumulator plants known to accumulate Cu and Cs, respec-tively. Hyperaccumulator plants can address contamination in shallow soils only, upto 24 inches in depth. If contamination is deeper, 6–10 feet, deep-rooted poplar treescan be used for phytoextraction of heavy metals. These trees can accumulate theheavy metals by sequestration. However, there are concerns specifically for treesthat include leaf litter and associated toxic residues being blown off site. This concernmay be tested in the laboratory to see whether uptake and translocation of the metalsinto the leaves exceed standards.

Hyperaccumulators have metal accumulating characteristics that are desirable,but lack the biomass production, adaptation to current agronomic techniques, andphysiological adaptations to climatic conditions required at many contaminated sites.It has been reported that harvesting at different seasons in a year had pronounceddifferences in accumulation levels. In the future, genetic manipulation techniquesmay provide better hyperaccumulator species. The success of phytoextractiondepends on the use of an integrated approach to soil and plant management: thedisciplines of soil chemistry, soil fertility, agronomy, plant physiology, and plantgenetic engineering are currently being used to increase the rate and efficiency ofheavy metal phytoextraction.

Chelates have been used not only to enhance metal uptake but also to avoidmetal toxicity. Metal accumulator plants have been studied extensively for organo-metallic complexes. It has been suggested that there is a relationship between metaltolerance and carboxylic acids. Organo-metallic complexes increase the translocationand tolerance of plants to the toxic effects of metals. For example, in

Sebertiaacuminata

citrate seems to be a detoxifying agent as well as an agent in transportingphytotoxic Ni from root systems to the leaves until leaf fall.

5,6

It has also beensuggested that in copper (Cu) and cobalt (Co) accumulator plants, Co existed as anoxalate complex within the leaf. The formation of Zn–citrate complexes in Zn-tolerant plants was the reason for high levels of organic acid accumulation. Reportshave indicated that histidine was responsible for accumulation, tolerance, and trans-port to shoots in nonaccumulating and hyperaccumulating (Ni) plant species.

11

In

Thlaspi

, a Zn hyperaccumulator plant species, it has been determined that themajority of Zn in the roots was coordinated with histidine, whereas organic acidswere involved in xylem transport and Zn storage in the shoots. Similarly in a Cr-accumulating plant,

Leptospermum scoparium

, it was found that soluble Cr in leaftissue was present as the trioxalatochromium (III) ion, [Cr (C

2

O

4

)

3

]

3–

. The functionof the Cr-organic acid complex was to reduce the cytoplasmic toxicity of Cr.

5

Adding ethylenediaminetetraacetic (EDTA) acid, citric acid, or oxalic acid tometal contaminated soils will significantly increase the metal concentrations in plantshoots and roots.

5

However, the application of these chelates during a full scaleremediation application has to be carefully controlled; if not, the increased solubilityof the metal chelates formed could drive these contaminants to migrate furtherdownward by leaching when plant uptake rates are not adequate. Controlling thepH and conditioning the soils for optimum pH is an important factor when dealingwith metals-contaminated soils.



The schematic of the process involved in heavy metal phytoextraction is shownin Figure 5.2. Translocation from the root to the shoot must occur efficiently forease of harvesting. After harvesting, a proper, regulartorily acceptable biomassprocessing step or disposal methods should be implemented.

5.3.2 Phytodegradation

Phytodegradation, also called phytotransformation, is the breakdown of contam-inants taken up by plants through metabolic processes within the plant, or thebreakdown of contaminants external to the plant through the effect of compounds(such as enzymes) produced by the plants. Pollutants are degraded, used as nutrients,and incorporated into the plant tissues. In some cases metabolic intermediate or endproducts are rereleased to the environment depending on the contaminant and plantspecies (phytovolatilization) (Figure 5.3).

Plants synthesize a large number of enzymes as a result of primary and secondarymetabolism and can quickly uptake and metabolize organic contaminants to lesstoxic compounds. Plant enzyme systems can be constitutive or induced and can playa role in solar driven transformations and plant adaptation and/or tolerance to adverse

Figure 5.2

Process schematic describing the various processes during phytoaccumulation ofheavy metals.



growth conditions resulting from contamination of the soils. Plant-formed enzymesthat are useful for phytodegradation are nitroreductases (for munitions and pesti-cides); dehalogenases (for chlorinated solvents and pesticides); phosphatases (forpesticides); peroxidases (for phenols); laccases (for aromatic amines); cytochromeP-450 (for pesticides and chlorinated solvents); nitrilase (for herbicides).

Plant transformation pathways can be of many different types and obviouslydepend on plant species and tissue type. In simplistic terms, these pathways can becategorized as reduction, oxidation, conjugation, and sequestration. The “green livermodel” has been proposed to describe the metabolic pathways of herbicides, pesti-cides, explosives, and other nitroaromatic compounds. Contaminant degradation byplant-formed enzymes can occur in an environment free of microorganisms (forexample, an environment in which the microorganisms have been killed by highcontaminant levels). Thus, phytodegradation potentially could occur in soils wherebiodegradation cannot.

The current state of science in phytodegradation (phytotransformation) is sum-marized below:

1,2

• Plant-formed enzymes that degrade organic contaminants have been isolated andmetabolic pathways can be predicted.

• Phytodegradation can be used for the treatment of soil, sediments, sludges, andgroundwater depending on contaminant type and concentrations.

Figure 5.3

Phytodegradation and phytovolatilization mechanisms associated with some othermechanisms essential for plant life.

O2

2CO

Photosynthesis

2+O

PhloemPhotosynthates Xylem

H O, Nutrients2

2H O Transpiration andVolatilization of VOCs

Dark Respiration

O2

2CO , H O

Lignification,MetabolitesSequestration

CO , H O2

2O

Root Respiration

Exudation

O , CH COOH, C H OHCometabolism

2 3 4 5 Contaminant CO , H O, ClMineralization

2 2

ContaminentUptake

2H O, Nutients, OTranspiration

2

Phytodegradation- Metabolism within the plant- Production of enzymes which help to catalyze degradation



• Mass balance and pathway analyses studies have been conducted to prove com-plete degradation; potential toxicity of intermediate compounds also can be pre-dicted.

• Differentiation between degradation by plant enzymes, rhizosphere microorgan-isms, and other breakdown processes is being performed.

• Development of engineered solutions based on the use of monocultures vs. mul-ticultures found in wetlands and terrestrial communities is being further investi-gated.

• Organic contaminants are the main category of contaminants with the highestpotential of phytodegradation. Inorganic nutrients are also consumed through plantuptake and metabolism. Phytodegradation outside the plant does not depend onlog K

ow

and plant uptake.• Axenic plant tissue cultures of the aquatic plant

Myriophyllum

and the periwinkle

Catharanthus

are being used for elucidating plant transformation pathways.

The aquatic plant parrot feather (

Myrioplillum aquaticum

) and the algae

Nitella

have been used for the degradation of TNT. The nitroreductase enzyme has alsobeen identified in other algae, ferns, monocots, dicots, and trees.

Degradation of TCE has been detected in hybrid poplars and in poplar cellcultures, resulting in production of metabolites and in complete mineralization of asmall portion of the applied TCE.

12,14

Poplars have been used to remove atrazineand inorganic nutrients.

2

Black willow (

Salix nigra

), yellow poplar (

Liriodendrontulipifera

), bald cypress (

Taxodium diskchum

), river birch (

Betula nigra

), cherry barkoak (

Quercus falcata

), and live oak (

Quercus viginiana

) have been known to supportdegradation of herbicides.

13

One recent study demonstrated that poplar trees, whichpossess cytochrome P-450s analogous to the oxygenases responsible for transfor-mation of compounds such as TCE in the mammalian liver, exposed to 100 mg/Lof TCE did uptake and chemically alter this contaminant. TCE and its metaboliteswere found in the roots and tissue of the study trees, but not in control trees or inthe soil used for potting the trees. In a subsequent study, poplar seedlings exposedto

14

C-labeled TCE were found to generate

14

C-labeled carbon dioxide. Intermediatecompounds generated during oxidation are thought to be 2,2,2-trichloroethanol, anddi- and trichloroacetic acid. Similar studies have shown positive results for tolueneand benzene.

A recent study using parrot feather showed positive results for phytotransforma-tion of perchlorate at concentrations of up to 20 ppm.

22

Based on the results of theseexperiments and ecological knowledge of parrot feather, this species is an excellentcandidate for future research on

in situ

phytoremediation of contaminated waterbodies. Parrot feather also is a good candidate for phytoremediation of contaminatedgroundwater temporarily held in artificial ponds.

5.3.3 Phytostabilization

Phytostabilization is the use of certain plant species to immobilize contaminantsin the soil and groundwater through absorption and accumulation by roots, adsorp-tion onto roots, or precipitation within the root zone and physical stabilization ofsoils. It is also used as a means to stabilize contaminated soil by decreasing wind



and water erosion and to decrease water infiltration and the subsequent leaching ofcontaminants. This process reduces the mobility of the contaminant and preventsmigration to the groundwater or air. This technique can be used to re-establish avegetative cover at sites where natural vegetation is lacking due to high metalconcentrations. Metal-tolerant species may be used to restore vegetation to suchsites, thereby decreasing the potential migration of contamination through winderosion, transport of exposed surface soils, and leaching of soil contamination togroundwater.

Implementation of phytostabilization involves reduction in the mobility of heavymetals and high molecular weight organics by minimizing soil erodibility, decreasingthe potential for wind blown dust, and reduction in contaminant solubility by theaddition of soil amendments. Containment using plants either binds the contaminantsto the soil, renders them nonavailable, or essentially immobilizes them by removingthe means of transport.

Erosion leads to the concentration of heavy metals because of selective sortingand deposition of different size fractions of the soil. Eroded material is often trans-ported over long distances, thus selectively extending the effects of contaminationand increasing the risk to the environment. Erosion can, therefore, cause the buildup of concentrations of normally nontoxic contaminants to toxic levels at locationswhere transported material is deposited.

Planting of vegetation at contaminated sites, particularly abandoned strip miningsites, will significantly reduce the erodibility of the soils by water and wind; densityof vegetation will effectively hold the soil and provide a stable cover against erosion.An excellent example of phytostabilization is everyone’s family garden where plantshelp to minimize erosion and enhance the stability of the soil.

Another element of phytostabilization is to supplement the system with a varietyof alkalizing agents, phosphates, organic matter, and biosolids to render the metalsinsoluble and unavailable to leaching. Materials with a calcareous character or ahigh pH, such as lime and gypsum, can be added to influence the acidity. Specificbinding conditions can be influenced by adding concentrated Fe, Mn or Al com-pounds. To maintain or raise the organic matter content in the soils, various materialssuch as humus or peat materials, manure, or mulch can be added.

This chemical alteration should be quickly followed by establishing a plant coverand maximizing plant growth. The amendments sequester the metals into the soil matrixand plants keep the stabilized matrix in place, minimizing wind and water erosion.

5.3.4 Phytovolatilization

Phytovolatilization is the uptake and transpiration of a contaminant by a plant,with release of the contaminant or a modified form of the contaminant to theatmosphere from the plant. Phytovolatilization occurs as growing trees and otherplants take up water and organic and inorganic contaminants. Some of these con-taminants can pass through the plants to the leaves and volatilize into the atmosphereat comparatively low concentrations (Figure 5.3). Many organic compounds tran-spired by a plant are subject to phytodegradation.



Thus far, phytovolatilization has mainly been applied to groundwater contami-nation. However, the potential exists for application to soil, sediments, and othercontamination and needs some careful applications.

2

The state of science with respectto phytovolatization can be summarized as follows:

2,17

• Contaminants could be transformed to less toxic forms (e.g., elemental Hg anddimethyl selenite gas).

• The contaminant or a hazardous metabolite might accumulate in vegetation.• Significant reductions of TCE, TCA, and carbon tetrachloride have been achieved

in experimental studies.• Poplars, alfalfa (

Medicago sativa

), and black locust species have been studied toevaluate phytovolatilization.

• Indian mustard and canola have been used in phytovolatilization studies of Se.

2

Selenium (as selenate) was converted to less toxic dimethyl selenite gas andreleased to the atmosphere. Kenaf and tall fescue have also been used to take upSe, but to a lesser degree than canola.

• A weed from the mustard family (

Arabidopsis thaliana

), genetically modified toinclude a gene for mercuric reductase, converted mercuric salts to metallic mercuryand released it to the atmosphere.

2

• Groundwater must be within the influence of plant (usually a tree) roots and soilmust be able to transmit sufficient water to the plant.

• Climatic factors such as temperature, precipitation, humidity, solar radiation, andwind velocity can affect transpiration rates and thus the rate of phytovolatilization.

• Improved methods for measuring phytovolatilization, diurnal and seasonal varia-tions, and precipitation vs. groundwater use need to be developed.

• Significant research needs to be focused on modeling impacts of vegetation suchas transpiration stream concentration factors, canopy effects, and root concentra-tion factors.

5.3.5 Rhizodegradation

Rhizodegradation (also called phytostimulation, rhizosphere biodegradation,enhanced rhizosphere biodegradation, or plant-assisted bioremediation/degradation)is the breakdown of contaminants in the soil through microbial activity enhancedby the presence of the rhizosphere (Figure 5.4). Microorganisms (yeast, fungi, and/orbacteria) consume and degrade or transform organic substances for use as nutrientsubstances. Certain microorganisms can degrade organic substances such as fuelsor solvents that are hazardous to humans and ecoreceptors and convert them intoharmless products through biodegradation. Natural substances released by plant roots— such as sugars, alcohols, and acids — contain organic carbons that act as nutrientsources for soil microorganisms; these additional nutrients stimulate their activity.Rhizodegradation is aided by the way plants loosen the soil and transport oxygenand water to the area. Plants also enhance biodegradation by other mechanisms suchas breaking apart clods and transporting atmospheric oxygen to the root zone.

Soil adjacent to the root contains increased microbial numbers and populations.

15

It is common knowledge that the number of bacteria in the rhizosphere is as muchas 20 times that normally found in nonrhizosphere soil (Figure 5.4). Short gramnegative rods (specifically

Pseudomonas, Flavobacterium,

and

Alcaligens

) are most



commonly found in the rhizosphere.

15

The increased microbial numbers are primarilydue to the presence of plant exudates and sloughed tissue that serve as sources ofenergy, carbon, and other growth factors. The products excreted by plants includeamino acids, carboxylic acids, carbohydrates, nucleic acid derivatives, growthfactors, and enzymes. The activity of microorganisms in the root zone stimulatesroot exudation further stimulating microbial activity.

16

Several studies have evaluated the effect of plants and the associated rhizosphereon the fate of petroleum contaminants.

2,4,15

For the most part, the presence of plantsenhanced the degradation of contaminants. Also, in studies using

14

C-labeled con-taminants in closed plant chambers, mineralization was greater in rhizosphere soilsthan in unvegetated soils, indicating that the bioavailability of the contaminant washigher in the rhizosphere.

15

Studies using deep rooted prairie grasses to remediate soils contaminated withPAH suggest that the roots of these perennial grasses may be more effective atstimulating the rhizosphere microflora due to their fibrous nature. Fibrous roots offermore root surface area for microbial colonization than other roots and result in alarger microbial population in the contaminated soil. Big bluestem (

Andropogongerardii

), indian grass (

Sorghastrum nutans

), switch grass (

Panicum virgatum

),Canada wild rye (

Elymus canadensis

), little bluestem (

Schizachyrium scoparius

),side oats grama (

Bouteloua curtipendula

), western wheatgrass (

Agropyron smithic

),and blue grama (

Bouteloua gracilis

) are some of the species known to enhancedegradation of petroleum compounds. Crested wheatgrass (

Agropyron desertorum

)is known to degrade PCP contaminated soils.

15

Alfalfa (

Meticago sativa

), fescue(

Festuca anundinacea

), big bluestem (

Andropogon gerardii

), and sudan grass

Figure 5.4

Rhizodegradation and associated processes in the root zone.

Enhanced rhizosphere biodegradation - Supply of nutrients, cometabolites - Transport and retention of water - Aeration

Root respiration

Sloughing

Enzymes dehalogenase nitroductase

Uptake

Root intrusion

Soil dessication



(Sorghum vulgare sudanense) are known to enhance the degradation of PAH com-pounds in the rhizosphere. The degradation rates among various PAHs studiedcorrelated with the water solubility of the compound with the more soluble com-pound, showing the highest degradation.

Cometabolic transformation of chlorinated solvents and other compounds alsohas been reported in the literature.2 Wherever significant cometabolic transforma-tions took place, the following enzyme systems were present: dehalogenase, nitrore-ductase, peroxidase, laccase, nitrylase, and oxygenase.

The rhizosphere is often divided into two general areas: the inner rhizosphereat the very root surface and the outer rhizosphere embracing the immediately adjacentsoil. The microbial population is larger in the inner zone where biochemical inter-actions are most pronounced and root exudates are concentrated. In addition to plantexudates, the rapid decay of fine-root biomass can also become an important additionof organic carbon to soils. A recent report considers some strategies for engineeringplants to improve bioremediation in the root zone. One of the simpler approachesis to make use of the organism Agrobacterium rhizogenes to induce a state called“hairy root disease.” Depending on virulence of the strain used, the extent of rootproduction is variable, but generally, infection leads to a significant enhancement ofrooting without obvious detrimental effects on the host plant. Increased root masshas the apparent advantage of increasing the surface area available for microbialcolonization. Root exudation may be increased in proportion to increase in root area.Such rhizosphere enhancements could improve bioremediation potential of the plant-microbial system. It is suggested that when water is not freely available in unlimitedquantities, increased root mass could lead to greater water uptake, and hence greatercontaminant mobilization and potential degradation.

Different plant species often establish somewhat different subterranean floras(Figure 5.5). The differences are attributed to variations in rooting habits, tissuecomposition, and excretion products of the plant. The primary root population is

Figure 5.5 Examples of different root depths.

Alfalfa 4-6 ft.Grasses 2 ft. Indian

Mustard 1 ft.

Poplar Trees 15 ft.



determined by the habitat created by the plant; the secondary flora, however, dependsupon the activities of the initial population. The age of the plant also alters themicrobial population in the rhizosphere. Roots also harbor mycorrhizae fungi, whichmetabolize organic contaminants. These fungi, growing in symbiotic associationwith the plant, have unique enzymatic pathways, similar to white rot fungus enzymesthat help to degrade organics that could not be transformed solely by bacteria.

In summary, plants provide exudates that offer an excellent habitat for increasedmicrobial populations and pump oxygen to roots, a process ensuring aerobic trans-formations near the root that otherwise may not occur in bulk soil. Due to thepresence of certain primary substrates in the exudate system, anaerobic cometabolictransformations may also take place in the rhizosphere. Typical microbial populationin the rhizosphere comprise: 5 × 106 bacteria, 9 × 105 actinomycetes, and 2 × 103

fungi per gram of air dried soil.The state of science in phytodegradation can be summarized as follows:

• Contaminant degradation can be achieved in situ, which is the biggest advantage.• Translocation of the contaminant to the plant or atmosphere is less likely than

with other phytoremediation techniques since degradation takes place at the sourceof contamination.

• There are low installation and maintenance cost(s) since no harvesting and disposalare required.

• Various microorganism species and enzymes have been isolated which degradedifferent contaminants.

• Analytical methods to better quantify treatment efficiency and success areimproving.

• Field management techniques for nutrients, water, and plant selection areadvancing.

• TPH and PAHs up to hundreds of ppm have been studied in the field with varyingsuccess.2

• Degradation of various pesticides (atrazine, metolachlor, parathion, diazinon, and2,4-D, 2,4,5-T herbicides) has been studied, again with mixed results.2

• TCE, PCP and PCB degradation have also been investigated — again with varyingsuccess.

• More research needs to be done to further elucidate: microbial metabolism in therhizosphere, toxicity towards plants, biodiversity in the rhizosphere, biogeochem-ical optimization in the rhizosphere, and interrelation between biological, chemicaland physical characteristics of the rhizosphere.

The following plants, in addition to the ones discussed previously, have beenused for successful implementation of phytodegradation at field sites:2 1) red mul-berry, crabapple, spearmint, and osage orange that are capable of stimulating PCBdegradation; 2) alfalfa, loblolly pine, and soybean for TCE degradation;3) alfalfafor TCA degradation; and 4) rye, St. Augustine, and white clover for TPH. Growthof hybrid poplar trees for the application of phytodegradation and rhizodegradationis shown in Figures 5.6a, b, and c.



5.3.6 Rhizofiltration

Rhizofiltration is the adsorption or precipitation of contaminants onto plant rootsor the absorption of contaminants into the roots when contaminants are in solution

Figure 5.6a Phytoremediation System, August 6, 1998.

Figure 5.6b Phytoremediation System, September 13, 1999.



surrounding the root zone. In some applications, the plants are raised in greenhouseshydroponically (with their roots in water rather than in soil). Once a large root systemhas been developed, contaminated water is diverted and brought in contact with theplants or the plants, are moved and floated in the contaminated water. The plantsare harvested and disposed as the roots become saturated with contaminants. Plantuptake, concentration and translocation might occur, depending on the contaminant.Exudates from the plant roots might cause precipitation of some metals. Rhizofil-tration first results in contaminant containment, in which the contaminants areimmobilized or accumulated on or within the plant; contaminants are then removedby removing the plant.

Aquatic plants and algae are known to accumulate metals and other toxic ele-ments from solution.18 There are large differences in bioremoval rates due to speciesand strain differences, cultivation methodology, and process control techniques. Inthe past, commercial systems have used immobilized algae biomass for removingradionuclides and other heavy metals in the aqueous phase.19

Naturally immobilized, plants such as attached algae and rooted plants, and thoseeasily separated from suspension, such as filamentus microalgae, macroalgae, andfloating plants, have been found to have high adsorption capacities. In a recent study,one blue green filamentous alga of the genus Phormidium and one aquatic rootedplant, water milfoil (Myriophyllum spicatum), exhibited high specific adsorption forCd, Zn, Ph, Ni, and Cu.18

It has been reported that porous beads containing immobilized biological mate-rials such as sphagnum peat moss can be used for extracting metals dissolved in theaqueous phase.20 The beads designated as BIO-FIX beads readily adsorbed Cd, Pb,and other toxic metals from dilute waters. In one recent study, it was reported that

Figure 5.6c Phytoremediation System, August 22, 2000.



Saccharomyces cerevisiae yeast biomass, when treated with a hot alkali, exhibitedan increase in its biosorption capacity for heavy metals.21 It was also reported thatcaustic treated yeast immobilized in alginate gel could be reactivated and reused toremove Cu, Cd, and Zn in a manner similar to the ion exchange resin.

Phytoremediation applications are summarized in Tables 5.2a and b based oncontaminant fate, degradation, extraction, containment type, or a combination ofthese applications. In the soil–plant–atmosphere continuum, a specific contaminantcan be remediated at specific points along this continuum by different phytoreme-diation mechanisms.

Table 5.2a Types of Phytoremediation for Organic Constituents

Type of Phytoremediation Process Involved Contaminant Treated

1. Phytostabilization Plants control pH, soil gases, and redox conditions in soil to immobilize contaminants. Humification of some organic compounds is expected.

Expected for phenols, chlorinated solvents (tetrachloromethane and trichloromethane) and hydrophobic organic compounds

2. Rhizodegradation (phytostimulation, rhizosphere bioremediation, or plant-assisted bioremediation)

Plant exudates, root necrosis, and other processes provide organic carbon and nutrients to spur soil bacteria growth by two or more orders of magnitude. Exudates stimulate degradation by mycorrhizal fungi and microbes. Live roots can pump oxygen to aerobes and dead roots may support anaerobes.

Polyaromatic hydrocarbons, BTEX, and other petroleum hydrocarbons, perchlorate, atrazine, alachlor, polychlorinated biphenyl (PCB), and other organic compounds

3. Rhizofiltration (contaminant uptake)

Compounds are taken up or sorbed by roots (or sorbed to algae and bacteria).

Hydrophobic organic chemicals

4. Phytodegradation (phytotransformation)

Aquatic and terrestrial plants take up, store, and biochemically degrade selected organic compounds to harmless byproducts, products used to create new plant biomass, or byproducts that are further broken down by microbes and other processes to less harmful products. Reductive and oxidative enzymes may be used in series in different parts of the plant.

Munitions (TNT, DNT, HMX, nitrobenzene, picric acid, nitrotoluene), atrazine, halogenated compounds (tetrachloromethane, trichloromethane, hexachloroethane, carbon tetrachloride, TCE, tetrachloroethane, dichloroethant), DDT and other chlorine and phosphorus based pesticides, phenols, and nitrites

5. Phytovolatilization Volatile organic compounds are taken up and transpired. Some recalcitrant organic compounds are more easily degraded in the atmosphere (photodegradation).

Chlorinated solvents (trichloroethane), organic VOCs, BTEX, MTBE



5.3.7 Phytoremediation for Groundwater Containment

Phytoremediation can be applied for containment of contaminated groundwaterunder the right hydrogeologic conditions such as sites with shallow groundwaterdepths. In general, favorable economics is one factor in phytoremediation’s favor,particularly in contrast to the high cost of operation and maintenance of conventionalgroundwater treatment systems. Furthermore, the high pumping rates of many deeprooted trees may make them more efficient at removing water at low permeability sites.

Phreatophytes (like willows, cottonwood, and hybrid poplar), which take up and“process” large volumes of soil water are good candidates for phytoremediationapplications specifically for groundwater containment. For example, a single willowtree on a hot summer day transpires more than 5000 gallons of water, and a hybridpoplar can transpire about 50 to 350 gallons per day.23

Phytoremediation of groundwater plumes is preferred when the contaminantsare water soluble, leachable organics, and inorganics present at concentrations thatare not phytotoxic. Hydraulic control by plants can occur only within the root zoneor within a depth influenced by roots; the placement depth of roots during plantingcan be varied. Root depth, early tree growth, and nutrient uptake were enhanced byplacing poplar tree root balls closer to shallow groundwater during planting.23

The primary considerations for selecting phytoremediation for hydraulic controlas the method of choice are the depth and concentration of contaminants that affectplant growth. Soil texture and degree of saturation are also influential factors.Planning technique and materials can extend the influence of plants through non-saturated zones to water-bearing layers.

As mentioned earlier, phreatophytes such as poplars are capable of extendingtheir roots into aerobic water tables. For example, the roots of poplars growing

Table 5.2b Types of Phytoremediation for Inorganic Constituents

Type of Phytoremediation Process Involved Contaminant Treated

1. Phytostabilization Plants control pH, soil gases, and redox conditions in soil to immobilize contaminants. Humification of some organic compounds is expected.

Proven for heavy metals in mine tailing ponds

2. Rhizofiltration (contaminant uptake)

Compounds are taken up or biosorbed by roots (or sorbed to algae and bacteria).

Heavy metals and radionuclides

3. Phytoaccumulation (phytoextraction or hyperaccumulation)

Metals and organic chemicals taken up by the plant with water, or by cation pumps, sorption and other mechanisms.

Nickel, zinc, lead, chromium, cadmium, selenium, other heavy metals radionuclides

4. Phytovolatilization Volatile metals are taken up, changed in species, and transpired.

Mercury and selenium



alongside streams can easily be observed intertwined in the stream bottom. Thedegree to which poplar roots would penetrate the saturated zone cannot be easilyestimated. If their access to soil moisture from precipitation is limited, poplars willdraw large amounts of water from the top of saturated aquifer. Evapotranspirationwill draw down the water table below the trees similar to a pump and treat system(Figures 5.7a and b). Simulations of a proposed design can be carried out based onextent of contamination, hydrogeological data, past precipitation and infiltrationrecords, and evapotranspiration data.

A big advantage of phytoremediation over conventional pump and treat systemsis the ability of the roots to penetrate the microscopic scale pores in the soil matrix.Contaminants adsorbed or trapped in these micropores are impacted minimally ornot at all by the pump and treat system. In the case of phytoremediation, the rootscan penetrate these micropores for contaminant removal.

5.3.8 Phytoremediation of Dredged Sediments

Dredged material is nothing more than displaced topsoil that enters and iseventually removed from navigable waterways. Contaminant discharges into water-ways over time result in contamination of bottom sediment. Dredged sediments areusually stored in confined disposal facilities (CDF).

24

The application of phytoremediation to dredged material presents some chal-lenges unique to dredged material. Dredged sediments come from an aquatic envi-ronment and are initially wet and anaerobic after placement in a CDF. Subsequentdrying and oxidation depend on dewatering and management techniques. Dryingand oxidation of surface layers may result in physicochemical changes that mayaffect plant establishment and contaminant mobility. Although the surface layer of

Figures 5.7a

Placement of root ball with time due to maturation of the tree.

AboveCapillary

Fringe

At CapillaryFringe

In Capillary Fringeand Groundwater Table



dredged sediments in a CDF may be dry and aerobic, deeper layers may remainanaerobic. Saltwater dredged sediments provide another level of difficulty for veg-etation and in most cases must be leached to reduce soluble salt levels. Dredgedmaterial management is further complicated by the potential of elevated concentra-tions of multiple contaminants. The selection of plant species and methods ofestablishment will be determined by these factors. Common contaminants presentin dredged sediments are metals, PAHs, polychlorinated phenols, PCBs and otherheavy molecular weight compounds. The current state of knowledge indicates thatphytoremediation of dredged sediments would not be as readily effective as appli-cation to more heavily contaminated industrial sites.

5.4 PHYTOREMEDIATION DESIGN

The design of a phytoremediation system varies according to contaminants,conditions at the site, level of cleanup required, and plants used. A thorough sitecharacterization should provide the needed data to design any type of remediationsystem. Clearly, phytoextraction has different design requirements from phytostabi-lization or rhizodegradation. Nevertheless, it is possible to specify a few designconsiderations that are part of most phytoremediation efforts (Figures 5.8a, b, andc). Site characterization data will provide the information required for the designerto develop a properly functioning system. The design considerations include con-taminant levels; plant selection; treatability; irrigation, agronomic inputs (P N, K,salinity, zinc, etc.), and maintenance; groundwater capture zone and transpirationrate; and contaminant uptake rate and clean-up time required.

Other factors to be considered during the evaluation, design, and implementationphases of phytoremediation at a contaminated site are:

Figures 5.7b Predicted groundwater flow conditions at maturation of tree growth.

Groundwaterflow

Groundwater table elevation contours

Zone of tree plantation

30.0

30.5

31.0

30.0

29.028.0



• Soil Water — The most crucial factor in a plant’s life is water, which links it tothe soil via roots and serves as a vehicle for nutrient transport. Water also controlsthe exchange of gases and moderates soil temperature changes. Plant availablewater is held in the soil between the field capacity and permanent wilting point.Plant roots can extract water at lower potentials, depending upon the plant typeand arable environment. Root growth rates are controlled by the presence ofcontinuing supplies of water to maintain hydrostatic pressure in the elongating

Figure 5.8a Decision tree for phytoremediation in soil.



cells of the root, and metabolites for cell wall construction. Water flows radiallyinto elongating root cells only when the cell’s total water potential is lower thanthe combined osmotic and matric potentials of the soil. Soil water content willinfluence plant biomass growth.

Figure 5.8b Decision tree for phytoremediation groundwater.

Decision Tree for Phytoremediation

Groundwater

YES N O

N O YES

YESN O

Will the plants be used for hydrauliccontrol ONLY (prevent water from

REACHING the contaminated zone)? YES

N OWill the water be mechanically pumped

and applied to the phytoremediationsystem? YES

N O

Will state regulations allow thistype of phytoremediation? YES

N O

N O

YES

YESN O

YES

N ON O YES

N O YES

N O YES

YESN O

YESN O

YES

N O

YESN O

YESN O

YESN O

N O YES

YES N O

Phytoremediation has the potential tobe effective at the site

Phytoremediation is NOT an optionat the site;consider other options

Will the climate support the proposed plants?

Is time or space a constraint?

Is the contaminant physically within the range of the proposed plant (typically lessthan 10- 20 feet bgs for Salix species - willows, cottonwoods, poplars) ?

Is the contaminant at phytotoxicconcentrations (this may require agreenhouse dose-response test)?

Will the rhizosphere microbes and plant-exuded enzymes degrade the targetcontaminants in the rhizosphere and are the metabolic products acceptable?

Will the plant degrade the contaminantafter uptake and are the metabolic

products acceptable?

Is the log Kow of the contaminant ormetabolic poducts between 1 and 3.5(will uptake occur)?

Will the plant accumulate the contaminantor metabolic products after uptake?

Will the plants transpire the contaminantor metabolic products?

Is the level of accumulation acceptable forthis site throughout the growth of the plant?

Can controls be put in place to prevent thetransfer of the contaminant or metabolic

products from a plant to humans/animals ?

Are the quantity and rate of transpirationacceptable for this site?

Can engineering controls make itacceptable?

Is the final disposition of the contaminantor metabolic products acceptable?

Can the contaminant or metabolic productbe immobilized to acceptable levels ?

Does the plant material constitute a waste if harvested?

Can the plant waste be economically disposed?



• Soil Air — Plants need molecular oxygen to respire and convert carbohydrates toCO2 and H2O. This is an exothermic reaction and releases respiratory energyutilized for many plant processes. The disappearance of O2 triggers a sequenceof changes in the biogeochemical properties of the soil; the absence of O2 aloneis sufficient to alter plant metabolism profoundly. Suboptimal concentrations of

Figure 5.8c Decision tree for phytoremediation sediments.

Decision Tree for Phytoremediation

Sediments

Will the climate support the proposed plants?YES NO

Is time or space a constraint?NO YES

Can the sediments be treated in place (wetlands)?YES NO Are the sedimentsto be dredged? YES

NO

Are there hotspots that can beremoved or treated? YES

NO Is the contaminant at phytotoxic concentrations (thismay require a greenhouse dose-response test)?NO

YES

Can the contaminant or metabolic productsbe immobilized to acceptable levels ?YES

NOIs the final disposition of the contaminantor metabolic products acceptable?YES

NO

Does the plant material constitute a waste if harvested? NO YES

Can the plant waste be economically disposed?YES NO

Phytoremediation has the potential tobe effective at the site

Phytoremediation is NOT an option at the site; consider other options

Will the regulatory statutes allow the dredged sediments to be treated as a soil?YES NO

Is there strong public support to treat the sediment as a soil?YES NO

Is the contaminant physically within the range of the proposed plant (typically less than1- 2 feet bgs)?

YES NO

Will the rhizosphere microbes and plant-exuded enzymes degrade the targetcontaminants in the rhizosphere and are the metabolic products acceptable?YES

NO

Is the log Kow of the contaminant or metabolicproducts between 1 and 3.5 (will uptake occur)?

NO YES Will the plant degrade thecontaminant after uptake and are

the metabolic products acceptable?YES

NO

Is the level of accumulation acceptable forthis site throughout the growth of the plant?YES

NO

Will the plant accumulate the contaminantor metabolic products after uptake?NO

YES

Can controls be put in place to prevent thetransfer of the contaminant or metabolic

products from a plant to humans/animals?YES

NO

Will the plants transpire thecontaminant or metabolic products?NO

YES

Are the quantity and rate oftranspiration acceptable for this site?YES

NO

Can engineering controls make itacceptable?YES NO



O2 in the soil occur because of interactions among soil properties such as porosity,water content, temperature, surface water infiltration, and continuity of air filledpores with biotic activity.

• Soil Temperature — Temperature influences plant processes at the cellular level,such as osmotic potential, hydration of ions, stomatal activity and transpiration,Gibbs free energy available for work, membrane permeability, solute solubilities,diffusion, and enzymatic activities. Temperature and cultivar strongly influencethe establishment of plants. Low temperatures also decrease metabolic activityand root growth.

• Physical Impedance — Physical impedance, sometimes called mechanical imped-ance or excessive soil strength, can severely affect normal root growth patterns. Suchimpedances result from increased soil bulk density, increased cohesion and frictionbetween soil particles, reduction in soil water content, frost-heave action of soil, andpresence of permafrost within the root zone. Under an excessive soil strength envi-ronment, roots enter the soil volume where pore sizes are larger than the root tip.Conversely, if pore sizes are too small for entry of the main root but not for thelaterals, then laterals proliferate and produce a highly branched root system.

• Topography — Topography is a critical factor because it is a key factor in deter-mining runoff velocity and erosion. In general, the amount of soil erosion increasesmanifold with increasing degree and length of slope. Contaminated sites withslopes greater than 10% are often not suitable for phytoremediation withoutsurface modification because of excessive erosion.

• Soil pH — Plant roots are damaged at pH lower than 4.0. The roots are shortened,thickened, fewer in number, and dull brown or gray in color. Salinity is anotherchallenge to phytoremediation applications in the field. Soluble salts reduce thetotal water potential of the soil solution, thus tending to reduce the potentialdifference between soil water and the atmosphere. Excessive soil salinity reducesroot elongation and upsets hormonal balance, as well as altering soil structurethat, in turn, affects plant growth.

5.4.1 Contaminant Levels

During the site characterization phase the concentration level of the contaminantsof concern will be established. High levels of contamination may eliminate phytore-mediation as a treatment option. Plants are not able to treat all contaminants. Thecomposition of organic compounds (structure, log Kow, degree of weathering andboiling point range) and degree of adsorption are important factors in phytoreme-diation. It is important to understand the range of contaminants that can be treatedusing phytoremediation. In addition to knowing contaminants and their concentra-tions, the depth of the contaminants must be known. The primary consideration inthis area is that the contaminant concentrations cannot be phytotoxic or causeunacceptable impacts on plant health or yield. Higher concentrations of contaminantsmight be tolerated more readily by plants than by soil microorganisms.

5.4.2 Plant Selection

The goal of the plant selection process is to choose a plant species withsuitable characteristics for growth under site conditions that meet the objectives of



phytoremediation. Native, crop, forage, and other types of plants that can grow underregional and climatic conditions should be preferred. Plants are selected accordingto the application and the contaminants of concern. For phytotransformation oforganic compounds, the design requirements are that vegetation is fast growing andhardy, easy to plant and maintain, utilizes a large quantity of water by evapotrans-piration, and transforms the contaminants of concern to nontoxic or less toxicproducts. In temperate climates, phreatophytes (e.g., hybrid poplar, willow, cotton-wood, and aspen) are often selected because of fast growth, a deep rooting abilitydown to the level of groundwater, large transpiration rates, and the fact that they arenative throughout most of the country. A screening test or knowledge from theliterature of plant attributes will aid the design engineer in selection of plants.

Plants used in phytoextraction include sunflowers and Indian mustard for lead;Thlaspi spp. (Pennycress) for zinc, cadmium, and nickel; and sunflowers and aquaticplants for radionuclides. Aquatic plants are used in constructed wetlands applica-tions. The two categories of aquatic plants used are emergent and submerged species.Emergent vegetation transpires water and is easier to harvest if required. Submergedspecies do not transpire water but provide more biomass for the uptake and sorptionof contaminants.

5.4.3 Treatability

Treatability or plant screening studies are recommended prior to designing aphytoremediation system. If the decision tree flowcharts indicate phytoremediationis an applicable technology for a site, a plant scientist should assist in the treatabilitystudies which assure concerned parties that the phytoremediation system will achievedesired results. Treatability studies provide toxicity and transformation and assessthe fate of the contaminants in plant system. Different concentrations of contaminantare tested with proposed plant species. Volatile organic compounds are oftentranspired to the atmosphere by plants; calculations will predict the amount and typeof material transpired.

5.4.4 Irrigation, Agronomic Inputs, and Maintenance

Irrigation of plants ensures a vigorous start to the system even in drought.Hydrologic modeling may be required to estimate the rate of percolation to ground-water during irrigation conditions. Irrigation should be withdrawn if the area receivessufficient rainfall to sustain the plants. Agronomic inputs include the nutrients nec-essary for vigorous growth of vegetation and rhizosphere microbes. The soil mustbe analyzed and then items such as nitrogen, potassium, phosphorous, aged manure,sewage sludge compost, straw, and/or mulch are added as required to ensure thesuccess of the plants. Maintenance of the phytoremediation system may includeadding fertilizer, agents to bind metals to the soil, or chelates to assure plant uptakeof the contaminants. Replanting may be required due to drought, disease, insects,or animals killing plants.



5.4.5 Groundwater Capture Zone and Transpiration Rate

For applications involving groundwater remediation, a capture zone calculationcan be used to estimate whether the phytoremediation pump (trees) can be effectiveat entraining the plume of contaminants. The goal is to create a water table depressionwhere contaminants will flow to the vegetation for uptake and treatment. Organiccontaminants are not taken up at the same concentrations in the soil or groundwaterbecause membranes at the root surface reduce the uptake rate. Although it is possibleto estimate the uptake rate of contaminants, the calculation is beyond the scope ofthis chapter.

REFERENCES

1. McCutcheon, S. C., USEPA, personal communications, 1999, 2000.2. USEPA, Introduction to Phytoremediation, EPA/600/R-99/107, Washington D.C.,

February, 2000.3. Schnoor, J. L. et al., Phytoremediation of organic and nutrient contaminants, Environ.

Sci. Technol., 29, 1620–1631, 1995.4. McCutcheon, S. C., Phytoremediation of organic compounds: science validation and

field testing, in Workshop on Phytoremediation of Organic Wastes, Kovalick, W. W.and Olexsey, R., Eds., Ft. Worth, TX, December, 1996.

5. Shahandeh, H. and Hossner, L. R., Enhancement of Cr (III) phytoaccumulation, Int.J. Phytoremed., 2, 269–286, 2000.

6. Brooks, R. R., Plants That Hyperaccumulate Heavy Metals, CAB International, NewYork, NY, 1998.

7. McCutcheon, S. C., The science and practice of phytoremediation, in Phytoremedi-ation: State of the Science Conf., Boston, MA, May, 2000.

8. Cornish, J. E. et al., Phytoremediation of soils contaminated with toxic elements andradionuclides, in Bioremedation of Inorganics, Hinchee, R. E. et al., Eds., BattellePress, Columbus, OH, 1995.

9. Brown, S. L. et al., Zinc and cadmium uptake by hyperaccumulator Thlaspi caerule-scens and metal tolerant Silene vulgaris grown on sludge amended soils, Environ.Sci. Technol., 29, 1581–1590, 1995.

10. Bishop, J. E., Pollution fighters hope a humble weed will help reclaim contaminatedsoil, Wall Street Journal, August 7, 1995.

11. Kramer, et al., Free histidine as a metal chelator in plants that accumulate nickel,Nature, 379, 635–638.

12. Newman, L. A. et al., Uptake and biotransformation of trichloroethylene by hybridpoplars, Environ. Sci. Technol., 31, 1062–1067, 1997.

13. Conger, R. M. and Portier, R., Phytoremediation experimentation with the herbicidebentazon, Remediation, 7, 19–37, 1997.

14. Narayanan, M., Davis, L. C., and Erickson, L. E., Fate of volentile chlorinated organiccompounds in a laboratory chamber with alfafa plants, Environ. Sci. Technol., 29,2437–2444, 1995.

15. Fiorenza, S., Oubre, C. L., and Ward, C. H., Phytoremediation of HydrocarbonContaminated Soil, Lewis Publishers, Boca Raton, Florida, 2000.

16. Alexander, M., Introduction to Soil Microbiology, John Wiley & Sons, New York,1977.



17. Stomp, A. M. et al., Genetic strategies for enhancing phytoremediation,

Ann. NYAcad. Sci.

, 721, 481–491, 1994.18. Want, T. C., Weissman, J. S., Ramesh, G., Varadarajan, R., and Benemann, J. R.,

Bioremoval of toxic elements with aquatic plants and algae, in

Bioremediation ofInorganics

, Hinchee, R. E. Means, J. L., and Burns, D. R., Eds., Battelle Press,Columbus, OH, 1995.

19. Feiler, H. D. and Darnall, D. W., Remediation of Groundwater Containing Radionu-clides and Heavy Metals using Ion Exchange and the Alga SORD Biosorbent System,Final Report under Contract No. 02112413, DOE/CH-9212, 1991.

20. Jeffers, T. H., Bennett, P. G., and Corwin, R. R., Biosorption of metal contaminantsusing immobilized biomass: field studies, Report of Investigations 9461, Bureau ofMines, US Department of the Interior, 1993.

21. Lu, Yongming and Wikins, E., Heavy metal removal by caustic-treated yeast immo-bilized in alginate, in

Bioremediation of Inorganics

, Hinchee, R. E., Means, J. L.,and Burris, D. R., Eds., Battelle Press, Columbus, OH, 1995.

22. Susarla, S. et al., Phytotransformation of perchlorate using parrot feather,

Soil andGroundwater Cleanup

, March, 1999.23. Gatliff, E., personal communication, 200024. DOE, Phytoreclamation of dredged material; a working group summary, Technical

Note, DOER-C9, November, 1999.

CREDIT

Figures 5.1, 5.8a,b,c, and Tables 5.2a,b were reproduced from

PhytoremediationDecision Tree

, prepared by Interstate Technology and Regulatory Cooperation WorkGroup, Phytoremediation Work Team, November 1999.

L1282_C05_frame Page 268 Thursday, March 6, 2003 12:14 PM

269

CHAPTER

6

Constructed Treatment Wetlands

CONTENTS

6.1 Introduction ..................................................................................................2706.1.1 Beyond Municipal Wastewater ........................................................2726.1.2 Looking Inside the “Black Box” .....................................................2736.1.3 Potential “Attractive Nuisances”......................................................2746.1.4 Regulatory Uncertainty and Barriers ...............................................275

6.2 Types of Constructed Wetlands ...................................................................2766.2.1 Horizontal Flow Systems.................................................................2766.2.2 Vertical Flow Systems......................................................................277

6.3 Microbial and Plant Communities of a Wetland.........................................2786.3.1 Bacteria and Fungi ...........................................................................2786.3.2 Algae ................................................................................................2796.3.3 Species of Vegetation for Treatment Wetland Systems...................279

6.3.3.1 Free-Floating Macrophyte-Based Systems.......................2826.3.3.2 Emergent Aquatic Macrophyte-Based Systems ...............2846.3.3.3 Emergent Macrophyte-Based Systems with Horizontal

Subsurface Flow ...............................................................2856.3.3.4 Emergent Macrophyte-Based Systems with Vertical

Subsurface Flow ...............................................................2856.3.3.5 Submerged Macrophyte-Based Systems ..........................2856.3.3.6 Multistage Macrophyte-Based Treatment Systems..........287

6.4 Treatment-Wetland Soils..............................................................................2876.4.1 Cation Exchange Capacity...............................................................2896.4.2 Oxidation and Reduction Reactions ................................................2906.4.3 pH .....................................................................................................2926.4.4 Biological Influences on Hydric Soils.............................................2926.4.5 Microbial Soil Processes..................................................................2926.4.6 Treatment Wetland Soils ..................................................................293

L1282/C06/frame Page 269 Monday, June 18, 2001 9:29 AM


6.5 Contaminant Removal Mechanisms ............................................................2946.5.1 Volatilization ....................................................................................2946.5.2 Partitioning and Storage...................................................................2956.5.3 Hydraulic Retention Time................................................................297

6.6 Treatment Wetlands for Groundwater Remediation....................................2996.6.1 Metals-Laden Water Treatment........................................................300

6.6.1.1 A Case Study for Metals Removal ..................................3026.6.2 Removal of Toxic Organics .............................................................306

6.6.2.1 Biodegradation ..................................................................3066.6.3 Removal of Inorganics .....................................................................3096.6.4 Wetland Morphology, Hydrology, and Landscape Position............309

References..............................................................................................................310

Creating or constructing a natural wetland sounds like an oxymoron, but thisdoesn’t mean that an “unnatural wetland” is by definition bad. It doesn’t meanwe can’t mimic Mother Nature in giving natural birth to a desirable wetland.Constructed rice paddies have been responsible for feeding more people thanany other enterprise on earth.

6.1 INTRODUCTION

Natural wetlands are land areas that are wet during part or all of the year becauseof their location in the landscape. Historically, wetlands were called swamps,marshes, bogs, fens, or sloughs, depending on existing plant and water conditionsand on geographic setting. Wetlands are frequently transitional between uplands(terrestrial systems) and continuously or deeply flooded (aquatic) systems. They arealso found at topographic lows (depressions) or in areas with high slopes and lowpermeability soils (seepage slopes). In other cases, wetlands may be found at topo-graphic highs or between stream drainages when land is flat and poorly drained(blanket bogs). In all cases, the unifying principle is that wetlands are wet longenough to alter soil properties because of the chemical, physical, and biologicalchanges that occur during flooding, and to exclude plant species that cannot growin wet soils.

1

The structural components of natural wetland ecosystems are shown in Figure6.1. These components are highly variable and depend on hydrology, underlyingsediment types, water quality, and climate. Starting with the unaltered sediments orbedrock below the wetlands, these typical components are

1

•

Underlying strata

— unaltered organic, mineral, or lithic strata, typically saturatedwith or impervious to water and below the active rooting zone of the wetlandvegetation

•

Hydric soils

— the mineral-to-organic soil layer of the wetland, infrequently tocontinuously saturated with water and containing roots, rhizomes, tubers, funnels,burrows, and other active connections to the surface environment


CONSTRUCTED TREATMENT WETLANDS 271

•

Detritus

— the accumulation of live and dead organic material in a wetland,consisting of dead emergent plant material, dead algae, living and dead animals(primarily invertebrates), and microbes (fungi and bacteria)

•

Seasonally flooded zone

— the portion of wetland seasonally flooded by standingwater and providing habitat for aquatic organisms including fish and other verte-brate animals, submerged and floating plant species that depend on water forbuoyancy and support, living algae, and populations of microbes

•

Emergent vegetation

— vascular, rooted plant species containing structural com-ponents that emerge above the water surface, including both herbaceous andwoody plant species

Natural wetlands have been used as convenient wastewater discharge sites foras long as sewage has been collected (at least 100 years in some locations). Examplesof old treatment wetland sites can be found in Massachusetts, Wisconsin, Florida,and Ontario.

Judging by the growing number of wetlands built for wastewater treatmentaround the world, this “natural” technology seems to have firmly established roots.After almost 30 years of use in wastewater treatment, constructed “treatmentwetlands” now number over 1000 in Europe and in North America.

1

Marsh-type“surface flow” systems are most common in North America, but “subsurface flow”wetlands, where wastewater flows beneath the surface of a gravel-rock bed, pre-dominate in Europe. This inexpensive, low-maintenance technology is reportedlyin high demand in Central America, Eastern Europe, and Asia. New applications,

Figure 6.1

Structural components of natural wetland ecosystems (adapted from Kadlec etal., 1996).

UnalteredSediment

Hydric Soils

Detritus

Rhizomes

EmergentVegetation

Cypress Kness

SubcanopyTree

Shrub

SeasonalHigh Water

SeasonallyFlooded Zone

Seasonal Low Water

CanopyTree

Buttressed Stem



from nitrate-contaminated groundwater to effluent from high-intensity livestockoperations, are also increasing.

In the U.S., treatment-wetland technology has not yet gained universal regulatoryacceptance; projects are approved on a case-by-case basis. Some states and EPAregions are eager to endorse them, but others are wary of this nontraditional methodof treating wastewater and contaminated groundwater. In part, this reluctance existsbecause the technology is not yet completely understood. Knowledge of how thewetland works is not far enough advanced to provide engineers with detailed pre-dictive models. Because wetlands are natural systems, their performance is variable,subject to the vagaries of changing seasons and vegetative cycles. These treatmentwetlands also pose a potential threat to wildlife attracted to this new habitat withinan ecosystem exposed to potentially toxic compounds.

When utilized for benign, pretreated wastewaters, wetlands do not generally posea threat to human or wildlife health. In these circumstances, there may be significantancillary benefits in terms of habitat creation and beneficial human use. In thosesituations where a potentially hazardous condition exists, the extra expense of agravel media is warranted.

1

Water and associated particulates, organisms, and sedi-ments are then located below ground, and thus out of reach of human and wildlifecontact. Subsurface wetland waters are typically anoxic or anaerobic, which isoptimal for some processes such as sulfide precipitation or denitrification, but unsat-isfactory for other processes, such as nitrification of ammonium nitrogen.

New efforts are underway, however, to place the technology onto firmer scientificand regulatory ground. Long-term demonstration and monitoring field studies arecurrently probing the inner workings of wetlands and their water quality capabilitiesto provide better data on how to design more effective systems. Researchers aredocumenting the fate of toxic compounds in wetlands and the extent to which wildlifemay be exposed to them. A recent study of U.S. policy and regulatory issuessurrounding treatment wetlands has recommended that the federal governmentactively promote this technology and clear the regulatory roadblocks to enable wideruse. Proponents argue that the net environmental benefits of constructed wetlands,such as restoring habitat and increasing wetland inventory, should be considered. Afederal interagency work group is grappling with that recommendation, trying tobalance the benefits and shortcomings of this increasingly popular technology.

6.1.1 Beyond Municipal Wastewater

Constructed wetland systems in North America have been designed predomi-nantly for large-scale treatment of municipal wastewater, ranging from 100,000 to15 million gallons per day.

1,2

The use of treatment wetlands is well established inEurope, where the technology originated with laboratory work in Germany 30 yearsago.

3

Subsurface-flow systems are the norm because they provide more intensivetreatment in a smaller space than marsh-type wetlands — an important designconstraint in countries where open space is limited. The European thrust has beenfor small-scale systems primarily for domestic wastewater treatment; for example,Denmark alone has 150 systems, most in small villages handling domestic waste-water. The term “reed beds” is commonly used for treatment wetlands in Europe.



Since the 1980s, constructed wetlands have also been built to treat other typesof wastewaters, including acid mine drainage, industrial wastewater, agricultural andstorm water runoff, and effluent from livestock operations.

1,2

The petroleum industryis using constructed wetlands to treat a variety of wastewaters from refineries andfuel storage tanks. Food processing and pulp and paper industries are relative new-comers to treatment wetlands. Stormwater runoff also has recently become a focusof research in using constructed wetlands as a treatment method.

While many of the early acid mine drainage treatment systems were marsh-likesurface flow systems, the most recent projects are “passive treatment systems” thatlink several different types of cells — vertical limestone drain as well as vegetatedcells — to sequentially treat particularly “nasty” wastewater with low pH and highmetals content.

2

A wetland system for the treatment of runoff from coal piles atcoal-fired power plants with a pH of 2 and high levels of metals uses a series ofsuccessive alkalinity-producing systems, a rich organic layer over an anoxic lime-stone drain, to reduce the acidity in the wastewater before it flows into wetland cells.

Landfill leachates are a subset of polluted waters requiring substantial levels oftreatment. Leachates vary considerably, depending upon the materials accepted atthe landfill. They may contain large concentrations of volatile and toxic organics,both as individual compounds and as COD, chlorinated organics, metals, and nitrog-enous compounds.

2

Wetland treatment of landfill leachates has been successfullytested at several locations. Cold climate systems are functioning properly in Norway,as well as at several locations in Canada; reed beds are used to treat leachate in theUnited Kingdom, Slovenia, and Poland.

4

Based on current understanding of the effectiveness of wetland treatment ofleachates, several U.S. projects are in planning and design phases. In addition, thereare about a half-dozen other projects in various locations, such as Mississippi,Indiana, Pennsylvania, and West Virginia. Wetlands have been proposed for controlof stormwater runoff from capped landfills.

1,2

Continued growth in the use of treatment wetlands is expected as a result of newregulatory initiatives on nutrient management, including the Clean Water Act’s totalmaximum daily load (TMDL) program. Small- to medium-sized communities tryingto meet new TMDLs in sensitive watersheds for phosphorus or ammonia needsomething that is cost-effective, and wetlands are a good option.

6.1.2 Looking Inside the “Black Box”

The rapid spread and diversification of treatment-wetland technology are runningahead of the mechanistic understanding of how they work. These complex naturalsystems are still, somewhat, a “black box,” according to many in the field. Forexample, the role of plants in transporting oxygen into the root zone to promotenitrification has been demonstrated in the laboratory but not convincingly in thefield, according to many researchers. There is very little data to say whether that isan important factor or whether the plants are more or less passive. It is likely,according to some researchers, that the ratio of open water to vegetated areas ismore important in creating aerobic conditions in a wetland.



Another issue quite often debated is how important the volume of water in awetland is to treatment performance. Is it the bottom of the wetland or the volumeof water that is more important? The data coming in now are on the side of thewetland bottom:

1,2

it apparently does not matter how deep the water is as long asthe soil is wet. That is a surprise to civil engineers, who, for years, have designedtreatment systems based on their volume and hydraulic residence time.

Numerous research efforts, both broad based and focused, are currently gener-ating a great deal of new information on treatment-wetland function.

1,2,5

The exten-sive research activities include gathering conventional water quality data; measure-ments of metals, biotoxicity, and organics; bird surveys; and macroinvertebratesampling. Expanding the species pallet of plants used in treatment wetlands isanother focus of research among researchers in this field. Most constructed wetlandsfor treatment have been built around herbaceous species so far, and many researchersare experimenting with a greater variety of plants to see how water quality changeswhen multispecies systems are used. Many have found that pathogen removal ishigher in a multispecies system than in a single species system. One of the thingsthat may be important in pathogen removal is having multiple types of wetlandcomponents, for example, a duckweed system followed by a subsurface wetland.

5

Looking deeper into the wetland, to the microbes in the soil and around the rootsystems of wetland plants, some researchers are studying the role that bacteria playin trace element removal. Researchers have found that bacteria in the root zone ofbulrush increase the plants’ ability to accumulate and volatilize selenium twofold.They are now working to identify which bacteria are most responsible, and will soonmove to mesocosm studies to see whether seeding the soil with those bacteriaincreases trace element removal.

Some researchers are experimenting with an innovative wetland design — avertical flow system — to solve the oxygen depletion problem and boost nitrifica-tion.

1,2

Effluent flows over a porous surface and percolates through a vegetated sandfilter, which is periodically allowed to dry to reintroduce oxygen to the system.

6.1.3 Potential “Attractive Nuisances”

Aside from research issues surrounding the design and performance of thetreatment wetlands black box, another scientific issue looms large for the future ofthe technology: do treatment wetlands pose a threat to wildlife?

1,5

This question isan important one, since many wetland projects are designed with habitat creationas one of their primary beneficial objectives. It is easier to justify the land use fora constructed wetland if it is also used for habitat restoration.

Research is also being directed toward several critical issues. Some researchersare working to find out exactly where toxic trace elements from wastewater end upin a treatment wetland. They are completing laboratory studies documenting traceelement uptake potential of various wetland plants and identifying where the ele-ments go in the plants: roots, stems, leaves, or plant litter. They are also monitoringseveral active treatment wetlands to track trace elements in the ecosystem: sediment,water, air, plant tissues, and animal tissues.



To address similar habitat-related issues, influent and effluent water have beenanalyzed for potential bioaccumulation and mutagenic activity from organic com-pounds.

5

Toxicity tests were designed to look for physiological impacts on biotaliving in the system. Work also continues on the control of an unplanned threat tohuman health: mosquitoes. Fish have been introduced to the wetlands to consumemosquito larvae, but the density of the particular bulrush variety used may preventthe fish from reaching certain parts of the wetland. Sections of the wetlands can bereconfigured and replanted to raise the water level and give the fish greater access.

6.1.4 Regulatory Uncertainty and Barriers

Treatment wetlands do not appeal to all wastewater engineers because they lackthe traditional “handles” of engineered pollution control systems, are not easy tocontrol, and may be hard to predict. Regulators in the U.S. have similar problemswith treatment wetlands because they do not fit easily into existing regulatorycategories. Surface-flow treatment wetlands can be a point source discharge and aprotected environment at the same time.

No national guidance on the use of treatment wetlands and no uniform acceptanceof them by states exist, according to researchers and consultants. In this atmosphereof regulatory uncertainty, questions abound. Concerns have been expressed that undera strict reading of the Clean Water Act, certain treatment wetlands could be considered“waters of the U.S.,” and thus discharges into them could be tightly regulated.

USEPA’s environmental technology initiative (ETI) treatment wetland policy andpermitting team of representatives from federal, state, and local agencies issued areport in January, 1997, that recommended “changes in regulation and/or policy thatwould facilitate, where appropriate, implementation of beneficial treatment wetlandprojects.”

6,7

It also advocated that “net environmental benefits” of habitat creation,reduced use of energy and treatment chemicals, and recreational value — not justthe water quality impact of a treatment wetland project — should be considered inapproving it.

The report catalogued numerous regulatory and policy issues. Should disinfec-tion of effluent be done at the inlet rather than the outlet of a wetland? When shoulda wetland be lined to protect groundwater? Should treatment wetlands be allowedto mitigate for permitted wetland losses? Under what conditions should constructedtreatment wetlands be considered “waters of the U.S.?” The report also noted thatmore research is needed concerning the “fate and effect of potential wastewatertoxins and ecological risks in treatment wetlands.”

The federal interagency work group, including representatives from USEPAwetlands and wastewater offices, the U.S. Army Corps of Engineers, the NationalOceanic and Atmospheric Administration, the Bureau of Reclamation, and the U.S.Fish and Wildlife Service, was created to take up these issues.

6,7

The question of where treatment wetlands should be sited has been a particularlydifficult regulatory issue, and consensus must be reached on the need to handlewetland systems differently depending on whether their primary purpose is watertreatment or habitat restoration. There is still some disagreement about the habitat



value of treatment wetlands and concerns about the negative impact they could haveon the environment.

USEPA currently is not developing the type of specific guidance documents andformal agency actions recommended in the ETI study to promote the use of treatmentwetlands. Nevertheless, wetlands experts are encouraged because the issues are nowbeing discussed at the national level.

6.2 TYPES OF CONSTRUCTED WETLANDS

6.2.1 Horizontal Flow Systems

The purposeful construction of treatment wetland ecosystems is a relatively newtechnology. Constructed wetlands for pollution control, wastewater treatment, and,recently, for contaminated groundwater treatment are divided into two basic types:free water surface (FWS) and subsurface flow (SSF) wetlands. Both types consistof a channel or a basin with some sort of barrier to prevent seepage and utilizeemergent aquatic vegetation as part of the treatment system. The difference betweenFWS and SSF wetlands is that SSF uses some kind of media as a major component(Figures 6.2a and b). In an FWS treatment wetland, soil supports the roots of theemergent vegetation; water at a relatively shallow depth of 6 to 24 inches flowsthrough the system with the water surface exposed to the atmosphere. Oxygen isprovided by diffusion through the water surface.

An SSF treatment wetland bed contains a suitable depth (1.5 – 3.0 feet) ofpermeable media, such as coarse sand or crushed stone, through which the waterflows. The media also support the root structure of the emergent vegetation. Thesurface of the flowing water is beneath the surface of the top layer of the medium,determined by proper hydraulic design and appropriate flow control structures. Inboth systems the polluted water undergoes physical, biological, and chemical treat-ment processes as it flows through the wetlands.

The rate at which organic contaminants move through wetlands can be deter-mined by several transport mechanisms. These mechanisms often act simultaneouslyon the organics and may include such processes as convection, diffusion, dispersion,and zero- or first-order production or decay.

Figure 6.2a

Free water surface (FWS) wetland.

Inlet

OutletWeir



Currently, constructed wetlands for municipal wastewater treatment are designedbased on the assumptions of plug-flow hydrodynamics and first-order biochemicaloxygen demand (BOD) removal kinetics. The first assumption implies that dispersionin the system is negligible and all the fluid particles have a uniform detention timetraveling through the system. The plug-flow model seems to give a reasonablyaccurate estimate of the performance of SSF-constructed wetlands.

However, some designers have recognized the limitation of using the plug-flowmodel for constructed wetlands design. Three types of hydraulic inefficiencies mayoccur in treatment wetlands: one caused by internal islands and topographical features,a second caused by preferential flow channels on a large-distance scale, and a thirdcaused by mixing effects, such as water delays in litter layers and transverse mixing.

6.2.2 Vertical Flow Systems

Vertical flow constructed wetlands are vegetated systems in which the flow ofwater is vertical rather than horizontal as in FWS and SSF wetlands (Figure 6.3).

Figure 6.2b

Subsurface flow (SSF) wetland.

Figure 6.3

Vertical flow constructed wetland.

Inlet

Effluent

Porous Media



Polluted water is applied at time intervals over the entire surface of the wetland.The water flows through a permeable medium and is collected at the bottom. Theintermittent application allows the cell to drain completely before the next applica-tion. This type of operation allows for much more oxygen transfer than typical SSFsystems and thus may be a good option for treatment of wastewaters with a relativelyhigh oxygen demand. This type of system has been recommended for removal ofhigh levels of ammonia through nitrification. High BOD levels may cause cloggingdue to biomass buildup; mineral buildup may also cause clogging. Intermittentapplication gives the advantage of greatly increasing the oxygen available for micro-bial reactions, but also greatly increases the mechanical and operational requirementsof the system over the more traditional wetland treatment processes.

6.3 MICROBIAL AND PLANT COMMUNITIES OF A WETLAND

Because of the presence of ample water, wetlands are typically home to a varietyof microbial and plant species. This biological diversity, from the smallest virus tothe largest tree, creates interspecies interactions resulting in greater diversity, morecomplete utilization of energy inflows, and ultimately, emergent properties of thewetland ecosystem.

1,6,7

The treatment wetland system designer should not expect tomaintain a system with just a few known species. The successful wetland designercreates the gross environmental conditions suitable for groups or guilds of species,seeding the wetland with diversity by planting multiple species, using soil seedbanks, and inoculating from other similar wetlands, and then using minimum externalcontrol to guide the wetland development.

1

This form of ecological engineeringresults in lower initial cost, lower operation and maintenance costs, and the mostconsistent system performance.

6.3.1 Bacteria and Fungi

Wetland and aquatic habitats provide suitable environmental conditions for thegrowth and reproduction of microorganisms, two important groups of which arebacteria and fungi. These organisms are important in treatment wetland systemsprimarily because of their role in the assimilation, transformation, and recycling ofchemical constituents present in contaminated waters.

Bacteria and fungi are typically the first organisms to colonize and begin thesequential decomposition of contaminants and wastes. Also, microbes typically havefirst access to dissolved constituents in the wastewater or contaminated groundwater.Some bacteria are sessile, while others are motile by use of flagella. In wetlands,most bacteria are associated with solid surfaces of plants, decaying organic matter,and soils. Bacteria also play a significant role in altering the biogeochemical envi-ronment because they are responsible for processes such as nitrification, denitrifi-cation, sulfate reduction and methanogenesis, etc.

Fungi represent a separate kingdom of eucaryotic organisms and include yeasts,molds, and fleshy fungi. Most fungal nutrition is saprophytic, which means it is



based on the degradation of dead organic matter. Fungi are abundant in wetlandenvironments and play an important role in treatment.

6.3.2 Algae

Algae are unicellular or multicellular photosynthetic bacteria and plants that lackthe variety of tissues and organs of higher plants. Algae are a highly diverse assem-blage of species that can live in a wide range of aquatic and wetland habitats. Majoralgae life forms typical in wetland environments are unicellular, colonial, filamen-tous, and macroscopic forms.

For the most part, algae depend on light for their metabolism and growth andserve as the basis for an autochthonous food chain in wetland habitats. Organiccompounds created by algae photosynthesis contain stored energy which is used formicrobial respiration or which enters the aquatic food chain and provides food to avariety of microbes. Alternately, this reduced carbon may be directly deposited asdetritus to form organic peat sediments in wetlands.

When light and nutrients are plentiful, algae can create massive populations andcontribute significantly to the overall food web and nutrient cycling of a treatmentwetland ecosystem. When shaded by the growth of macrophytes, algae frequentlyplay a less important role in wetland energy flows and treatment (Figure 6.4).

Filamentous algae mats are sometimes a dominant component of the plantbiomass in wetland systems. The mats are made of a few dominant species of greenor blue-green filamentous algae in which individual filaments may include thousandsof cells. Filamentous algae mats first develop below the water surface on the substrateof wetland in areas with little emergent vegetation. During the day, entrained gasbubbles (primarily pure oxygen resulting from photosynthesis) may cause the matsto move up through the water column and float at the surface. During the night, themats sink again to the wetland substrate.

1

Filamentous algae that occur in wetlands as periphyton or mats may dominatethe overall productivity of the wetland, controlling DO and CO

2

concentrationswithin the treatment wetland water column. Wetland water column DO can fluctuatediurnally from near zero during the early morning following a night of high respi-ration to well over saturation (>15 mg/L) in high algae growth areas during a sunnyday. Dissolved carbon dioxide and consequently the pH of the water varies propor-tionally to DO because of the corresponding use of CO

2

by plants during photosyn-thesis and release at night during respiration. As CO

2

is stripped from the watercolumn by algae during the day, pH may rise by 2 to 3 pH units (a 100- to 1000-fold increase in H

+

concentration). These daytime pH changes are reversible, andthe production of CO

2

at night by algae respiration frequently returns the pH to theprevious day’s value by early morning (Figure 6.5.)

1

6.3.3 Species of Vegetation for Treatment Wetland Systems

The term macrophyte includes vascular plants with tissues that are easily visible.Vascular plants differ from algae through their internal organization into tissuesresulting from specialized cells (Figure 6.6). The U.S. Fish and Wildlife Service has



Figure 6.4

Major energy sources and ecological niches affecting the occurrences of algaein wetlands (adapted from Kadlec and Knight, 1996).

Figure 6.5

Typical diurnal plots of DO concentration and pH in a wetland dominated byfilamentous algae.

pH

Dis

solv

ed O

xyge

n (m

g/L)

pH

DO

TIME

12MN

6PM

12NOON

6AM

12MN

10

0

20

10

0



more than 6700 plant species on their list of obligate and facultative wetland plantspecies in the U.S. Obligate wetland plant species are defined as those which arefound exclusively in wetland habitats, while facultative species are those that maybe found in upland or in wetland areas.

1

Wetland macrophytes are the dominant structural component of most wetlandtreatment systems. The vascular macrophytes can be categorized morphologically

Figure 6.6

Growth forms of rooted wetland and aquatic vascular plants (adapted fromKadlec and Knight, 1996).

EmergentHerbaceous

CattailDuck Potato

a.

EmergentWoody

ButtonbushShrub

b.

Floating Leaved

Water Lilly

c.

Submerged

Hydrilla

d.



by descriptors such as woody, herbaceous, annual, perennial, emergent, floating, andsubmerged. Woody species have stems or branches that do not contain chlorophyll.Since these tissues are adapted to survive for more than one year, they are typicallymore durable or

woody

in texture.

Herbaceous

species have aboveground tissuesthat are leafy and filled with chlorophyll-bearing cells that typically survive onlyone growing season. Woody species include shrubs that attain heights of up to sixto ten feet and trees that are generally more than ten feet in height when mature.

1

The terms emergent, floating, and submerged refer to the predominant growthform of a plant species. In

emergent

plant species, most of the aboveground part ofthe plant emerges above the waterline and into the air. These emergent structuresmay be self-supporting or may be supported by other physical structures. Emergentplant species are important because they provide surface area for microbial growthimportant in many of the contaminant assimilation processes in treatment wetlandsystems.

1,2

Floating

species have leaves and stems buoyant enough to float on thewater surface.

Submerged

species have buoyant stems and leaves that fill the nichebetween sediment surface and the top of the water column. Floating and submergedspecies may appear in treatment wetlands when water depths exceed the tolerancerange for rooted, emergent species.

Aquatic macrophyte-based wetlands treatment systems may be classified accordingto the life form of the dominating macrophyte into 1) free-floating macrophyte-basedtreatment systems; 2) rooted emergent macrophyte-based wastewater treatment systems;3) submerged macrophyte-based wastewater treatment systems; and 4) multistage sys-tems consisting of a combination of the above-mentioned concepts and other kinds oflow-technology systems (e.g., oxidation ponds and sanitary filtration systems).

6.3.3.1 Free-Floating Macrophyte-Based Systems

Free-floating macrophytes are highly diverse in form and habit, ranging fromlarge plants with rosettes of aerial and/or floating leaves and well-developed sub-merged roots (e.g., water hyacinth,

Eichhornia crassipes

) to minute surface-floatingplants with few or no roots (e.g., duckweeds

, Lemna, Spirodella, Wolffia

sp.)(Figure 6.7a).

2

Water Hyacinth-Based Systems:

The water hyacinth is one of the most prolificand productive plants in the world. This high productivity is exploited in wetlandtreatment facilities. Two different concepts are applied in water hyacinth-based

Figure 6.7a

Schematic description of a free-floating water hyacinth-(


)based treatment wetland system.

Influent

Effluent

L1282/C06/frame Page 282 Tuesday, June 19, 2001 1:10 PM


wastewater treatment systems: 1) tertiary treatment systems (i.e. nutrient removal) inwhich nitrogen and phosphorus are removed by incorporation into the water hyacinthbiomass, which is harvested frequently to sustain maximum productivity and to removeincorporated nutrients. Nitrogen may also be removed as a consequence of microbialdenitrification; and 2) integrated secondary and tertiary treatment systems (i.e., BODand nutrient removal) in which degradation of organic matter and microbial transfor-mations of nitrogen (nitrification-denitrification) proceed simultaneously in the waterhyacinth ecosystem. Harvesting of water hyacinth biomass is only carried out formaintenance purposes. The latter system should include aerators, that is, areas with afree water surface where oxygen can be transferred to the water from the atmosphereby diffusion and where algal oxygen production can occur. The retention time in thesystems varies according to wastewater characteristics and effluent requirements, butis generally on the order of from 5 to 15 days.

2

The role of water hyacinths in the process of suspended solids removal is welldocumented. Most suspended solids are removed by sedimentation and subsequentdegradation within the basins, although some sludge might accumulate on the sed-iment surface. The dense cover of water hyacinths effectively reduces the effects ofwind mixing and also minimizes thermal mixing. The shading provided by the plantcover restricts algal growth, and hyacinth roots impede the horizontal movement ofparticulate matter. Furthermore, electrical charges associated with hyacinth roots arereported to react with opposite charges on colloidal particles such as suspendedsolids, causing them to adhere to plant roots, where they are removed from thewastewater stream and slowly digested and assimilated by the plant and microor-ganisms. The efficiency of water hyacinths in removing BOD and in providing goodconditions for microbial nitrification is related to their capability of transportingoxygen from the foliage to the rhizosphere. The extensive root system of the waterhyacinth provides a huge surface area for attached microorganisms, thus increasingthe potential for decomposition of organic matter.

1,2

Water hyacinth-based wetland treatment systems are sufficiently developed tobe applied successfully in tropics and subtropics. Water hyacinths are severelyaffected by frost; the growth rate is greatly reduced at temperatures below 10

°

C.Consequently, in temperate regions, water hyacinth-based systems can only be usedin greenhouses or outdoors during summer. Pennywort (

Hydrocotyle umbellate

), onthe other hand, has a high growth rate and a high nutrient uptake capacity evenduring relatively cold periods in subtropical areas.

2

It has been suggested that waterhyacinths and pennywort can be alternately cultured, winter and summer, in orderto maintain performance at a high level year-round.

Duckweed-Based Systems:

Duckweeds (

Lemna, Spirodella

, and

Wolffia

sp.)have not been investigated as much as water hyacinths for use in wetlands treatment.Duckweeds, have a much wider geographic range than water hyacinths, however,as they are able to grow at temperatures as low as 1 to 3

°

C. Compared to waterhyacinths, duckweeds, play a less direct role in the treatment process because theylack extensive root systems and therefore provide a smaller surface area for attachedmicrobial growth.

2

The main use of duckweeds is therefore in recovering nutrientsfrom secondary treated wastewater. A dense cover of duckweed on the surface ofwater inhibits both oxygen entering the water by diffusion and the photosynthetic



production of oxygen by phytoplankton because of poor light penetration. The waterconsequently becomes largely anaerobic, which in turn favors denitrification. Thelight absorption of duckweed cover restricts growth of phytoplankton and thereforethe production of suspended solids.

Duckweed-based systems may be plagued by problems, as high winds can pilethe duckweed into thick mats and eventually completely sweep the plants from thewater. Therefore, in large systems, it is necessary to construct some kind of barrieron the water surface to prevent this. The retention time in duckweed-based wetlandtreatment systems depends on wastewater quality, effluent requirements, harvestingrate, and climate, but it varies typically from 30 days during summer to severalmonths during winter.

6.3.3.2 Emergent Aquatic Macrophyte-Based Systems

Rooted emergent aquatic macrophytes are the dominant life form in wetlandsand marshes, growing within a water table ranging from 18 inches below the soilsurface to a water depth of 60 inches or more. In general, they produce aerial stemsand leaves, and an extensive root and rhizome system. The depth penetration of theroot system, and thereby the exploitation of sediment volume, is different for dif-ferent species. Typical species of emergent aquatic macrophytes are the commonreed (

Phragmites australis

), cattail (

Typha latifolia

), and bulrush (

Scirpus lacustris

).

2

All species are morphologically adapted to growing in a waterlogged sediment byvirtue of large internal air spaces for transportation of oxygen to the roots andrhizomes. Most species of emergent aquatic macrophytes possess an extensive inter-nal lacunal system that may occupy 50 to 70% of the total plant volume. Oxygenis transported through the gas spaces to the roots and rhizomes by diffusion and/orby convective flow of air. Part of the oxygen may leak from the root system intothe surrounding rhizosphere, creating oxidized conditions in the otherwise anoxicsediment and stimulating both decomposition of organic matter and growth ofnitrifying bacteria. Emergent macrophyte-based wetland treatment systems can beconstructed with different designs; see Figure 6.7b for an example. These types ofsystems are also currently applied for the precipitation and removal of dissolvedheavy metals under anaerobic conditions as a sulfide or carbonate precipitate.

Figure 6.7b

Emergent macrophyte treatment wetland system with surface flow (adapted fromMohiri, 1993).



6.3.3.3 Emergent Macrophyte-Based Systems with Horizontal Subsurface Flow

Design typically consists of a bed planted with the common reed

Phragmitesaustralis

and underlain by an impermeable membrane to prevent seepage if required.The medium in the bed may be soil or gravel. During the passage of wastewater orcontaminated groundwater through the rhizosphere of the reeds, organic matter isdecomposed microbiologically, nitrogen may be denitrified, and phosphorus and heavymetals may be fixed in the soil. The reeds have two important functions in the process:1) to supply oxygen to the heterotrophic microorganisms in the rhizosphere, and 2) toincrease and stabilize the hydraulic conductivity of the soil. The quantitative signifi-cance of the uptake of nutrients in the plant tissue is negligible, as the amount ofnutrients taken up during a growing season constitutes only a few percent of the totalcontent introduced with the wastewater. Moreover, nutrients bound in the plant tissueare recycled in the system upon decay of the plant material. Surface runoff is a generalproblem in soil-based treatment facilities because it prevents the wastewater fromcoming into contact with the rhizoshere. Furthermore, the oxygen transport capacityof the reeds seems to be insufficient to ensure aerobic decomposition in the rhizosphereand deliver the oxygen needed for quantitatively significant nitrification.

6.3.3.4 Emergent Macrophyte-Based Systems with Vertical Subsurface Flow

In a vertical flow system the requirements for sufficient hydraulic conductivityin the bed medium and improved rhizosphere oxygenation can be established. Adesign consisting of several beds laid out in parallel with percolation flow andintermittent loading will increase soil oxygenation several-fold compared to hori-zontal subsurface flow systems. During the loading period, air is forced out of thesoil; during the drying period, atmospheric air is drawn into the porespaces of thesoil, thus increasing soil oxygenation. Furthermore, diffusive oxygen transport tothe soil is enhanced during the drying period, as the diffusion of oxygen is approx-imately 10,000 times faster in air than in water. This design and operational regimeprovides alternating oxidizing and reducing conditions in the substrate, therebystimulating sequential nitrification–denitrification and phosphorus adsorption (Fig-ure 6.7c). The limited information available on the treatment performance of suchsystems indicates good performance with respect to suspended solids and aerobicallybiodegradable organics, ammonia, and phosphorus.

6.3.3.5 Submerged Macrophyte-Based Systems

Submerged aquatic macrophytes have their photosynthetic tissue entirely sub-merged (Figure 6.7d). The morphology and ecology of the species vary from small,rosette-type, low-productivity species growing only in oligotrophic waters (e.g.,

Isoetes lacustris

and

Lobelia dortmanna

) to larger eloedid-type, high-productivityspecies growing in eutrophic waters (e.g.,

Elodea canadensis

).



Submerged aquatic plants are able to assimilate nutrients from polluted waters.However, they only grow well in oxygenated water and therefore cannot be used inwastewater with a high content of readily biodegradable organic matter because themicrobial decomposition of the organic matter will create anoxic conditions. Theprime potential use of submerged macrophyte-based wastewater treatment systemsis therefore for “polishing” secondarily treated wastewaters, although good treatmentof primary domestic effluent has been obtained in an

Elodea nuttallii

-based system.The presence of submerged macrophytes depletes dissolved inorganic carbon in thewater and increases the content of dissolved oxygen during the periods of highphotosynthetic activity. This results in increased pH, creating optimal conditions forvolatilization of ammonia and chemical precipitation of phosphorus. High oxygenconcentrations also create favorable conditions for the mineralization of organicmatter in the water. The nutrients assimilated by the macrophytes are largely retainedwithin the rooting tissues of the plants and by the attached microflora. Losses fromthe foliage of plant nutrients upon senescence of the macrophyte tissues are readilytaken up by the periphytic community so that very little leaves the littoral detritusand macrophyte-epiphyte complexes. Much of the detrital matter produced in thesesystems will be accumulated and retained in the sediments.

The use of submerged macrophytes for wastewater treatment is still in theexperimental stage, with species like egeria (

Egeria densa

), elodea (

Elodea canaden-sis

and

Elodea nuttallii

), hornwort (

Ceratophyllum demersum

), and hydrilla

Figure 6.7c

Emergent macrophyte based treatment wetland system with vertical percolation.

Figure 6.7d

Schematic description of a submerged macrophyte-based treatment wetlandsystem.



(

Hydrilla verticillata

) being the most promising. Present knowledge suggests thattheir prime area of application will be as a final step in multistage systems.

6.3.3.6 Multistage Macrophyte-Based Treatment Systems

Different types of these macrophyte-based wastewater treatment systems maybe combined with each other or with conventional treatment technologies. Forexample, a multistage system could consist of 1) a mechanical clarification step forprimary treatment; 2) a floating or emergent macrophyte-based treatment system forsecondary treatment; and 3) a floating, emergent, or submerged macrophyte-basedstep for tertiary treatment. The types of secondary and tertiary treatment step will,among other factors, depend on wastewater characteristics, treatment requirements,climate, and amount of available land.

6.4 TREATMENT WETLAND SOILS

Several individual component wetland processes combine to provide the observedoverall treatments. Sedimentation and filtration remove solids. Chemical precipitation(abiotic or microbially induced), ion exchange, and plant uptake remove metals. Nutri-ents are utilized by plants and algae, and cycled to newly formed sediments.

The definition of hydric soils indicates that any upland soil utilized for construc-tion of a wetland treatment system will become a hydric soil following a short tolong period of flooding and continuous anaerobiosis.

1

Hydric soils are defined as soils that, in their undrained condition, are saturated,flooded, or ponded long enough during the growing season to develop anaerobicconditions favoring the growth and regeneration of hydrophylic vegetation.

8

Since most wetlands are constructed in former uplands, most constructed wet-lands are initially dominated by mineral soils. As constructed wetland treatmentsystems mature, the percent of organic matter in the soil generally increases, and insome systems, soils might eventually cross the arbitrary line between mineral andorganic (Figures 6.8a and b). Mineral soils are classified by particle size distributions,color, depth, and a number of other factors. The three major mineral soil classes areclays, silts, and sands.

Clays are soils with very fine particles packed closely together. Because of theirvery fine texture and low hydraulic conductivity, clays may function as aquitards.The existence of many natural wetlands depends on impermeable clay lenses insedimentary or wind-blown (loess) deposits. Clays typically have the highest adsorp-tion potential of any soils because of their high surface area to volume ratio resultingfrom their small particle size distribution. When water in a wetland is in contactwith underlying clays or when water percolates through the bottom of a clay-linedwetland, the presence of clays may greatly increase treatment potential for conser-vative ions such as phosphorus and metals.

Organic soils, called peat, muck, or mucky peat may be classified by their extentof decomposition. Those soils with the least amount of decomposition (less thanone third decomposed) are called peat. Fibric peats have more than two thirds of



the plant fibers still identifiable. Saprists or mucks have greater than two thirds ofthe original plant materials decomposed. Hemists (mucky peat or peaty mucks) arebetween saprists and fibrists.

Due to their fibrous nature, organic soils may shrink, oxidize, and subside whenthey are drained. Fire may also accelerate this oxidation process, and agricultural

Figure 6.8a

The types of soils present in a newly planted treatment wetland system (adaptedfrom Kadlec, 1996).

Figure 6.8b Types of soil layers developed after a period of maturation in a treatment wetlandsystem (adapted from Kadlec, 1996).

Young Plants

Water

SaturatedMineral Soils

Aerobic

MildlyAnaerobic(Positive REDOX)



practices (drainage, cropping, harrowing, and burning) are known to result in soilsubsidence in highly organic soils such as those in the Everglades agricultural areawhere subsidence rates have been estimated at about 3 cm/yr.

Drying organic soils promotes oxidation and gasifies carbon, but not the mineralnutrients associated with those soils. Although the available nutrient content of apeat or muck is often quite low, there are large amounts of nitrogen, phosphorus,sulfur, and other mineral constituents organically bound in unavailable forms. Oxi-dation destroys these recalcitrant organics and releases the associated substances.Upon reflooding, those substances can dissolve and provide relatively high concen-trations of nutrients and other dissolved minerals.

Organic soils cannot easily be characterized by grain size because the necessaryact of drying destroys the physical-chemical structure. The general range of hydraulicconductivity for soils found in sedge, reed, and alder wetlands is 0.1 to 10 m/d,placing these materials in the range of other mineral soils. However, this is true onlyfor fully saturated soils; even a slight degree of unsaturation lowers the hydraulicconductivity by two orders of magnitude, due to the extremely large capillary suctionpressure created in the micropores. This means that organic soils and sediments arevirtually undrainable; they retain a very high percentage of water. Organic soils aretypically dark in color, ranging from black mucks to brown peats.

Soil chemical properties are primarily related to chemical reactivity of soilparticles and the surface area available for chemical reactions. Chemical reactivityis related to the surface electrical charge of the soil particles, is typically highest inclays and organic soil particles.

6.4.1 Cation Exchange Capacity

Wetland soils have a high trapping efficiency for a variety of chemical constit-uents; they are retained within the hydrated soil matrix by forces ranging fromchemical bonding to physical dissolution within the water of hydration. The com-bined phenomena are referred to as sorption.

A significant portion of chemical binding is cation exchange, which is replace-ment of one positively charged ion, attached to the soil or sediment, with anotherpositively charged ion. The humics substances found in wetlands contain largenumbers of hydroxyl and carboxylic functional groups, which are hydrophilic andserve as cation binding sites. Other portions of these molecules are nonpolar andhydrophobic in character. The result is the formation of micelles, groups of humicmolecules with their nonpolar sections combined in the center and their negativelycharged polar portions exposed on the surface of the micelle. Protons or otherpositively charged ions may then associate with these negatively charged sites tocreate electrical neutrality.

Micelles are one form of ligand that can bind metal ions. The cumulative processof binding a metal ion to a ligand (L) to form a complex may be described by achemical equation; here, it is illustrated for the binding of a divalent metal ion (M):1

2HL + M2+ ⇔ ML2 + 2H+ (6.1)



The number of ligands per gram of dry solid is determined from the number ofmetal ions that can be sorbed by a fully protonated sample. This is referred to asthe cation exchange capacity (CEC) of the material, usually measured in milliequiv-alents per gram. Peats have CEC values of approximately 1.0 to 1.5 meq/g. For aheavy metal such as copper, this can translate to a large binding capacity, on theorder of a few percent by weight.

Clearly, the pH of the soil or sediment has a large influence on the partitioningof a metal to the ligand because excess hydrogen ions drive Equation 6.1 towardthe ionic form of the metal.

Drying of the organic material will destroy some of the character of the highlyhydrated micellular chemical-physical structures, therefore destroying some of thesorption capacity of the material. The sorption capacity of dried, harvested peats isnot as large as that of wet, living peats.

6.4.2 Oxidation and Reduction Reactions

Wetlands are ideal environments for chemical transformations because of therange of oxidation states that naturally occur in wetland soils. Free oxygen decreasesrapidly with depth in most flooded soils because of the metabolism of microbes thatconsume organic matter in the soil and through chemical oxidation of reducedsubstances. This decline in free oxygen, in other words the depletion of oxidizingpotential, is measured as an increasingly negative electric potential between a stan-dard platinum electrode and the concentration of oxygen in the soil. This measureof electric potential is called reduction-oxidation or REDOX potential (ORP) andprovides an estimate of soil oxidation or reduction potential (Figure 6.9).

Figure 6.9 Typical depth profile for potential oxidation-reduction reactions taking place in atreatment wetland system.



When ORP >100 mV, conditions are termed aerobic because dissolved oxygenis available. When ORP <–100 mV, conditions are termed anaerobic because thereis no dissolved oxygen. Some authors refer to intermediate conditions (near-zeroDO) as transient or anoxic.

The reduction of ferric iron (Fe3+) to ferrous iron (Fe2+) is a REDOX half reactiontypical of anaerobic sediments:

Fe3+ + e– = Fe2+ (6.2)

The typical complementary half reaction for this reduction of ferric to ferrousiron is the oxidation of reduced sulfur (hydrogen sulfide, H2S) to sulfate (SO4

2– ) bythe half reaction:

H2S + 4H2O = SO42– + 10H+ + 8e– (6.3)

for an overall oxidation-reduction reaction of:

8Fe3+ + H2S + 4H2O = 8Fe2+ + SO42– + 10H+ (6.4)

However, the real situation in treatment wetlands is far more complicated becausethere are more species present and more REDOX routes than those used here forillustration. For instance, iron reduction can be accompanied by sulfate reduction— not sulfate production — and the ferrous ion, Fe (II), will be precipitated outas FeS.

The REDOX potential of many wetland soils decreases with vertical depth intothe sediments because the only source of free oxygen is from atmospheric diffusionat the top of the sediment layer (Figure 6.9). The typical oxygen gradient in wetlandsediments includes a thin (less than a few centimeters) oxidized surface horizon atthe sediment-water interface, underlain by increasingly reduced conditions withdepth based on the amount of biological and chemical reducing activity in thesediments. Vertical REDOX gradients in treatment wetland soils will vary in responseto distance from the point of wastewater loading.

Northern wetlands are sometimes sealed by an ice cap in winter that preventssupply of oxygen to the water and/or soils. This then shifts the REDOX profile tomuch lower Eh values, causing sulfate reduction and methanogenesis to dominateeven the upper soil horizons. Gaseous sulfur compounds cannot escape and mayreach lethal levels for wetland biota.

Treatment wetlands are often subjected to waters with higher oxygen demandsexerted by both carbonaceous and nitrogenous compounds. This causes a greaterdepletion of electron acceptors such as oxygen, nitrate, sulfate, and iron in both thewater column and the underlying soils. The REDOX potential in treatment wetlandsis therefore typically lower than for natural wetlands, ranging from the denitrifyingregime downward to the methanogenesis regime.



6.4.3 pH

Prior to flooding, soils may have widely varying pH conditions ranging fromabout 3 to 10 units. Following flooding, pH in wetland soils may initially declinedue to aerobic decomposition liberating carbon dioxide into the interstitial water.However, this initial pH swing is generally transient and is followed by a typicaltrend in both acidic and alkaline soils toward pH neutrality (pH 6.7 to 7.2 units)over time. This typical trend is considered to be the result of ferric iron reductionunder flooded soil conditions. In some highly organic soils, pH may remain verylow, even following long periods of flooding. This result is likely due to the slowoxidation of organic sulfur compounds resulting in production of sulfuric acid andto the presence of humic acids.

6.4.4 Biological Influences on Hydric Soils

Biological, chemical, and physical properties of wetland soils are interdependent.Microbial wetland fauna make up a significant fraction of the organic carbon occur-ring in hydric soils. These tiny organisms are competing for sometimes limited andrapidly shifting supplies of energy containing compounds and nutrients, and theirgrowth and death have a very significant effect on the fate and transport of themajority of soil chemical constituents. In addition to the microbial populations,macrophytic plants diversify soil structure through the growth and death of rootsand creation of decaying plant litter, and wetland animals dig, burrow, scrape, andotherwise cause bioturbation of wetland sediments on an almost continuous basis.Some of the major interactions between wetlands biology and sediments aredescribed below.

6.4.5 Microbial Soil Processes

Soil microbial populations have significant influence on the chemistry of mostwetland soils. Important transformations of nitrogen, iron, sulfur, and carbon resultfrom microbial processes. These microbial processes are typically affected by theconcentrations of reactants as well as the REDOX potential and pH of the soil.

Organic nitrogen is biologically transformed to ammonia nitrogen through theprocess of mineralization, which results as a consequence of organic matter decom-position that results from actions of both aerobic and anaerobic microbes. Ammoniais in turn converted to nitrite and nitrate nitrogen through an aerobic microbialprocess called nitrification. Nitrate nitrogen can be further transformed to nitrousoxide or nitrogen gas in anaerobic wetland soils by the action of another group ofmicrobes (denitrifiers). Nitrogen gas can also be transformed to organic nitrogen bybacterial nitrogen fixation in some aerobic and some anaerobic wetland soils.

Bacteria can transform reduced iron and possibly manganese to oxidized forms.These chemosynthetic processes utilize oxygen as an electron acceptor and usuallyare accelerated by acidic conditions typical of acidic coal mine drainage waters.

Sulfate can be reduced to sulfide by anaerobic bacteria in wetlands. Thesulfate serves as an electron acceptor in the absence of free oxygen at low REDOX



potentials. Sulfides can provide a source of energy for chemautotrophic and photo-synthetic bacteria in aerobic wetlands, resulting in the formation of elemental sulfurand sulfate. The formation of sulfide under anaerobic conditions enables the precip-itation of dissolved heavy metals in treatment wetlands.

Organic carbon is microbially degraded to carbon dioxide by aerobic respirationwhen oxygen is available and by fermentation under anaerobic conditions. Greaterenergy is released under aerobic respiration, resulting in more efficient assimilationof organic matter into microbial cellular material. In fermentation, organic matterserves as the terminal electron acceptor, forming acids and alcohols. Methane canbe formed in wetlands due to the action of bacteria using carbon dioxide as anelectron acceptor at very low REDOX potentials.

6.4.6 Treatment Wetland Soils

The sediments that form in treatment wetlands are often different from thosethat form in natural wetlands, for a number of reasons. First, the enhanced activityof various microbes, fungi, algae, and soft-bodied invertebrates leads to a greaterproportion of fine detritus compared to leaf, root, and stem fragments. There issignificant formation of low-density biosolids (sludge). Second, there may be aprecipitation of metal hydroxides, carbonates, or sulfides, which add mineral flocsto the sediments. Finally, there is often a high ionic strength associated with treatmentof effluents reflected in high dissolved salt content. The effect of high ionic strengthis to alter the structure of the highly hydrated organic materials comprising wetlandsediments and soils.

The considerations listed previously show that wetland soils are a vital andintegral link in processes that govern water quality. Therefore, it seems reasonableto expect consideration of wetland soils to be an important part of treatment wetlanddesign; however, that is not the case. Antecedent soils are altered and replaced bynew organic soils. The sorption capacity of the antecedent soils is re-equilibratedwith the new water quality of the incoming water, perhaps along a gradient frominlet to outlet. If there are leachable chemicals, they are depleted and exit the wetland.Roots and rhizomes of new plantings (constructed wetlands) or replacement species(natural wetlands) repopulate the top 30 cm of the wetland and set up a newbiogeochemical cycle.

The long hydroperiods of treatment wetlands are conducive to the buildup oforganics: first litter and microdetritus, then the sediments formed from their decom-position, and, finally, the organic soils generated from those sediments and depositedmineral solids.

In short, the wetland rearranges itself to accommodate the environment createdby the designer. The functioning of the wetland after such adaptation is no longerdependent upon the previous condition and type of soils, hydrology, and biota; it isnow totally dependent upon the new soils, hydrology, and biota. It is this newsustainable mode of wetland operation that is the target of most designs.

Available data indicate that the final state of a treatment wetland, and the accom-panying suite of water quality functions, are largely independent of the initialcondition of the real estate upon which it is built. During the interim period of



adaptation, antecedent conditions are important because they dictate the short-termperformance of the wetland. That period of adaptation appears to extend for up to2 years for newly constructed wetlands and longer for alteration of natural wetlandsto a treatment function.1,2

If soils are unsatisfactory for plant or microbial growth, the wetland treatmentsystem is liable to have inadequate plant cover. A knowledge of the physical andchemical composition of site soils is essential to predict accurately some internalchemical and biological processes in treatment wetlands. The rate of soil accretionin wetland treatment systems affects the potential removal of conservative elementssuch as phosphorus and heavy metals and also is an important consideration duringdesign of berm height above the wetland substrate. Some sorbed materials can bereleased if exposed to lower concentrations in incoming water. Violent hydrologicevents are capable of resuspending particulate deposits. It is therefore incumbent onthe designer to minimize or prevent such occurrences.

The main environmental factor that influences the nature of wetlands soils isdissolved oxygen concentration. Vertical oxygen gradients are typically establishedin wetland soils due to bacterial respiration and chemical oxidation demand and dueto the greatly reduced rate of oxygen diffusion in saturated soils compared tounsaturated soils. These oxidation gradients result in a chain of oxidation-reductionreactions that provide many wetlands with their typical profile of declining REDOXpotentials with depth (Figures 6.10a and b). REDOX, in turn, affects the microbialprocesses important in most aspects of wetland use for water quality improvement,especially including removal of organic carbon and nitrogen.

Wetland soils are as dynamic in character as all other aspects of the wetlandecosystem. Soils in constructed wetlands built on upland sites undergo gradualtransformation, resulting in accumulations of organic carbon and reduced elementssuch as iron and sulfur typical of natural wetland soils. Many of the changes thatoccur during wetland development and succession are the result of biological factorsthat occur in wetlands such as growth of bacteria and fungi, algae and macrophyticplants, micro- and macroinvertebrates, and larger animals. While many of thesenatural biological processes are not within the control of the wetland treatmentsystem designer and operator, their effects should be considered when trying tomaximize chances for success with a wetland project.

6.5 CONTAMINANT REMOVAL MECHANISMS

6.5.1 Volatilization

A shallow wetland water body provides the opportunity for air stripping ofvolatiles. The efficiency is not as great as for mechanical devices, but the differenceis more than compensated for by the long detention times and large surface areasin the wetland. Half-lives (time to volatilize one half the substance) range from 2–4days for compounds such as benzene, toluene, and naphthalene, and are projectedto be less for more volatile compounds, such as vinyl chloride and chloromethane.9,10



More soluble, less volatile organics are less likely to volatilize, but then have anopportunity for enhanced biodegradation to occur. As a result, half-lives for sub-stances such as phenol, tetrahydrofuran, and methanol are also fairly short; thesehave been measured to be in the range of 10 to 40 hours in wetlands.11

The half-lives of low molecular weight alkanes (<C8) and mono- and bicyclicaromatics (including PCBs) have been projected to be less than 12 hours at one-meter water depth.9 Therefore, lighter weight molecules are very likely to be effec-tively stripped in wetlands designed to remove other constituents with equal or lowerrate constants. Experimental studies verified the strong effect of mixing in the waterphase, and established a diffusion-only half-life of about 8 hours for benzene,toluene, TCE, and PCE.12

6.5.2 Partitioning and Storage

Many organics are known to sorb strongly to the organic sediments and substratesin wetlands. In addition, they may be complexed with dissolved organic matter

Figure 6.10a Idealized macrophyte root structure in a wetland soil illustrating development ofan oxidized rhizophere (adapted from Kadlec and Knight, 1996).



(DOM). For example, the partition coefficient of hexachlorobenzene (HCB) to wet-land sediments has been investigated and it has been determined that the associationof dosed HCB with DOM is a fast reaction that equilibrates in one hour.

9

The role of wetland vegetation in buffering air emissions from waste sitesand the role of partitioning between the plant–air interface and resulting impactcycling of contaminants have also been investigated.

13

The impact of the qualityof organic matter on the magnitude of desorption–resistance has been studied.Most studies indicate the presence of a sizeable desorption–resistant-phase inwetland soils and an increase in the size of the desorption–resistance in “older”organic matter.

Volatile substances are gasified. Many materials undergo microbial transforma-tions. These processes all lead to the transformation and transfer of a “removed”pollutant either to the atmosphere or to the wetland sediments and soils. The vege-tation is extremely important for nutrient transformations and transfers because itplays a key role in the cycling and temporary storage of many substances.

Figure 6.10b

Pathways of nitrogen transformations in the immediate vicinity of a wetland plantroot (adapted from Reddy and Graetz, 1988).

Root

Organic C

CH4

Root

Aerobic

4CH

2N

2O

Organic N

Anaerobic

4NH +NO -3

2N

3NO

NH +4



6.5.3 Hydraulic Retention Time

Removals proceed over the time water is held in the wetland. Therefore, detentiontime (or the equivalent hydraulic loading rate) becomes the key design variable. Thelonger the water is held in the wetland, the better the treatment — so long as it isnot at the expense of added depth, which contributes little to the active transfers andstorages. However, wetlands possess irreducible background concentrations of somesubstances: about 5 mg/L of BOD and TSS, for instance. Typical detention timesrange from one to ten days for existing treatment wetlands, which corresponds tohydraulic loading rates (rainfall equivalent) of 1–10 cm/day.

This technology requires land instead of mechanical devices to accomplishtreatment. If the necessary land is available, it typically offers capital savings overcompetitive processes. The passive nature of wetlands technology typically offers avery large advantage in operating costs because operation is simple and maintenanceis very low.

It has been assumed by some people that wetlands are easily conceived and builtand are low technology devices; therefore, they should be easily described in termsof simple equations. In fact, they are exceedingly complex “ecoreactors” that requirecomplex descriptions if they are to meet designed expectations. A design modelmust first do an adequate job of predicting wetland hydraulilcs. The basic tool isthe interior water budget. Stream inflows and outflows are typically measurable andcontrollable in design. The wetland often will not communicate with groundwaterdue to existing or constructed clay seals or liners. The uncontrollable elements ofthe water budget are precipitation and evapotranspiration.

The fundamental hydraulics question for treatment wetlands is conveyancecapacity and the related issues of depths and slopes. These issues are answered bythe interior water budget, together with data on the flow resistance of the wetland.FSF and SSF wetlands are a bit different with respect to frictional characteristicsand should be treated separately. The hydraulic sizing of the wetland is then set sothat operation will occur within the desired parameters of depth and flow.1

Other tools necessary to determine the size of the treatment wetland are theparameters required to achieve the biogeochemical and physical processes for con-taminant mass reduction. These are a description of internal flow paths, the interiorchemical mass balance, and the reaction rate constants. Detailed description of thedesigned process of a treatment wetland system is out of the scope of this chapterand book, but a few lessons from expert practitioners are described below.1,2

A pictorial description of the energy balance in a wetland system is shown inFigure 6.11. In addition, Figure 6.12 portrays mass balance objectives; meeting thisobjective is the primary function of a treatment wetland system. For the sake ofbrevity, only parameters of importance for the design are listed below:1

• Evapotranspiration, water losses, and gains• Area requirements, number of cells, and cell shape• Overland flow and geohydrology



• Water depth, subsurface wetland hydraulics, bed friction, hydraulic conductivityand changes with deposition, and clogging

• Surface water elevation profiles, dynamic responses with precipitation, shadowand dead zones

• Nonideal flow patterns, vertical and transverse• Mixing flow velocity and residence time• Biochemical reaction rates, residence time variations based on flow paths, substrate

additions to achieve desired treatment reactions, plug flow vs. CSTR reactor models• Mass transfer mechanisms• Temperature, dissolved oxygen, and pH profiles• Suspended solids

Figure 6.11 Components of the wetland energy balance (adapted from Kadlec and Knight,1996).

Figure 6.12 Components of a chemical budget and associated terminology (adapted fromKadlec and Knight, 1996).

IncomingSunlight

(Short Wave) Clouds Absorband Reflect

Evapotranspiration

RadiativeHeat Loss

(Long Wave)

Outgoing WaterHeat Content

Conductive Transferto/from Ground

Incoming WaterHeat Content

Convective Transfer

to/from Air

Surface Reflects

TranspirationEvaporation

Change in Storage

Surface Area = A

Volumetric Outflow = QConcentration = CMass Outflow = Q C

Mass Removal

Volumetric Flow = QConcentration = CMass Inflow = Q C

ii

i i

eee e



• Nutrients demand and cycling• Flood protection• Earth moving, dikes, berms, and levees• Plant selection and planting, which includes factors such as geographic location,

physiographic features, regional hydrology, climatic conditions, soil characteris-tics, water treatment objectives, type and intensity of adjacent land use impacts,wildlife, and aesthetic preferences

6.6 TREATMENT WETLANDS FOR GROUNDWATER REMEDIATION

Utilization of wetlands for the purpose of remediating contaminated groundwa-ters is a relatively new concept. The evolution of ideas has been from an independentpoint of view, not based upon other pre-existing aspects of the technology. Con-structed wetlands in this context could be a subset of options under the generalheading of phytoremediation in the previous chapter. Several processes are envi-sioned as effective in pollutant reduction: phytoextraction, phytostabilization, tran-spiration, stripping, and rhizofiltration.

Phytoextraction refers to plant uptake of toxicants, which is known to occur andhas been studied in the stormwater and mine water treatment wetland context.However, in many cases the contaminant is selectively bound up in below groundtissues, roots, and rhizomes, and is not readily harvested. Phytostabilization refersto the use of plants as a physical means of holding soils and treated matrices inplace. This process is also one of the chief underpinnings of treatment wetlands asit relates to sediment trapping and erosion prevention in those systems. Wetlandplants possess the ability to transfer significant quantities of gases to and from theirroot zone and the atmosphere (Figure 6.10). This ability is part of their adaptationrequired for survival in flooded environments. Stems and leaves of wetland plantscontain airways (aerenchyma) that transport oxygen to the roots, and vent watervapor, methane, and carbon dioxide to the atmosphere. There may also be transportof other gaseous constituents, such as dinitrogen and nitrogen oxides, and volatilehydrocarbons. The dominant gas outflow is water vapor, creating a transpiration fluxupward through the plant.

Rhizofiltration refers to a set of processes that occur in the root zone, resultingin the transformation and immobilization of some contaminants. Wetlands are ver-tically stratified with respect to REDOX potential and other important chemicalattributes (Figure 6.9). These vertical gradients provide for interzonal processes withboth oxic and anoxic character. The microenvironments in the root zone of wetlandplants also have a great deal of structure, leading to extremely different biogeochem-ical conditions in zones in close proximity. Typically, oxidized conditions are foundadjacent to roots, while anaerobic conditions may exist only millimeters away. Thisvaried environment makes possible diffusional transfers between chemically differ-ent regions. A prime example is the precipitation of dissolved heavy metal ions assulfides in the anaerobic zones of wetlands.

Despite this wealth of scientific knowledge about individual contaminant removalprocesses, there have been few attempts to implement wetlands for groundwater



remediation. Therefore, the potential of the technology must currently be assessedfrom other applications. The next sections briefly address the state of knowledge inclosely related applications.

6.6.1 Metals-Laden Water Treatment

A very large application area for constructed wetlands is the treatment of acidcoal mine drainage. Hundreds of wetlands are now in operation serving this function.The contaminants of interest are typically pH, iron, and manganese. Despite thelarge number of such wetlands, there is not yet a clearly stated design methodologyavailable for acid mine drainage treatment.

Individual metals have been targeted at mining sites that involve them. Forexample, a good deal is known about wetland treatment of copper,14 aluminum,15

and zinc. There are other reports on removal efficiencies of a number of wetlandsreceiving a variety of mine effluents. Moderately high efficiencies, usually in the60–90% range are reported from full and pilot scale projects.16,17

Many treatment wetland field studies have investigated removal efficiencies formultimetal wastewaters, mostly at low to moderate concentrations in domestic/indus-trial combinations and urban stormwaters.18-21 Moderately high efficiencies, usually inthe 60–90% range are also reported in these studies. Laboratory and mesocosm scalestudies bear out these results under more controlled, but less realistic conditions; forexample, fast and high reductions in copper and chromium (VI) in mesocosms.11

Metals in wastewater and contaminated groundwater must be removed prior tofinal discharge to protect the environment form toxic effects, but the use of wetlandsto accomplish this goal must be examined cautiously. The potential for economicaltreatment is nevertheless quite attractive, and the concept of metal removal inwetlands has undergone considerable investigation.

Metals are removed by cation exchange to wetland sediments, precipitation assulfides, carbonates, and other insoluble salts, and plant uptake. The anaerobicsediments provide sulfate reduction to sulfide and facilitate chemical precipitation.As a result, good removals of metals are reported for operating wetland facilities.Removal of heavy metals such as Ni, Nz, Cu, Fe, Mn, Co, and other metals hasbeen reported at varying removal efficiencies. Under optimum operating conditions,some of these metals have been removed at efficiencies of up to 96–98%. At mostof these treatment wetland systems designed for heavy metal removal, a majorityof the mass reduction has taken place at deeper depth, indicating that anaerobicconditions are a must for the precipitation mechanisms. Researchers have shownthat spent mushroom or other organic substrates enhance the formation of anaerobicconditions in both SSF and FWS systems.1,2

A pretreatment system with limestone and peat mixture beds has been used ata few treatment wetland systems for the treatment of acid mine drainage or coalstockpile drainage waters. Limestone is used to neutralize the pH of the inputdrainage water and provide for increased alkalinity and precipitations as metalhydroxides. It has been shown that these drainage waters (with a pH of 2–3) enterthe actual treatment zone with a pH greater than 6.0 from the limestone pretreatmentbeds (Figures 6.13 and 6.14).



The peat mixture beds are placed downstream of the limestone, with the flowdispersed laterally to increase the contact area and encourage subsurface flow (SSF).The subsurface flow through the peat mixture encourages the development of anaer-obic conditions for sulfate reduction activities. In addition, the dynamics of thephysical/chemical and biological components of peat/wetland systems utilize othermetal removal mechanisms such as adsorption, cation exchange, and complexationof soil organic matter (SOM). Some practitioners have added spent mushrooms withpeat to increase the SOM. Removal efficiencies of around 95–98% have beenreported for similar systems. It is important to pay attention to the selection of the

Figure 6.13 Typical peat/wetland treatment system, designed for heavy metal removal fromstockpile drainage (adapted from Moshiri, 1993).

Figure 6.14 General schematic of staged aerobic constructed wetlands.

Limestone Bed

Flow Control Culvert

Peat Mixture

Open Water Pool

Open Water Pool

Open Water Pool

Open Water Pool

Water Flow Direction

Limestone Bed

Peat Mixture Treatment

Intermittent Creek

Present Extent of Wetland

Control Culvert

Berm

LimestoneDrain

DeepPond Deep

Marsh

ShallowMarsh Rock

FilterAlkaline Polishing

CellDischarge

Stormwater Control ChemicalsTSS RemovalCompliance

pH IncreaseMn RemovalMn RemovalFe RemovalFe Removal

Mn CoprecipitationTSS Removal

+Alkalinity

Aeration



right type of peat for these systems. The most important factors are high hydraulicconductivity and available organic content.

Similar removals are obtained for other metals; for instance, chromium is reducedby more than 70% in about 70 hours in surface flow wetlands.1,2 The area designapproach selects the wetland area required to remove a given amount of metal.Recommended areas for iron and manganese reduction are in the range of 100–500square meters per kilogram of metal removed per day. Longer detention timesproduce even greater reductions. Less is known about some elements and com-pounds, such as boron, arsenic, cyanide, and selenium. However, enough is knownthat the correct conditions for removal may be intentionally designed into thewetland.22

In summary, metal removal in wetlands stems from a variety of biogeochemicalprocesses, including aerobic and anaerobic processes in the water column, on thesurface of living and decaying plants, and in sediments. The most significant mech-anisms include:

• Sorption and/or exchange onto organic matter (detritus)• Filtration of solids and colloids• Formation of insoluble metal sulfides• Formation of carbonates• Association with iron and manganese oxides• Metal hydrolysis (catalyzed by bacteria under acidic conditions)• Reduction to nonsoluble, nonmobile forms (also catalyzed by bacteria)

These processes are involved to various degrees depending on specific circum-stances, and it is critical to identify their relative importance in specific treatmentwetlands. Identifying and quantifying these processes provides a basis for the rationaldesign of treatment wetlands and informs on the stability and biological availabilityof the contaminants they retain.

6.6.1.1 A Case Study for Metals Removal*

The site is a former fibers manufacturing site located in the coastal plain ofVirginia. A technical solution was needed to reduce zinc loadings to meet NPDESor Publicly Owned Treatment Works (POTW) zinc limits for discharge from theimpacted area. After a feasibility study and pilot test, a constructed treatment wet-lands was retrofitted into an impoundment area holding zinc containing sediments.The area also receives a discharge of zinc and iron laden landfill leachate from anadjacent industrial 36-acre landfill. A passive treatment system was preferred to amechanical treatment system as a cost-effective and long-term solution to rehabilitatethe impacted area and provide treatment of the water in the impounded area.

At the start of the restoration project, the wetlands contained no vegetation as aresult of the acidic water. Early efforts included pH adjustment by adding limestonegravel and slurry to raise pH. Attempts were made to replant the area with several

* Courtesy of Joseph McKeon, BASF Corporation, Remediation Manager, Ecology and SafetyDepartment.



species to determine which would survive. Phragmites australis volunteered andpopulated the entire wetlands, crowding out most of the replanted species. ThePhragmites community continues to thrive and has spread to areas with hospitablemoisture conditions, as well as repopulating areas where vegetation was removedfor construction.

The wetlands was designed as a two stage aerobic/anaerobic treatment systemto remove the iron in an oxic environment and the zinc as sulfide in an anaerobicenvironment. A partitioning dike was built across the wetlands to separate the aerobicone third of the 16.2-acre wetlands. The characteristics of the leachate coming fromthe adjacent landfill approximates high strength acid mine drainage with an acidicpH and high iron concentrations. The leachate is passively intercepted in a series ofanoxic limestone drains (ALDs), which are mitigation chambers installed below thewater table. Each ALD contains a specified volume of high calcium carbonatelimestone that dissolves upon contact with the acidic leachate, thus raising the pHand increasing the alkalinity. At the point of ALD discharge into the wetlands, thepH is about 7.0 and contains both iron (II) and zinc. The iron oxidizes to Iron (III)Oxide and precipitates to the substrate of the wetlands. Some zinc coprecipitateswith the iron and some is transported in solution to the anaerobic treatment com-ponent of the wetlands. Sulfate naturally occurring in the leachate and supplemen-tally added in the compost cell is reduced to hydrogen sulfide and forms a relativelyinsoluble and nonbiologically available precipitate of zinc sulfide.

Spent mushroom compost is used to drive the wetlands anaerobic at the aero-bic/anaerobic boundary and to add alkalinity and sulfate to the water. One of theprinciple components of the compost is manure which serves as a carbon (food)source for the biologically mediated reduction process. Limestone in the compostadds alkalinity, providing additional buffering capacity to stabilize pH. Other nutri-ents in the compost contribute to the health of the vegetation in the wetlands. Acompost cell is also installed in the area receiving the largest amount of leachateand zinc loadings to further condition the water before it enters the anaerobictreatment area.

Attenuation of zinc is dependent on the reduction of sulfate to produce hydrogensulfide. The rhizosphere of the Phragmites provides the substrate to support thesulfate reducing bacteria. Phragmites also transpires large volumes of water duringthe warm months to control water levels in the impounded wetlands area. Thisspecies of plant is not an important part of the food chain so uptake of metals is notan issue for managing the passive treatment system. Phragmites does provide habitatfor insects, reptiles, and amphibians and is used by deer as a bedding area.

Zinc concentrations as high as 400 mg/l are introduced into the wetlands in thelandfill leachate and are reduced to as low as 0.3 mg/l enabling discharge to thelocal POTW without further treatment. Treatment efficiency is higher in the warmermonths when biological activity is at maximum and lower during the cold months.Water level management helps to minimize discharge during the colder months, thusenabling more contact time with the treatment area and thus assisting zinc attenua-tion. The treatment system was designed with excess treatment capacity to manageother zinc containing streams, if necessary. Various sections of the treatment wetlandsystem is shown in Figures 6.15a-d.



Figure 6.15a

Headwaters of the wetlands. (Photo courtesy of Joseph McKeon.)

Figure 6.15b

The aerobic or oxic area. (Photo courtesy of Joseph McKeon.)



Figure 6.15c

Spent mushroom compost at the beginning of the anoxic or anaerobic area. (Photo courtesy of Joseph McKeon.)

Figure 6.15d

Vegetation in the anaerobic area. (Photo courtesy of Joseph McKeon.)



6.6.2 Removal of Toxic Organics

Organic contaminants partition strongly to solid organic substrates. Wetlandsediments are therefore excellent sinks for organics. Subsequent to such partitioning,the organic chemical may diffuse downward, or undergo biodegradation. The wet-land environment is complicated by the existence of biofilms on submerged plantparts and litter. Although small in terms of mass per unit volume, these biofilms arevery active in biodegradation, and consequently serve as important sinks for organ-ics.23 In this aspect, wetlands resemble conventional attached growth bio-treatmentprocesses. The number of toxic organics that can potentially create environmentalproblems is very large. Not all of these have been studied with respect to wetlandtechnology. Those that have been investigated show removals due to volatilization,storage, and biological degradation in the wetland environment.

6.6.2.1 Biodegradation

Anthropogenic organic chemical additions to wetlands for treatments are mod-ifications to an exceedingly complex background array of hydrocarbons and organicchemical reactions. Hydrocarbon molecules are susceptible to fragmentation andchemical conversion in the wetland environment, predominantly via microbiallymediated pathways. Partial conversion may occur via hydrolysis, dealkylation andring cleavage, or the removal of amino, nitro, chlorine, hydroxyl, acid, or thio groupsfrom the parent molecule. Oxidative processes ultimately produce carbon dioxideand water, while anaerobic processes will enhance reductive dechlorination and mayterminally result in the evolution of methane. It is therefore important to identifythe byproducts of degradation.

Trichloroethene (TCE) and 1,1,2,2-Tetrachloroethane (PCA), at concentrationranges of 100 to 1000 ppb, were naturally attenuated when contaminated ground-water from an aerobic, sandy aquifer discharged through anaerobic sediments in afreshwater wetland at a U.S. army base in Maryland.24 Microcosm experimentsconfirmed field observations that Cis 1,2-DCE, and VC are the dominant daughterproducts from anaerobic biodegradation of both TCE and PCA in the wetlandsediments. Cis 1,2-DCE was produced by reductive dechlorination of TCE anddihaloelimination of PCA. VC was produced by reductive dechlorination of Cis 1,2-DCE and dihalaelimination of the 1,1,2-TCA that was produced through reductivedechlorination of PCA. Parent and daughter concentrations in the microcosmsdecreased to less than 5 ppb in less than 35 days, showing extremely rapid degra-dation rates in these organic rich sediments.

Wetlands have all the basic elements needed for the attenuation of chlorinatedalkenes and alkanes, which include high organic carbon content in the sediments tobind the contaminants; high microbial density and diversity in the sediments tobiodegrade contaminants; and both anaerobic and aerobic conditions to ensure thatcontaminants can be fully degraded without the accumulation of potentially toxicintermediates such as VC. Treatment wetlands can be designed and constructedspecifically to enhance the natural physical and biochemical processes to degradechlorinated organic compounds (Figure 6.16). Some basic steps to be taken into



consideration on such a project include site suitability assessment, risk management,modeling and conceptual design;,permitting, design, construction, and operation andmaintenance.

Though anaerobic environments, such as high organic carbon wetland sediments,are often sufficiently reducing to make reductive dechlorination reactions thermo-dynamically favorable, the transfer of electrons from reduced species to a chlorinatedsolvent is often kinetically constrained. The reduction rate may be enhanced in thepresence of compounds capable of facilitating transfer of electrons from the bulkreductant to the chlorinated solvent of interest. Nickel and copper complexes withAldrich humic acid have recently been shown to be effective electron mediators forthe reductive dechlorination of TCE.25 Porewaters from the high organic carbonsediment zone of the West Branch Canal Creek Wetland at Aberdeen ProvingGrounds, Maryland, were used. Using titanium (III) citrate as the bulk reductant,Ni and Cu complexes (of DOC) present at the site sediments rapidly reduced TCE.The reaction was pseudo-first order with a half-life of approximately one hour forboth DOC-metal complexes. Reaction rates were comparable to the Aldrich humicacid-transition metal systems. Dechlorination was complete with the dominant prod-ucts being ethene and ethane, the mass balance was near 100%, and chlorinatedintermediates were either absent or at extremely low concentrations.

The partitioning of compounds from the aqueous phase into biota is not the onlysignificant process that occurs after the initial discharge of xenobiotics into wetlandsystems. Partitioning of xenobiotics from the aqueous phase into the sediment phasemay be of equal significance, attested to by the structural range of organic compounds(PAHs, PCBs, PCDDs, Chloro Phenols, etc.) recovered from contaminated sediments.

It is well established that many heavy molecular weight and high Kd compoundsafter introduction into the wetlands environment are not readily accessible to chem-ical recovery. This does not necessarily imply, however, that they are of no environ-mental significance: the degree to which they are desorbed and therefore becomeaccessible to biota is a central issue that has implicaitons both for toxicity of

Figure 6.16 Biofilms dominate the sediment–water interface and the surfaces of litter andstanding dead material.

Biofilm formation



xenobiotics and for their resistance to microbial attack. Results that indicate decreas-ing recoverability and irreversible sorption from the sediment phase with increasingtime from deposition may be accommodated under the general description of aging.26

Details of the mechanisms by which xenobiotics are “bound” to components of thesediment phase have not been fully established, although several plausible hypoth-eses have been put forward. Proposed mechanisms of interaction include ionic andcovalent binding, long range (Van der Waals) forces, or sorption by undefinedmechanisms. A recent study27 indicated 30–62% of chlorobenzene and 47–83% ofphenathrene of the adsorbed mass residing in the desorption-resistant fraction.

In another recent study partitioning of hexachlorobenzene (HCB), lower chlori-nated benzenes, and hexachlorbutadiene (HCBD) between the plant–air interfacewas measured in field and laboratory studies using bark, leaf, and leaf litter of baldcypress (Taxodium distichum) and black willow (Salix nigra).28 A first order kineticmodel was applied to the experimental results to determine a plant–air partitioncoefficient (KPA). The role of wetland vegetation in buffering air emissions fromwaste sites was established.

With high microbial populations compared to terrestial environments, wetlandsoils are highly effective at consuming and processing labile, complex organics. Thesupply of low molecular weight organic substrates from decomposition of organicmatter for dechlorination in organic rich wetland sediments is important to under-standing the potential of using sediment microorganisms for natural degradation ofchlorinated organic compounds. A recent study29 indicated that 2,3,5,6-TetrachloroPhenol can be dechlorinated in wetland sediments and that 2,4,5-Trichloro Phenol,and 2,5-and 3,4-dichlorophenols are major intermediates before the eventual dechlo-rination to 3-chlorophenol is achieved.

In another study30 intrinsic biodegradation of PAHs was evaluated using twomicrobiological analyses: heterotrophic bacterial production using 3H-Leucine incor-poration to estimate the growth rate of the entire community and PAH mineralizationrates using degradation of 14C-naphthalene, phenanthrene, and fluoranthene to 14CO2.Bacterial growth was generally reduced at locations that had PAH concentrationsabove 100 ppm. However, PAH mineralization rates were significant across the siteand PAH mineralization accounted for up to 15% of the heterotrophic bacterialgrowth substrate demand.

Additional investigations revealed that complete mineralization of hexachlo-robenzene (HCB) is possible in heavily anaerobic, methanogenic wetland sedimentenvironments.31 HCB was reductively dechlorinated to isomes of dichlorobenzene(DCB) and monochlorobenzene. After further reductive dechlorination of DCB, allthe monochlorobenzene was degraded to CO2 under methangenic conditions. Thisstudy also found that uncomplexed Cd and Pb in the sediment porewater wasinhibitory to the reductive dechlorination reactions.

Research focused on the transformations of alachlor and atrazine (members ofthe chloroacetamide and chloro-s-triazine herbicide classes, respectively)32 revealedthat degradation is promoted by naturally occurring inorganic sulfur nucleophiles inanoxic salt marsh porewaters. Bisulfide (HS) and polysulfides (Sn

2– ) have beenreported in salt marsh porewaters as high as 5.5 mm and 0.43 mm, respectively. Therates of sulfhydrolysis in these environments may greatly exceed rates of other



removal processes (biotic or aerobiotic) for the above compounds and may representan important sink for these herbicides in salt marsh environments.

Studies have examined the relationship between irreversible sorption (aging)and bioavailability in the vegetated wetland environment using plant uptake as anendpoint. In a recent study using black willow (Salix nigra), and Scirpus olney, itwas found that plant uptake of about 4–6% of the phenanthrene was observed insediment with phenanthrene in a fully desorption resistant state.33 Both plantsappeared to access the desorption resistant phase to some extent. Two mechanismsappear to be important: the direct uptake of the porewater in equilibrium with thedesorption resistant phase, and the sorption of the organic onto the root followedby uptake.

Other studies have reported that phenolic compounds at high concentrations canbe remediated in treatment wetlands using plants such as Schoenoplectus and Phrag-mites. Trinitrotoluene (TNT), cyclotrimethaylenetrinitramine (RDX), and cyclotet-ramethylenetetranitramine (HMX) at concentrations of 4, 4, and 0.10 ppm, respec-tively, were treated in a treatment wetland planted with canary grass, woolgrass,sweet flag, parrot feather, sago pond weed, water stargrass, and elodea. The treatmentefficiencies were greater than 95%.34

6.6.3 Removal of Inorganics

Reduction of phosphorus and nitrogen compounds requires the longest detentionof any pollutants. Approximately 90% of the incoming inorganic load can be elim-inated by 14 days detention in a surface flow wetland.1 These substances are required,at low loading rates, to sustain a healthy wetland. Nitrogen removal may requireactive or passive reaeration to promote nitrification; denitrification requires a carbonsource to fuel the denitrifer population.

In a recent study using native wetland plants such as bulrushes (Scirus spp.),cattails (Typha spp.), and sedges (Carex spp.), an influent water containing 62 mg/Lof NO3

– was treated to consistently nondetectable levels of effluent quality.7 Thistest is continuing to see whether perchlorate also can be treated by the same system.

It is not possible to forecast the types of substances that will be contained inleachates over the entire history or anticipated discharge. Treatment wetlands havecapabilities for removing a wide variety of substances, and hence operate as “broadspectrum” treatment technology. They have the further property of pulse averagingbecause of long detention times. A brief “spike” of a given substance will, at aminimum, be diluted by the relatively large volume of water in the wetland. Aver-aging is approximately over the detention time of the wetland.

6.6.4 Wetland Morphology, Hydrology, and Landscape Position

Wetlands are often constructed to remove pollutants from ground or surfacewaters. The intention is that constructed wetlands will degrade or sequester pollutantsso they do not pose a risk to humans and wildlife downstream of pollutant sources.Pollutants retained in constructed wetlands may be converted to less toxic forms orsequestered in the sediments or biota of wetlands. Ultimately, pollutants may be



removed from wetlands by harvesting soils and biota and transferring them to long-term storage sites such as landfills.

While wetlands may remove pollutants from ground and surface waters, pollut-ants may leave wetlands through aquatic-terrestrial food chain links. Many birdsand mammals feed on larger aquatic invertebrates and vertebrates that may accu-mulate significant amounts of pollutants. Thus, factors that promote the accumulationof pollutants in larger aquatic organisms can lead to risks to wildlife and, ultimately,human health. By managing the factors that promote accumulation of pollutants inorganisms inhabiting constructed wetlands, a manager can reduce the ecological riskthese prey items represent to terrestrial organisms.

The relationships among wetland landscape position, hydrology, morphology,and toxic accumulation in fish in natural depression wetlands can be manipulatedin constructed wetlands to reduce risk to wildlife from accumulated pollutants. Thedistribution of wetlands in the landscape and the amount of time they hold waterduring the annual hydrological cycle determines the amount of time fish will bepresent; wetlands close to other aquatic habitats and holding water for much of theyear are more likely to have fish populations. Fish in wetlands with shallow maxi-mum depth and greater water level fluctuations accumulate significantly more toxins.These relationships suggest: 1) constructed wetlands should be located in the land-scape such that the chance of colonization by fish is minimized; 2) when possible,stable water levels should be maintained in constructed wetlands; and 3) wetlandsshould be constructed with steep sides and flat bottoms.

REFERENCES

1. Kadlec, R. H. and R. L. Knight, Treatment Wetlands, CRC Press, Boca Raton, FL,1996.

2. Moshiri, G. A., Constructed Wetlands for Water Quality Management, Lewis Pub-lishers, Chelsea, MI, 1993.

3. Birkbeck, A. E., D. Reil, and R. Hunter, Application of natural and engineeredwetlands for treatment of low-strength leachate, in Constructed Wetlands in WaterPollution Control, P. F. Cooper and B. C. Findlater, Eds., Pergamon Press, Oxford,UK, 1990, 441–418.

4. Maehlum, T., Treatment of landfill leachate in on-site lagoons and constructed wet-lands, Proc. 4th Int. Conf. Wetland Syst. Water Pollut. Control, Guangzhou, China,553–559, 1994.

5. USEPA, Constructed Wetlands for Wastewater Treatment and Wildlife Habitat; 17Case Studies, EPA 832-R-93-005, 1995.

6. CH2M-Hill and Payne Engineering, Constructed Wetlands for Livestock WastewaterManagement, EPA Gulf of Mexico Program, Nutrient Enrichment Committee: StennisSpace Center, January, 1997.

7. Cole, S., The emergence of treatment wetlands, Environ. Sci. Technol., 218–222, 1998.8. U.S. Soil Conservation Service (USSCS), Hydric Soils of the United States, National

Technical Committee for Hydric Soils, Washington, D.C., 1987.9. Mackay, D. and P. J. Leinonen, Rate of evaporation of low-solubility contaminants

from water bodies to the atmosphere, Environ. Sci. Technol., 9, 1178–1180, 1975.



10. Shugai, D. et al., Removal of priority organic pollutants in stabilization ponds, WaterRes., 28, No. 3, 681–685, 1994.

11. Srinivasan, K. R. and R. H. Kadlec, Wetland Treatment of Oil and Gas Well Waste-waters, Report to U.S. Dept. of Energy on Contract DE-AC22-92MT92010, 55, 1995.

12. Peng, J., J. K. Bewtra and N. Biswas, Effect of turbulence on volatilization of selectedorganic compounds from water, Wat. Env. Res., 67, No. 1, 101–107, 1995.

13. Boethling, R. S. and D. Mackay, Handbook of Property Estimation Methods ofChemicals, Lewis Publishers, Boca Raton, FL, 2000.

14. Eger, P. et al., The use of wetland treatment to remove trace metals from minedrainage, Constructed Wetlands for Water Quality Improvement, G. A. Moshiri, Ed.,Lewis Publishers, Boca Raton, FL, 1993, 171–178.

15. Reily, J. M. and H. A. Wojnar, Treating and reusing industrial wastewater, WaterEnviron. Technol., 52–53, 1992.

16. Haffner, W. M., Palmerton zinc superfund site constructed wetlands, paper presentedat American Society for Surface Mining and Reclamation, Duluth, MN, 1992.

17. Noller, B. N., P. H. Woods and B. J. Ross, Case studies of wetland filtration of minewaste water in constructed and naturally occurring systems in Northern Austrailia,Water Sci. Technol., 29, No. 4, 257–265, 1994.

18. Crites, R. W., R. C. Watson, and C. R. Williams, Removal of metals in constructedwetlands, conference preprint, Water Environ., 68th Ann. Conf., Miami, FL, 1995.

19. Delgado, M., M. Biggeriego and E. Guardiola, Uptake of Zn, Cr, and Cd by WaterHyacinths, Water Res., 27, No. 2, 269–272, 1993.

20. Strecker, E. W., R. R. Horner, and T. E. Davenport, The Use of Wetlands for Con-trolling Stormwater Pollution, The Terrene Institute, Washington, D.C., 1992, 66.

21. Zhang, T., J. B. Ellis, D. M. Revitt, and R. B. E. Shutes, Metal uptake and associatedpollution control by Typha latifolia in urban wetlands, in Constructed Wetlands inWater Pollution Control, P. F. Cooper and B. C. Findlater, Eds., Pergamon Press,Oxford, UK, 1990, 451-459.

22. Masscheleyn, P. H., R. D. Delaune, and W. H. Patrick, Jr., Arsenic and seleniumchemistry as affected by sediment redox potential and pH, J. Environ. Qual., 20,522–527, 1991.

23. Alvord, H. H. and R. H. Kadlec, The interaction of atrazine with wetland sorbents,Ecol. Eng., 5, No. 4, 469–479, 1995.

24. Lorah, M.M. and L. D. Olsen, Natural attenuation of chlorinated volatile organiccompounds in a freshwater tidal wetland, Wetlands Remed.: Int. Conf., Salt LakeCity, November, 1999.

25. Burris, D. R. and E. J. O’Loughlin, Reductive dehalogenation of trichloroethenemediated by wetland DOC-transition metal complexes, Wetlands Remed.: Int. Conf.,Salt Lake City, November, 1999.

26. Neilson, A. H., Organic Chemicals: An Environmental Perspective, CRC/Lewis Pub-lishers, Boca Raton, Florida, 1999.

27. Pardue, J. H. and W. S. Shin, Desorption resistance of organic compounds in wetlandsould, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999.

28. Leppich, J., J. H. Pardue, and W. A. Jackson, Plant – Air partitioning of chloroben-zenes in wetland vegetation at a superfund site, Wetlands Remed.: Int. Conf., SaltLake City, November, 1999.

29. Chiang, S. V. and F. M. Saunders, Intrinsic microbial dechlorination of 2,3,4,6-tetrachloro phenol in anaerobic wetland sediment slurry, Wetlands Remed.: Int. Conf.,Salt Lake City, November, 1999.



30. Montgomery, M. T. et al., Measuring intrinsic bacterial degradation of PAHs in a saltmarsh, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999.

31. Jackson, W. A. and J. H. Pardue, Natural attenuation case study for chlorogenzenesin a forrested wetland, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999.

32. Roberts, A. L. and K. A. Lippa, Reactions of herbicides with reduced sulfur speciesin salt marshes, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999.

33. Gomez-Hermosillo, C., Bioavailability of desorption resistant phenanthrene to wet-land plants, Wetlands Remed.: Int. Conf., Salt Lake City, November, 1999.

34. L. L. Behrends et al., Phytoremediation of explosives contaminated groundwaterusing constructed wetlands, Wetlands Remed.: Int. Conf., Salt Lake City, November,1999.


313

CHAPTER

7

Engineered Vegetative Landfill Covers

CONTENTS

7.1 Historical Perspective on Landfill Practices................................................3147.2 The Role of Caps in the Containment of Wastes........................................3157.3 Conventional Landfill Covers ......................................................................3167.4 Landfill Dynamics........................................................................................3177.5 Alternative Landfill Cover Technology .......................................................3217.6 Phyto-Cover Technology..............................................................................321

7.6.1 Benefits of Phyto-Covers over Traditional RCRA Caps.................3267.6.2 Enhancing

In Situ

Biodegradation...................................................3267.6.3 Gas Permeability

......................................................................327

7.6.4 Ecological and Aesthetic Advantages

.........................................327

7.6.5 Maintenance, Economic, and Public Safety Advantages

..............329

7.7 Phyto-Cover Design .....................................................................................3297.7.1 Vegetative Cover Soils .....................................................................3307.7.2 Nonsoil Amendment ........................................................................3317.7.3 Plants and Trees ...............................................................................331

7.8 Cover System Performance..........................................................................3327.8.1 Hydrologic Water Balance ...............................................................3327.8.2 Precipitation .....................................................................................3357.8.3 Runoff...............................................................................................3357.8.4 Potential Evapotranspiration — Measured Data .............................3377.8.5 Potential Evapotranspiration — Empirical Data .............................3397.8.6 Effective Evapotranspiration

......................................................340

7.8.7 Water Balance Model

................................................................343

7.9 Example Application....................................................................................3447.10 Summary of Phyto-Cover Water Balance....................................................3477.11 General Phyto-Cover Maintenance Activities .............................................348

7.11.1 Site Inspections ................................................................................3487.11.2 Soil Moisture Monitoring ................................................................349



7.11.2.1 Drainage Measurement.....................................................3507.11.3 General Irrigation Guidelines ..........................................................3527.11.4 Tree Evaluation ................................................................................356

7.11.4.1 Stem ..................................................................................3567.11.4.2 Leaves ...............................................................................357

7.11.5 Agronomic Chemistry Sampling .....................................................3587.11.6 Safety and Preventative Maintenance..............................................3597.11.7 Repairs and Maintenance.................................................................359

7.12 Operation and Maintenance (O&M) Schedule............................................3597.12.1 Year 1 — Establishment ..................................................................3607.12.2 Years 2 and 3 — Active Maintenance.............................................3607.12.3 Year 4 — Passive Maintenance .......................................................361

7.13 Specific Operational Issues..........................................................................3627.13.1 Irrigation System Requirements ......................................................3627.13.2 Tree Replacement.............................................................................362

References..............................................................................................................362

Maintaining and enhancing the closed landfill as a bioreactor requires modifi-cation of design and operational criteria normally associated with traditionallandfill closure…

7.1 HISTORICAL PERSPECTIVE ON LANDFILL PRACTICES

The practice of using shallow earth excavations, or landfills, for disposal of liquidand solid waste has a very long history. Landfill practices basically followed thedesign philosophy of “out of sight, out of mind” in that a pit or trench was excavatedinto the ground, waste was placed into the excavation, and, when it was full, theexcavation was covered with soil and abandoned. If thought was ever given to thematter, it was likely assumed that the soil surrounding the waste effectively preventedcontaminant migration from the burial zone.

It was not until 1976, with the passage of the Resource Conservation andRecovery Act (RCRA), and 1980, with the passage of Comprehensive EnvironmentalResponse, Compensation, and Liability Act (CERCLA) that federal and state regu-lations mandated much improved methods for disposal of waste in landfills. Todaythere are a plethora of federal and state regulations controlling all aspects of landfilldisposal of municipal, radioactive, and hazardous waste. The problem in the U.S.,however, is that hundreds of thousands of landfills were operated and then decom-missioned prior to the requirements of current regulations. Many of these old landfillsnow come under the closure requirements of RCRA or CERCLA, depending on theagreements between the responsible parties. In 1989, U.S. Environmental ProtectionAgency (USEPA) stated that there are 226,000 sanitary landfills in the U.S. requiringevaluation for potential risks to human and environmental receptors.

1

Regardless of the corrective action imposed on these old sites, almost all of themwill require installation of a new cover as a final step in the closure process. The


ENGINEERED VEGETATIVE LANDFILL COVERS 315

design of most landfill covers in the U.S. has been based on criteria developed byEPA for use in closing either RCRA subtitle C (hazardous waste) or subtitle D(municipal solid waste) landfills. Two major themes emerge in reviewing recentwork in landfill cover design:

2

1) there has been an overemphasis on regulatorycompliance, thus inhibiting innovative and creative design that looks at the entirelandfill system as a holistic biogeochemical environment, and 2) there are fewpublished data on field performance of constructed cover systems and their impactson the biogeochemistry of the groundwater within the footprint of the landfill.

7.2 THE ROLE OF CAPS IN THE CONTAINMENT OF WASTES

Because of the expense and risk associated with treating or removing largevolumes of landfill wastes, remediation usually relies upon containment, whichrequires the construction of a suitable cover. Both regulators and the public usuallyaccept covers as part of the presumptive remedy for final landfill remediation;therefore, covers are likely to be included in the optimal remedial actions for closureof most landfills.

The intent of landfill remediation is to protect the public health and the environ-ment. In keeping with this intent, a modern philosophy has evolved requiring con-taminants in the waste to be isolated from receptors and contained within the landfill.As a result, landfills have become warehouses in which wastes are stored for anindefinite time, possibly centuries.

There are fundamental scientific and technical reasons for placing a cover onlandfill sites. Although regulations are often the most apparent influence governingthe selection and design of landfill covers today, these regulations were draftedbecause of specific environmental concerns and were based upon scientific andtechnical understandings. The three primary requirements for landfill covers are to:

•

Minimize infiltration:

water that percolates through the waste may dissolvecontaminants and form leachate, which can pollute both soil and groundwater asit travels from the site.

•

Isolate wastes:

a cover over the wastes prevents direct contact with potentialreceptors at the surface and prevents movement by wind or water.

•

Control landfill gas:

landfills may produce explosive or toxic gases, which, ifallowed to accumulate or to escape without control, can be hazardous.

Landfills have been covered by barriers for years, usually built with little regardfor the monetary and environmental costs associated with constructing and main-taining them. A typical landfill cover design consists of a sequence of layeredmaterials to control landfill gas infiltration and promote internal lateral drainage.The uppermost layer of a landfill cover consists of a vegetative soil layer to preventerosion, promote runoff, and insulate deeper layers from temperature changes. Thelandfill cover is not a single element but a series of components functioningtogether.

3



Landfill covers are designed to minimize infiltration of rainfall and melting snowinto the landfill in order to minimize postclosure leachate production. This objectiveis achieved by converting rainfall into surface runoff and infiltration into evapotrans-piration and lateral drainage without compromising cover integrity. Secondary per-formance objectives of landfill cover design include the following:

3

minimize post-closure maintenance; return the site to beneficial use as soon as possible; make thesite aesthetically acceptable to adjacent property owners; accommodate post-closuresettlement of the waste; address gas and vapor issues; provide stability againstslumping, cracking, and slope failure; provide resistance to disruption by animalsand plants; and comply with landfill closure regulations.

The design features of a landfill cover are varied to affect changes in the overallwater balance within the landfill to meet primary landfill cover objectives. The designadopted must take into account numerous other considerations, including costs, longterm maintenance implications, and construction risks. The relatively large areasthat landfill covers protect, and the thickness and number of individual layers withinthem, make covers a cost-intensive component of landfill facility design.

7.3 CONVENTIONAL LANDFILL COVERS

Nearly all conventional landfill covers in current use incorporate a barrier withinthe cover. The “impermeable” barrier layer is intended to prevent water from movingdownward in response to the force of gravity. In effect, these covers are designedto oppose the forces of nature. Barrier-type covers commonly include five layersabove the waste (Figure 7.1).

1

The top layer consists of cover soil typically two feetthick and supports a grass cover that provides erosion control. The barrier layerconsists of either a single low-permeability barrier or two or more barriers incombination. The fourth layer is designed to remove landfill gases as they accumulateunderneath the barrier layer. The bottom layer is a foundation layer of variablethickness and material; its purpose is to separate the waste from the cover and toestablish sufficient gradient to promote rapid and complete surface drainage fromthe finished cover.

The barrier layer is the defining characteristic of conventional landfill covers. Itmay be composed of compacted clay, a geomembrane, a clay blanket, or two ormore layers of materials in combination. A compacted clay layer is frequentlyspecified to have a maximum saturated hydraulic conductivity (K)

≤

1

×

10

–7

cm/sec.In contrast, both the drainage and gas collection layers are constructed to enhanceflow and commonly contain washed and selectively sieved sand, gravel, or speciallydesigned synthetic materials.

The soil in the top layer of barrier-type covers is usually too thin or has inadequatewater holding capacity to store infiltrating precipitation during a large storm. Thesecovers rely on barrier layers and rapid drainage through lateral drainage layers toprevent precipitation from reaching the waste. Barrier-type covers must accommo-date specific site conditions, and supplemental components are sometimes added.For example, gravel may be added to the surface soil in desert regions to control



wind erosion, or a layer of cobbles may be used with the cover to discourage animalburrowing into the waste.

7.4 LANDFILL DYNAMICS

Landfills that contain a large amount of organic, putrescible materials (such asmunicipal solid waste) literally function as bioreactors. Most “landfill bioreacters” ingeneral contain anaerobic and/or facultative microorganisms. Landfill leachate is gen-erated as a result of the percolation of water or other liquids through the waste andalso due to the accumulation of moisture generated as a result of microbial degradationof waste. Leachate is a concentrated fluid containing a number of dissolved andsuspended materials, specifically, high concentrations of organic compounds (organicacids, hydrocarbons, etc.) and certain inorganic compounds (ammonia, sulfates, dis-solved metals, etc. characteristic of the parent waste materials) from which it is derived.In addition, natural microbial activity in landfills also results in the generation of gasessuch as methane, carbon dioxide, ammonia, and hydrogen sulfide, a fraction of whichwill be dissolved in the leachate and may be introduced into the groundwater.

Numerous landfill investigation studies

4

have suggested that the stabilization ofwaste proceeds in sequential and distinct phases. The rate and characteristics ofleachate produced and biogas generated from a landfill vary from one phase toanother and reflect the processes taking place inside the landfill. These changes aredepicted in Figure 7.2.

Figure 7.1

Typical single barrier cover system.

Waste

Foundation Layer

ProtectiveCover Layer

Vegetative Layer0'-6"

1'-18"

2'-30"(min.)

Vegetation

Topsoil

Common Borrow Material

Geocomposite(Textile-Net-Textile)

40-mil LLDPE



The initial phase is associated with initial placement of waste and accumulationof moisture within landfills. Favorable biochemical conditions are created for thedecomposition of waste. During the next phase, transformation from an aerobic toanaerobic environment occurs, as evidenced by the depletion of oxygen trappedwithin and introduced to landfill media and continuous consumption of nitrates andsulfates. Subsequent phases involve the formation of organic acids and methane gas.During maturation phase, the final state of landfill stabilization, available organiccarbon and nutrients become limiting, and microbial activity shifts to very low levelsof activity. Gas production dramatically drops and leachate strength remains constantat much lower concentrations than in earlier phases.

Biochemical decomposition of putrescible solid waste is shown below by anexample (Equation 7.1). Typical landfill gas composition during peak activity as abioreactor is: 60% methane, 40% carbon dioxide, 5–10% other gases, and 0.3–1.0%VOCs and non-monitored organic compounds. Gas generation rates during peakactivity typically fall within the ranges of 5–15 ft

3

per pound of refuse per year.

9

(7.1)

Due to very high gas pressures generated at the source areas within the landfill(up to 4 atmospheres), migration of dissolved contaminants into the gaseous phasecould be a serious concern. Contaminants transferred into the gas phase could bereadsorbed in the waste above the water table, or dissolve in the moisture, condensein the waste zone, or migrate away from the landfill. The potential for contaminantmigration from the dissolved phase into the landfill gas can be evaluated as shownin Equation 7.2, and Figure 7.3.

Figure 7.2

Description of leachate and gas concentration changes during landfill lifecycle.

C H O H O CH CO6 10 5 2 4 23 3AnaerobicBacteria

→ +



Under non-equilibrium conditions:

(7.2)

where

= transfer rate from gas to water

K = phase transfer coefficientH = Henry’s Law Constant of VOCC

g

= gas phase concentration of VOCC

w

= water phase concentration of VOCA = gas/liquid contact area

The progress toward final stabilization of any landfill and the organic waste init is subject to physical, chemical, and biological factors within the landfill environ-ment, age and characteristics of the waste, operation and management controlsapplied, as well as site-specific external conditions.

Although barrier layers are sometimes referred to as “impermeable” layers, inpractice this is seldom true. An extensive review of failures and failure mechanismsfor compacted soil covers in landfills was performed and emphasized that “…naturalphysical and biological processes can be expected to cause [clay] barriers to fail inthe long term.”

5

Another study discussed a field test conducted in Germany in which,

Figure 7.3

Equilibrium mass transfer conditions of contaminants into landfill gas.

dmdt

K C HC Ag w= −( )

dmdt



seven years after construction, percolation through the compacted clay was almost200 mm/yr and increasing. Geomembrane barriers are also prone to leak.

6

Othershave traced most leaks in geomembranes to holes left by construction.

7,8

A modification of the typical barrier cover is the subtitle D cover (Figure 7.4)that relies on compaction to create a layer of soil with reduced K value. Usedprimarily for municipal landfills in dry regions, its use and components are specifiedin subtitle D of RCRA (40 CFR, Part 258.60), hence the name. From the surfacedownward, the cover includes an erosion control layer and a layer of compactedsoil. A major advantage of the subtitle D cover is that its construction cost is lowerthan for an RCRA subtitle C cover. Even though it has gained regulatory and publicacceptance, the subtitle D cover cannot ensure long-term protections against infil-tration of water into the waste, even in dry regions, because 1) the topsoil layer haslimited water holding capacity, 2) there is no drainage layer, 3) few roots can growin the barrier layer to remove water, and 4) soil freezing and root activity are likelyto increase the K value of the barrier soil layer over time.

Figure 7.4

Subtitle D cover for municipal solid waste landfills.

0.47 m

0.15 m

Waste

Soil BarrierK ≤1 x 10-5 cm/sec

Topsoil

Precipitation

Foundation -Gravel

Runoff



7.5 ALTERNATIVE LANDFILL COVER TECHNOLOGY

Alternative covers to the RCRA subtitle C or D design include evapotranspiration(ET) covers and capillary barriers. The ET cover uses no barrier or horizontaldrainage layers; it is designed to work with the forces of nature rather than attemptingto control them. An ET cover in its simplest form is a vegetated soil cover with asufficiently deep soil profile so that infiltrated water is stored until removal byevaporative losses from the soil surface and by plant roots at depth in the profile. Acapillary barrier also relies on water removal by ET, but is designed such that waterstorage near the surface is enhanced to promote the efficient removal of infiltratedwater by the ET process. Optimization of material types and thicknesses for capillarybarriers is critical to their effective performance. The use of sands or clays as thefine-soil component in the capillary barrier has proven to be less effective in storingwater than silt loams. Capillary barriers can be thought of as enhanced ET covers— alternative cover systems that work best in semi- and/or arid environments wherehigh ET rates and low precipitation make it possible to remove all infiltrated waterby ET. However, even in arid environments there are situations where ET coversand capillary barriers can allow excessive percolation, particularly where the soilused in the cover design has insufficient storage capacity to accommodate wintersnow melt events.

7.6 PHYTO-COVER TECHNOLOGY

The phyto-cover is the most popular application of the ET cover and is anengineered agronomic system that harnesses the natural transpiration process ofplants to limit percolation to the groundwater. A phyto-cover relies on shallow- anddeep-rooted plants to create a thick root zone from which the plants can extractavailable moisture. In effect, the plants serve as natural, solar-powered “pumps” towithdraw soil moisture and either convert it into biomass or evaporate it throughtheir leaves. The withdrawal rate of the botanical pumps is limited by the availableenergy (sunlight), rate of growth, and available soil moisture; withdrawal virtuallyceases during winter dormancy. Accordingly, the depth and composition of the rootzone must be sufficient to store accumulated water like a sponge and hold it untilthe plants remove it. Properly designed, this “sponge and pump” water removalsystem (Figure 7.5) can limit water from percolating below the root zone and canbe equally protective of groundwater as a RCRA cap. Thus, a phyto-cover servesas a functional alternative to natural clay, geocomposite, or geosynthetic membranecap, yet offers several advantages over those technologies.

The effectiveness of poplars in maintaining low soil moisture levels was firstdocumented by data collected from a phyto-cover application in Iowa.

10

The phyto-cover consistently maintained soil moisture levels substantially below the soil’s fieldcapacity (i.e., the amount of water that soil can retain without allowing percolation)of 40–45%. Soil dryness was maintained by the trees’ prodigious water extractingability. The capacity of certain trees such as hybrid poplar and willow trees to extractsoil moisture has been demonstrated by monitoring data from landfill at many sites.



The poplars are employed at this site not as a cover, but to treat collected landfillleachate, which is applied to the poplars during the growing season. The total amountof water extracted from the soil in one growing season by these two- and three-year-old poplars was equivalent to about 62 inches of precipitation.

10

One of the most important design considerations for a phyto-cover is choosingappropriate tree species and varieties. Selected trees must be capable of achieving thedesired treatment objective and adapt to the irrigation water, soils, and climate of thesite. Typically, achieving the highest possible rate of evapotranspiration is an importantgoal. Critical site conditions for plant selection include soil chemistry, irrigation wateror groundwater chemistry, and adaptation to pests and diseases of the area. Any factorthat compromises tree health and growth will reduce performance. For example, hybridpoplar clones that include either

trichocarpa

or

maximowiczii

parents are quite sus-ceptible to Septoria canker if used in the U.S. midwest.

10

Especially for the commonly used Salicaceae, a number of different types ofplant materials may be used. These include stem cuttings, whips or poles, and bareroot or potted material. Use of larger or rooted plant material will result in morerapid establishment and reduced weed competition, but plant material and plantingcosts are much higher than for smaller material. Whips and poles are commonlyused for deep planting applications. Economics, especially planting costs, drive mostlarger installations (>5 ha) toward short stem cuttings.

Certain varieties may result in a more valuable final wood product because ofstraighter stems or better paper processing properties. Significant differences indamage from voles has been observed among hybrid poplar trees at phytoremediationsites. Salt tolerance is a very important selection criterion, as differences betweenspecies and varieties can be significant. Only limited data for Salicaceae are currentlyavailable to guide design, but a number of relevant research programs are ongoing.

For an increasing number of sites, use of non-native species is unacceptable forlocal community groups and sometimes for regulators. Use of native material willgenerally ensure resistance to local pests and disease, but may not afford the greatestefficiency.

Figure 7.5

Conceptual phyto-cover diagram.

Infiltration

Evaporation

Precipitation Poplars

RootDepth

Evapo-transpiration

Surface Runoff

Topsoil Storage

Native Soil Storage

Daily Cover

Waste

10'-0"



Once the tree system forms a complete canopy, spacing has little effect onevapotranspiration or nutrient requirements. The impact of spacing on hydraulic andnutrient loading is primarily an early establishment phase concern. Establishingdense initial plantings with the intention of thinning may provide small increases inearly capacity, but thinning operations are often neglected and the resulting maturetree stands are excessively dense. Enough space must be left between tree rows toallow planned maintenance activities such as mowing or spraying.

The engineered phyto-cover system consists of densely planted, deep-rootedtrees and understory vegetation (perennial rye grass and clover). Photographs ofhybrid poplar tree stands of varying ages are shown in Figure 7.6. The water-holdingroot zone (“sponge”) includes the existing topsoil and fill at the site (includingintermediate and daily cover soil) supplemented with additional soil or soil amend-ments as dictated by design calculations. A phyto-cover will provide a protective,living “skin” for a landfill that permanently heals the wound to the landscapeoriginally created by anthropogenic activities. This skin can equal the percolation-blocking performance of a “rain coat” RCRA cap while being substantially morecost effective and providing additional benefits. The final design of a phyto-coveroften includes provisions for monitoring soil moisture levels to ensure that perfor-mance criteria are met.

Engineered phyto-cover systems have been applied to contain spilled petrochem-icals and cover landfills, as well as buffers to remove nitrogen from industrial andmunicipal wastewater. Sites where phyto-covers have been installed and recentresearch and demonstration sites for phyto-cover systems include the following:

2,10-13

• A 15-acre construction debris landfill in Beaverton, OR was covered with treesin 1990 as an alternative to excavation of the fill, the installation of a liner, andthen recovering with a geomembrane. The phyto-cover is serving to protectgroundwater cost-effectively. The owner has continued to expand the cover as newareas are closed.

Figure 7.6

Phyto-covers: comparison of two-year-old and four-year-old growth of a phyto-cover (courtesy of Licht, 1998).

Four-year Old Poplar Trees

Two-year Old Stand of Poplars



• From 1992 to 1993, the Riverbend Landfill in McMinneville, OR planted a 17-acre phyto-cover to manage landfill leachate water and soluble compounds. Allnutrient and water cycling results indicate the cap is achieving all regulatoryrequirements for ammonia treatment and ground protection.

• From 1993 to 1995, a 15-acre perimeter buffer was planted to reduce infiltrationfrom upgradient runoff, grow a visual and sound barrier, and intercept downgra-dient leachate seepage. Data collected at the site indicate the cap has been suc-cessful in stopping all leachate.

• At the Grundy county Landfill in Grundy Center, IA, a two-acre cap and perimeterbuffer was planted from 1993 to 1994 to reduce leachate formation by installinga phyto-cap over a completed subtitle D cap. The cap also provides a visual andsound barrier, and intercepts downgradient leachate drainage.

• A three-acre poplar cap was planted in 1994 at the Bluestem 1 Landfill in CedarRapids, IA over a pre-subtitle D cap. The cap serves to reduce leachate formationvertically and intercepts downgradient leachate drainage. The data collected atthis site have been used in writing specifications for soil and compost coverrequirements and use of MSW waste as a planting media.

• A five-acre cap and perimeter boundary were planted in 1994 over a pre-subtitleD cap at the Bluestem 2 Landfill in Marion, IA to reduce leachate formation.Moisture management data from this cap have been used in subtitle D equivalencecomparison between a soil/clay cover and the “sponge and pump” concept fordeep rooted trees in four feet of rootable soil.

• In 1995, a ten-acre area was planted with poplar trees and a clover/grass understoryover a subtitle D cell filled with MSW and industrial waste. The Department ofEnvironment Quality and governor’s office were interested in future phytoclosuresfor many funded pre-RCRA landfills in Virginia that have been abandoned andare creating potential environmental risk. The trees are growing well and are beingmaintained by the owner. A soil moisture measurement system using time domainreflectometry (TDR) is used to monitor the impact of tree roots on vadose zonewater content. A drip irrigation system using collected storm water can controlthe water stress during periods when moisture in the root zone has been exhausted.

• At a railroad RCRA site in Oneida, TN, a one-acre area impacted by coal andcreosote from past manufacturing activities was covered with poplar trees andgrass in 1997. The site soils were amended with compost and mineral fertilizer,then trenched in the root zone. The trees and grass managed to accelerate biomassgrowth with resulting water uptake and

in situ

constituent removal. The sitegroundwater is monitored by a university research grant to measure groundwaterelevation and the containment of constituents. The concept is similar to landfillcapping where the phytosystem pumps water at high rates during the growingseason and minimizes groundwater movement during the dormant season.

• The Woodburn WasteWater Treatment Plant in Woodburn, OR has been a dem-onstration site since 1995; a full-scale installation took place in 1998. This site isthe first full-scale tertiary treatment of secondary municipal wastewater effluentand is being designed for no leakage through the root zone in the summer months.

• The Mid-Lakes Co-op site in Bonduel, WI used an aesthetically pleasing poplarcover over a spill plume to contain pollutant migration, make use of all availableprecipitation, protect public exposure, and remove constituents from the ground-water . Closure requirements for this site included planting trees, monitoring thedepth to groundwater, and monitoring groundwater quality over a three-year-period.



• In Staten Island, NY a phyto-cover consisting of poplars, willows, paulowia, andgrasses is being used to prevent constituent migration and formation of leachate.Enhancement of existing vegetation is expected to establish hydraulic control ofgroundwater by reducing water infiltration through the landfill materials.

• Evidence collected at a closed landfill in Elmore, OH indicates naturally occurringtrees have created a treatment barrier for leachate seeps. An evaluation of on-sitebox elder and osage orange trees yielded evidence of TCE uptake. An evaluationof the existing cover for supplemental enhancement for additional groundwaterremediation and restoration was then conducted.

• A poplar tree phyto-cover was installed in 1996 at a landfill in Acme, NC. Thetrees were planted in the most downgradient area of the landfill to stop leachatemigration. Groundwater constituent concentrations have dropped substantially inthe area of the poplar trees but not in areas where trees were not planted.

• From 1992 to 1993, over 2000 poplar trees were planted at a site in Anderson,SC to be used for processing waste from mining ore material. The waste was usedto fill low areas over six acres of the site. Data collected at the site indicateinfilitration and leachate formation is being controlled.

• In 1991, a succession of trees (willow and black locust), legumes, and grasseswere planted to dewater slurry waste at a site in Baton Rouge, LA. The wastematerial was in a slurry state from a depth of 6 inches to 30 feet below groundsurface. The planted vegetation reduced the hydrated state of the waste and theoccurrence of leachate through the impoundment.

• A process waste was placed as a slurry into an impoundment at a site in TexasCity, TX. Naturally occurring trees (osage, orange, and mulberry) and vegetationhave reduced the hydrated state of the top ten feet of the waste. Research on thesite has found that dewatering and net water removal are directly correlated to thesize of the trees.

• Ongoing research, funded by the USEPA Great Plain and Rocky Mountain Haz-ardous Substance Research Center involves planting trees at CERCLA sites tocontrol erosion and leaching of zinc, arsenic, lead, and cadmium.

• A grass/soil cover system is one of five alternative covers being evaluated bySandia National Labs in NM as part of an alternative landfill cover demonstrationstudy. Similar phyto-cover systems are being considered as potential demonstra-tion sites by USEPA ORD at sites in Tulsa, OK; Beatty, NV; and Hill Air ForceBase in CA.

• A phyto-cover system has been proposed and designed at the 95% completionlevel for the F.E. Warren Air Force Base Superfund Site in Cheyenne, WY. Thissite is currently being considered as a technology demonstration candidate byUSEPA-Region VIII.

• Pfitzer junipers have been used in a landfill cover field demonstration at Beltsville,MD. The juniper phyto-cover was installed over a clay layer to add to the “robust”cover development, but not as a replacement of the low-permeability layer. Theobjective of the demonstration study was to document the influence of junipersas water scavengers, yet maintain the water runoff performance of the low-permeability cap. Compared to a reference soil, the “bioengineered” juniper coverreduced infiltration; it was demonstrated that the mature plant system improvedthe system’s resilience to failure.

• Research regarding the establishment of sufficient vegetation to provide adequatebiomass growth with resultant evapotranspiration is being conducted at IdahoNational Engineering Lab, Idaho Falls, ID. This research focuses on four plant



species to deplete soil moisture, and the configuration of a capillary barrier and rootzone to prevent deep percolation during wet periods. The use of such phyto-covershas been demonstrated to be applicable to landfill sites in the semi-arid west.

• In Ljubljana, Slovenia, a ten-acre cover was planted with poplar trees in1993–1994 with the primary goal of protecting groundwater by reducing leachateformation through municipal and industrial wastes. Installation of the cover hasgreatly improved the aesthetics of the area and increased the value of the wildlifehabitat. The design concept is being considered as a model that will becomenational policy.

• The author also knows of many other phyto-cover applications in MA, OH, MD,NC, MI, PA, NY, NJ, and IL.

In 1998, USEPA began an effort to establish a design database and improve numer-ical prediction methods for alternative landfill covers. The initial task of the AlternativeCover Assessment Project (ACAP) was to catalog past and existing research effortsinto measurement of cover performance and to describe the current state of numericalprediction methods. The primary criterion was to measure percolation directly. Severalresearch sites operated by branches of the federal government were included in thisstudy. These sites include the national laboratories at Hanford, Sandia, Los Alamos,Savannah River, and Idaho Falls, and DOE locations at Monticello, UT, Nevada TestSite, DoD locations at California, Hawaii, Colorado, Utah, and others.

7.6.1 Benefits of Phyto-Covers over Traditional RCRA Caps

In addition to satisfying the critical antileaching requirement, phyto-covers pro-vide a number of significant pollution control, ecological, and economic benefitswhen compared to traditional RCRA caps:

• A phyto-cover actually enhances natural biodegradation processes, instead ofinterfering with them, as a RCRA cap would.

• A gas-permeable phyto-cover allows for passive venting of gaseous byproductsof biodegradation and allows oxygen to move into the fill to facilitate additionalbiodegradation.

• A phyto-cover provides a forest ecosystem and an attractive alternative to anRCRA cap.

• A phyto-cover can be installed with less cost and less risk to public safety thana RCRA cap and, once the cover is established, the system has a natural stabilitythat minimizes long-term maintenance requirements.

7.6.2 Enhancing

In Situ

Biodegradation

Installing a phyto-cover at a site has the potential to enhance biodegradation ofwaste materials, including organic waste and contaminants, in the root zone. Innatural ecosystems, high concentrations of indigenous soil microorganisms are foundin association with plant roots, because the roots exude a wide variety of compoundssuch as sugars, amino/acids, carbohydrates, and essential vitamins. These com-pounds not only sustain the microbial consortia, which can degrade many organiccompounds directly, but also enhance and accelerate cometabolic degradation of



other pollutants resistant to direct degradation. In addition, the plants themselveswill take up and metabolize or volatilize some of the organic contaminants in thefill. Finally, exuded organic acids also help in sequestering and immobilizing anymetals present in the root zone. By contrast, a RCRA cap provides no stimulationto natural biodegradation and would be expected to substantially change bio-geochemical conditions in the fill by trapping the gaseous byproducts of biodegra-dation (e.g., methane, carbon dioxide, and ammonia), thereby affecting factorscritical to natural attenuation mechanisms, such as pH and REDOX potential.

The main reason for the enhanced

in situ

biodegradation in landfills with phyto-covers is the ability of the atmospheric oxygen to transfer into the landfill. Theprimary mechanism transferring oxygen into the landfill is diffusion into the soilfrom the atmosphere, based on an excellent summary shown below:

14

The exchange of gases between the soil and the atmosphere … is facilitated by twomechanisms: mass flow and diffusion. Mass flow of air, which is due to pressuredifferences between the atmosphere and the soil air, is less important than diffusionin determining the total exchange that occurs. It is enhanced, however, by fluctuationsin soil moisture content. As water moves into the soil during a rain … air will beforced out. Likewise, when soil water is lost by evaporation from the surface or istaken up by plants, air is drawn into the soil. Mass flow is also modified slightly byother factors such as temperature, barometric pressure, and wind movement. Mostof the gaseous interchange in soils occurs by diffusion.

The minimum rate of oxygen diffusion at the bottom of the root zone wasestimated to be 5

×

10

–8

grams per centimeter, squared per minute, or 2340 poundsper year per acre.

14

The maximum rate could be up to 9200 pounds per year peracre. Over the surface of a 30-acre landfill, this translates into at least 70,000 poundsof oxygen per year into the landfill, which facilitates stabilization of the waste. Bycontrast, the single-barrier cap would admit only an estimated 75 pounds of oxygenor about one tenth of one percent of the influx that could support the aerobic naturalattenuation mechanisms (Figures 7.7a and b).

15

7.6.3 Gas Permeability

Unlike RCRA caps, which are essentially impermeable to gases and thereforerequire elaborate gas venting systems to deal with gases and vapors generated bybiodegradation of the fill, a phyto-cover is porous and permeable to gas. A phyto-cover can thus eliminate the need for a gas collection system at many sites. Equallyimportant, a phyto-cover will allow oxygen to migrate into the fill, which will helpto support additional aerobic biodegradation and thereby hasten the completion ofthe waste life cycle.

7.6.4 Ecological and Aesthetic Advantages

Both a phyto-cover and an RCRA cap are designed to be vegetated on the surface,but vegetation on a phyto-cover has the appearance of a tree farm and, eventually,



a forest, and serves the same ecological function as a forest while the RCRA cap iscovered with grass and, in order to protect the impermeable liner, must be preventedfrom functioning like local natural ecosystems. Specifically, maintenance of theintegrity of the RCRA cap’s impermeable layer dictates that deep-rooted plantspecies, such as trees and shrubs, not be allowed to colonize the site through naturalsuccession. Moreover, protection of the impermeable liner also requires that smallburrowing mammals, such as those normally associated with a meadow, must beperpetually monitored for and eliminated when found. By contrast, the trees of aphyto-cover provide nest sites for birds and other arboreal species and readily acceptin-fill by shrubs and native tree species, as deemed appropriate under site manage-ment criteria. Because no animal is likely to excavate below the deep root zone, itis not necessary to prevent native fauna from inhabiting the phyto-cover. Besidesoffering a preferred natural ambiance, the phyto-cover forest would also serve thecommunity as a noise pollution buffer and assist incrementally with global climateissues by fixing substantially more carbon dioxide from the atmosphere than a grassRCRA cap.

Figure 7.7a

Biogeochemical conditions and mass balance for a presumptive remedy.

Figure 7.7b

Biogeochemical conditions and mass balance for a holistic remedy.



7.6.5 Maintenance, Economic, and Public Safety Advantages

The ongoing maintenance requirements for an established phyto-cover are min-imal. Although relatively intensive monitoring for disease and pests is needed duringthe three growing seasons that the trees need to become fully established, mainte-nance activities thereafter are expected to be minimal because of the self-healingnature of the phyto-cover. Like a natural forest, the phyto-cover is expected to beresistant to wind and water erosion. Unlike a RCRA cap, which can suffer cracks,rips, and tears due to factors such as differential settling or physical intrusion, thephyto-cover maintains its integrity and actually heals itself with new root growth inresponse to physical disturbances. Thinning of trees may be undertaken in the futureto avoid crowding as the trees reach their mature size. However, the trees cut in athinning operation represent a valuable forestry crop, so revenue from their saleshould compensate for the operation’s costs.

The lower economic cost of the phyto-cover compared to the RCRA cap is accom-panied by lower noneconomic social costs in the form of safety risks. Studies haveshown that remedy implementation imposes risks of injury and death to site workers,neighbors, and the public using transportation routes. These risks are more certain andtypically substantially greater in magnitude than risks to the public from exposure tosite contaminants. For example, assuming that bulk construction materials can be foundat an average distance of 15 miles (i.e., 30 miles round trip) and using the U.S. truckfatality rate of 4.7

×

10

–8

/mile, construction of RCRA “C” cap at a 30-acre landfillsite could lead to an estimate of transportation fatalities risk of 0.033.

16

This estimatewill be further increased if the nontruck driver fatalities estimate is combined withthis. Since phyto-covers require less site work and fewer truckloads of importedmaterial, such as borrow soil and gas collection layer sand, constructing a phyto-coverinstead of an RCRA cap would involve less risk of an accidental injury or fatality toa construction worker and lower risks of traffic incidents associated with truckloadsof construction materials carried over local roads

Finally, unlike a RCRA cap, which locks the site into an “impermeable barrier”strategy, the phyto-cover system can be operated in a flexible way to take into accountthe results of ongoing performance monitoring data. For example, if performancedata show that native species perform as well as hybrid poplars in preventinginfiltration, then the natural transition to native species can be accelerated, to enhancethe ecological service provided by the area. By the same token, in the unlikely eventthat performance data show that a part of the cover is not performing to expectation,then a supplementary measure such as additional “sponge” or denser planting wouldbe available to improve performance.

7.7 PHYTO-COVER DESIGN

The typical components of an engineered phyto-cover system consist of vege-tative cover soils (existing and supplementary), soil amendments, nonsoil amend-ments, understory grasses and plants, and trees. An irrigation system is an optionalcomponent to ensure sufficient water for tree growth in case of drought. Irrigated



trees grow more rapidly, thus meeting closure objectives in less time; however, thereis often lack of a convenient water source or on-site operation and maintenancepersonnel to make an irrigation system feasible at a site. The need for on-siteirrigation should be based upon the expected water consumption of the trees.

7.7.1 Vegetative Cover Soils

The existing cover soil at many sites is sufficient to support an adequate rootsystem for healthy tree growth. This is evidenced by the vigorous growth of treesoften seen at abandoned landfills (Figure 7.8); however, the ability to grow trees isnot evidence that percolation and leachate production are controlled. Typically,natural stands of vegetation are not effective at controlling percolation. Therefore,sufficient soil and nonsoil amendments may need to be added to meet the require-ments for tree growth, and to achieve minimum land surface slopes to promotesurface drainage and to provide sufficient soil water holding capacity for storage tofunction as an adequate “sponge.” The amount of soil and nonsoil amendmentsnecessary must be determined by site-specific information, often collected in thelater stages of design.

Any supplemental cover soil added to achieve the required grades, as well assufficient water storage capacity, will comprise common borrow soils. Supplementalsoil should be placed in 6-inch thick and loose lifts, and be lightly compacted toachieve the minimum slope and thickness. This material is typically available fromseveral sources in the vicinity of most sites; the specific local source usually dependsupon availability during the construction period. The surficial lift of supplementalsoil and existing cover, depending upon which is exposed at the final grade surface,

Figure 7.8

Existing root penetration of a tree at a landfill site.



must be ripped in two directions following final grading to assure noncompactionand to prepare the surface to receive the nonsoil amendments.

7.7.2 Nonsoil Amendment

The addition of nonsoil amendments will increase the water-holding capacityand nutrient transfer properties of the common borrow soils. Typical nonsoil amend-ments include compost, chipped wood, digested sewage biosolids, lime-stabilizedsludge, manure, and other organic biomass. The incorporation of this type of organicmatter into the existing and supplemented soils will greatly increase the tilth, fertility,and water-holding capacity of the soil, and further reduce percolation through thecover. Biosolids compost and lime-stabilized sludge are readily available through acompost contractor. Typically a minimum 6-inch thick layer of organic amendmentsneeds to be applied to the soil surface after achieving final grade. This material isspread evenly in a six-inch thick layer on the area to be planted with the engineeredphyto-cover system and is ripped into the surficial soils to a depth of 14 to 18 inches.Ripping is performed in both an east-west and north-south orientation in order toachieve a uniform mixing within the soil profile. Finally, the site is tilled in prepa-ration for planting.

If the organic materials used for the nonsoil amendment have a high carbon tonitrogen ratio, fertilizer is added along with organic amendments to aid in stabilizingthese amendments and to provide sufficient nutrients to the rooting plants. Thisorganic amendment is expected to supply all micronutrients required by the plants.A mineral fertilizer is also added as needed, based on nutrient analyses of the appliedcompost, to supplement the macronutrient reserves of nitrogen, phosphorous, andpotassium. All amendment addition, ripping, and tilling is completed prior to under-story planting in the fall and before the trees are installed in the followingearly spring.

7.7.3 Plants and Trees

The area to be planted will generally exhibit a minimum 2% or greater grade;therefore, stabilization of the site cover material remains necessary to prevent ero-sion. Understory planting will be established for early erosion control and wateruptake during the first year. Understory establishment is a combination of annualand perennial grasses, such as varieties of rye, oats, wheat, barley, and fescue, appliedat a rate of 20 to 40 pounds per acre. This mixture of seed is designed to meet theshort- and long-term objectives of the understory. Annual species will be fast growingto control near-term erosion; perennial grasses will be deep-rooted species selectedas the primary long-term understory for the site. The long-term effectiveness of theoverstory is dependent upon establishment and long-term maintenance of the under-story, which understory depletes shallow soil moisture, turn causing tree roots togrow deeper to meet water requirements. As discussed earlier, the success of a phyto-cover is dependent upon establishment of deep-rooted trees to create a sufficientsponge to store soil moisture in the dormant season.



The trees normally selected for construction of a phyto-cover are hybrid poplarsof the variety

Deltoides

x

nigra

.

10,13

The candidate varieties, DN-21, DN-34, OP 367and others, are planted based on demonstrated growth ability and hardiness in theenvironment. The poplars are installed as either rooted plants or whips at a densityof approximately 1200 trees per acre.

15

The rows are located by measurement andflagged, and the trees installed by a tractor and mechanical planter. These trees aretypically planted with an in-row spacing of 3 feet and a row spacing of 10 to 13feet. They are planted in rows positioned along the land elevation contours, perpen-dicular to slopes to aid in reducing sheet flow velocities and surface erosion.

7.8 COVER SYSTEM PERFORMANCE

The engineered phyto-cover system should be designed to meet the post-closureand remediation objectives established for any landfill site as specified below:

• Minimize infiltration of precipitation through the cover system into the waste toprotect groundwater quality at the site.

• Resist surface soil erosion by wind and precipitation.• Minimize long-term maintenance.• Protect human health and the environment.• Offer post-closure and future beneficial use.

The achievement of these objectives is outlined in this and subsequent sections.

7.8.1 Hydrologic Water Balance

The engineered phyto-cover system is designed to mature into a remedial systemthat exceeds the hydrologic performance of more conventional cover systems. How-ever, instead of acting as a constructed barrier layer, which diverts precipitation fromthe cover area as surface runoff or internal drainage, this system intercepts and usesthe water for plant growth. In other words, the engineered phyto-cover functions asa sponge and pump system, with the root zone acting as the sponge, and trees actingas the solar-driven pumps. In contrast to restrictive permeability barrier design, theengineered phyto-cover design involves the storage of free water in soil pores andthe extraction of stored water by the tree roots.

The effectiveness of engineered phyto-cover systems as landfill closure systemshas been demonstrated at sites in the U.S. At sites in various climates with engineeredand agronomically optimized growing conditions, rapidly growing poplar trees arecapable of transpiring all natural precipitation that infiltrates into a site. While theperformance of engineered phyto-cover systems has been demonstrated, a proventool to design phyto-covers is not available. Therefore, to support the design anddemonstrate the effective performance of phyto-cover systems, this section discussessome fundamental scientific methods of water balance analysis.

As discussed previously, the phyto-cover system utilizes specially selected treeswith a grass understory to optimize evapotranspiration and achieve the equivalent



performance of a conventional barrier cover system. This alternative landfill coversystem has been designed to minimize percolation to the waste by incorporating alandfill soil cover with sufficient evapotranspirative and water holding capacity tostore precipitation temporarily in the nongrowing season for subsequent evapotrans-piration by vegetation in the growing season. The two key design elements inengineering a phyto-cover system are 1) determining the thickness and materialcomposition of the soil cover system required to provide sufficient water storagecapacity; and 2) incorporating a supportive phyto-cover system to access water storedin the soil cover system for evapotranspiration to the atmosphere.

Moisture flow and moisture content in a landfill are extremely important to thedynamic processes of decomposition and potential leachate generation. The funda-mental means to assess the moisture conditions is through an evaluation of variousprocesses comprising a water mass balance. A water mass balance analysis is an“accounting” procedure for tracking the moisture inputs to storage and the moistureoutputs that influence the potential flux of water through the cover into the waste.The primary elements of a water mass balance include precipitation, surface runoff(R/O), potential evapotranspiration (PET), infiltration (I), soil moisture storage (ST),actual evapotranspiration (AET), and flux (or percolation) of water through thesystem. The water shedding efficiency of a cap is then derived by calculating thepercentage of flux relative to total precipitation. The phyto-cover system designconcept involves maximizing efficiency by optimizing ET, runoff, and soil moisturestorage to minimize infiltration, flux, and potential leachate generation. The waterbalance accounting for a phyto-cover can be summarized by the following equationand Figure 7.9:

Percolation = Precipitation – Runoff – Evapotranspiration – Moisture Storage(7.3)

The water mass balance processes within a landfill are typically evaluated usingthe hydrologic evaluation of landfill performance (HELP) model, developed by theWaterways Experiment Station.

17

The applicability of this model to design andevaluation of an engineered phyto-cover system has been reviewed, and it has beendetermined that the HELP model is inappropriate for this analysis because of severalcomputational deficiencies.

18,19

The HELP model was developed based on assump-tions pertaining to water management through low permeability soil covers withvegetative covers comprising short-rooted grasses. No opportunity exists for userinput of higher ET values more representative of plant species with significantlyhigher potential water uptake than the short grasses assumed by the HELP model.Therefore, the model significantly underestimates evapotranspiration from trees andother deeply rooted vegetation that are key elements of a phyto-cover system. Thisapplication limitation of the HELP model results in an overestimation of infiltrationand coincident underestimation of efficiency (overestimation of drainage) if themodel were to be applied to an evaluation of a phyto-cover system.

A detailed assessment of various computer models used for landfill cover designsduring the early phases of the alternative cover assessment program (ACAP) cameto similar conclusions.

11,12

Of the four codes tested, HELP was the most widely used



for landfill design, and the most user-friendly and highly dependable. HELP predic-tions consistently provided the highest estimates of drainage. Three concerns withHELP were 1) a nonrealistic response of increased drainage as available watercapacity increased, 2) insensitivity of drainage to thickness of the cover surfacelayer, and 3) consistent overprediction of drainage. EPIC was also relatively easyto use, but consistently underpredicted drainage in comparison to other codes. Thestudy suggested that Richards’ equation-based codes (HYDRUS–2D, UNSAT–H)were better able to capture the behavior of alternative landfill covers than simplewater balance codes such as HELP and EPIC.

Although the HELP model itself cannot accurately simulate the hydraulic effectsof an engineered phyto-cover system, the water balance method that is the funda-mental principle applied within the HELP model has been employed to evaluate theperformance of vegetative cover systems.

20

These same scientific principles areemployed to design and evaluate the performance of an engineered phyto-coversystem with a new software tool called PHYTOSOLV.

15,21

In using the water balancemethod, the first step is to acquire accurate precipitation records applicable to thesite and encompassing various extreme wet and dry periods. The second step is todetermine the quantity of surface water runoff and infiltration (which are functions

Figure 7.9

Diagram of a phyto-cover (modified from Licht, 1998).

Depth ofCapillaryZone Draw

RootDepth

Storage

Infiltration

Soil Evaporation

Surface Evaporation

Leaf Transpiration

Surface EvaporationSurface CoverInterception

CanopyInterception

Precipitation

PotentialInfiltration

1) Surface Litter or Compost

2) Soil and Nonsoil Amendments

3) Waste Rootable UpperLayer (Contributes towardStorage)

Thickness of the Arrow isProportional to the Volume of Water



of the site soils, slope, and surface texture). Infiltration is computed as the differencebetween precipitation rates for the site and surface-water runoff from the soil cover.The third step is to apply PHYTOSOLV, assuming a variety of soil cover depths, togenerate a range of annual hydrologic water balances using daily precipitation data.Finally, a supporting phyto-cover system is designed that would access infiltratedsoil water throughout the entire extent of root growth (the “sponge”), and thenecessary evapotranspiration rate (the “pump”) required to deplete soil moistureduring the growing season is computed. This iterative water balance analysis is usedto select the appropriate soil cover design to best achieve desired hydraulic perfor-mance, thereby minimizing generation of leachate. The measure of performancefor the designed phyto-cover is compared to the water-shedding efficiency of tradi-tional barrier cover systems. Presented below is a discussion of each of these stepsand the basis for the general engineered phyto-cover system design.

7.8.2 Precipitation

Long-term precipitation data need to be assembled from the closest weatherstation to evaluate local hydrologic conditions. There are no established regulatorilyapproved procedures or protocols to evaluate the hydrologic performance of a phyto-cover design. Therefore, long-term data are needed in order to characterize the long-term precipitation trends and extremes. Typically, precipitation can vary widely fromsite to site for a given year, season, or month. To demonstrate this variability, datashould be assembled summarizing average monthly and annual precipitation fordecades at weather stations near any given site. For example, during a long periodat a site in Maryland, the average annual precipitation varied from a minimum of26.29 inches in 1965 to a maximum of 62.36 inches in 1996. Similar variability canalso be observed in monthly precipitation totals. To the extent practical, thesedynamics must be accounted for in the design of the phyto-cover system to demon-strate adequate performance under extreme weather conditions. The application ofthese data to evaluate the phyto-cover design assumes that daily precipitation totalsare the result of individual storm events.

7.8.3 Runoff

Runoff from the designed phyto-cover is computed using the USDA Soil Con-servation Service (SCS) curve number model.

22

The model computes direct runofffrom an individual storm event as a portion of total precipitation (Figure 7.10). Themethod was developed from field studies by measuring runoff from various soilcover, land slope, and soil type combinations. Curve numbers were developed basedupon each of the combined hydrologic effects of these factors and enable the modelto be applied to any area within the U.S. The curve number model is widely usedand is incorporated into the HELP model and other agronomic models to computerainfall runoff and other elements comprising a water balance. The major deficiencyin this model is that it underestimates runoff from small precipitation events. Thisdiscrepancy results in overestimates of infiltration and the amount of water that mustbe managed by the cover system.

3

Consequently, the resultant engineered



phyto-cover is overdesigned and conservative: the engineered phyto-cover has theability to control more infiltration than it is designed to manage.

15

(7.4)

where

Q = runoff (in)P = precipitation (in)S = potential maximum retention after runoff begins (in)I

a

= initial abstraction (in)

The initial abstraction is all water loss before runoff begins. It includes waterdetained in surface depressions, as well as water intercepted by vegetation, evaporation,and infiltration. The initial abstraction is highly variable but from data collected fromsmall agricultural watersheds, I

a

was approximated using the following equation:

I

a

= 0.2 S (7.5)

By eliminating I

a

as an independent parameter, this approximation allows use ofa combination of retention storage (S) and precipitation (P) to predict a unique runoffamount. Substituting Equation 7.5 into Equation 7.4 gives

(7.6)

Figure 7.10

Runoff — SCS Method.

Dire

ct R

unof

f (Q

), in

ches

8

7

6

5

4

3

2

1

00 1 2 3 4 5 6 7 8 9 10 11 12

Rainfall (P), inches

Curves on this sheet are forthe case Is = 0.2S, so that

Q = (P - 0.2s)2

P + 0.8S

Curve

Num

ber =

100

90

80

70

60

50

40

QP I

P I Sa

a

=−( )

−( ) +

2

QP S

P S=

−( )+

0 2

0 8

2.

.



where S is related to the soil and cover conditions of the watershed through thecurve number CN. CN has a range from 30 to 100, and is related to S by the followingequation:

(7.7)

The use of the SCS runoff equation for this analysis assumes that the differencebetween precipitation and runoff is infiltration.15

The curve number can be estimated by either using the HELP model or othercomputations. The HELP model computes a curve number based upon final grade,soil type, and vegetative cover. Using this model is recommended because it objec-tively estimates a curve number based upon the final design. In addition, the methodsin the HELP model were developed and approved by the USEPA. When using themodel, the minimum final grade should be specified for the land surface slope anda good vegetative cover be assumed for the understory. These two assumptionsensure that the selected curve number is conservative (minimize runoff/maximizeinfiltration).

7.8.4 Potential Evapotranspiration — Measured Data

Potential evapotranspiration (PET) is a measure of the maximum rate at whichevapotranspiration can occur when adequate soil moisture is available for utilizationby vegetation. These data are measured in the field utilizing lysimeters planted withsingle species covers (usually perennial grasses). Soil moisture levels are maintainedat optimum levels and evapotranspiration is measured by weighing the lysimeter.Data collected through these methods are the most reliable and most defendableestimates of potential evapotranspiration; however, measured site-specific data arenot readily available for most sites.

The monthly potential evapotranspiration rates measured for grasses are adjustedto best represent the supplemental evapotranspiration available from the trees. Thisstep is performed by incorporating a consumptive-use coefficient (Kc) applicable tothe trees utilized in the phyto-cover design. A consumptive-use coefficient of 1.35was measured for areas with cottonwood trees, willows, and grass.15,23 This value isnear the low end of the range for consumptive-use coefficient values derived fordensely planted trees in other parts of the U.S.; it has been reported that consumptive-use coefficient values for densely planted trees vary from 1.3 to 1.6.24 The cotton-wood is one species used to develop the hybrid poplar trees selected for the phyto-cover system. Hybrid poplar trees have been developed specifically to achieve a highconsumptive use coefficient, in addition to disease resistance and high growth rates;therefore, the selection of a consumptive-use coefficient of 1.35 is conservative forthis engineered phyto-cover system.

The consumptive-use coefficient is used to adjust the measured potential evapo-transpiration for grasses only during the growing season for trees — April throughOctober (Figures 7.11a and b). The growing season begins with approximately 10%

SCN

= −100010



leafing in April, 50% leafing in May, and 100% leafing in June, July, and August.Thereafter, leaf falloff occurs at a level of 70% in September and 20% in October.15

No leafing is assumed during the dormant season. Factoring these together, a com-bined consumptive-use coefficient for grasses and trees can be developed and usedto compute the monthly potential evapotranspiration rate for trees and grasses forthe phyto-cover design. It should be noted again, though, that these PET rates cannotbe achieved unless adequate soil moisture is available.

Figure 7.11a Measured potential ET for a site with cool wet climatic conditions.

Figure 7.11b Potential ET of phyto-cover at the same site adjusted for consumptive use.

Total 46.18 Inches

Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

10

9

8

7

6

5

4

3

2

1

0

Pot

entia

l Eva

potr

ansp

iratio

n (in

ches

)

Total 55.62 Inches

Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

10

9

8

7

6

5

4

3

2

1

0

Pot

entia

l Eva

potr

ansp

iratio

n (in

ches

)



7.8.5 Potential Evapotranspiration — Empirical Data

Modern equations to compute potential ET normalize estimated PET to a refer-ence crop evapotranspiration rate (Erc; mm/day). The reference crop ET rate isdefined as the rate of ET from an idealized grass crop with a fixed crop height of0.12 m, albedo of 0.23, and a surface resistance of 69 m/s. The reference crop closelyresembles an extensive surface of short green grass cover of uniform height that isactively growing, completely shading the ground, and not short of water. The nor-malization of PET rates facilitates the comparison of computed rates from differentmethods and equations. This also extends the application of crop specific consump-tive use coefficients for estimating PET between methods.

The potential evapotranspiration from a site can be estimated empirically utiliz-ing a radiation-based equation with the general form

(7.8)

where:

Erc = potential ET [mm/day]α = constant∆ = gradient of the saturation vapor curve as a function of temperture

(kPa/˚C)γ = psychrometric constant (kPa/˚C)Rn = net radiation exchange for the crop cover (mm/day)G = soil heat flux (mm/day)

Each of the constants in Equation 7.5 can be evaluated using average dailytemperature, mean elevation of the site above sea level, and atmospheric constants.Substantial evidence supports the application of Equation 7.8 for determining PETfor areas with uniform vegetation cover. The following radiation-based equationsfor estimation of potential evapotranspiration are recommended:15,25

(7.9)

for arid locations with relative humidity less than 60% in the month having peakevaporation, and

(7.10)

for all other humid locations.

E R Gr c n=+

−( )αγ

∆∆

E R Gr c n=+

−( )1 74.∆

∆ γ

E R Gr c n=+

−( )1 76.∆

∆ γ



The variables in Equations 7.9 and 7.10 are evaluated using the followingexpressions:

(7.11)

where

T = temperature (˚C)es = saturation vapor pressure (kPa)

(7.12)

γ = 0.0016286

P = atmospheric pressure (kPa) estimated from the site elevation using therelationship

P = 101.0 – 0.011 5E + 5.44 χ10–7 E2 (7.13)

E = land surface elevation (meters)λ = latent heat of evaporation of water (MJ/kg)

λ = 2.501 – 0.002361 T (7.14)

G = 0.38(Tday2 – Tday1) (7.15)

(7.16)

n = bright sunshine hours per dayN = total day length in hoursS0 = extraterrestrial radiation (MJ/m2/day)ed = vapor pressure (kPa) = relative humidity (fraction) times the es

(computed above)σ = Stefan-Boltzmann Constant (4.903 × 10–9 MJ/m2/˚K4/day)

7.8.6 Effective Evapotranspiration

The actual or effective ET is calculated by adjusting the PET value to accountfor the reduction in the ET rates as soil moisture is depleted. This adjustment isperformed using a standard model of ET as a function of soil moisture.15,26 Themodel is shown graphically in Figure 7.12. The effective ET rate occurs at the PET

∆ =+( )

4098

237 3 2

e

Ts

.

eTTs =

+( )

0 6108

17 27237 3

. exp..

Pλ

RnN

SnN

e Tn d= +

− +

−( )0 25 0 50 0 9 0 1 0 34 0 140

4. . . . . . σ



rate until soil moisture content is at a percentage of field capacity, then declineslinearly at soil moisture levels drier than this value until it is approximately zero atthe wilting point of the plants. Mathematically, the relationship between soil moistureis expressed as follows:

(7.17)

where

Er = PET when Θi ≤ Θ ≤ Θs (7.18)

(7.19)

Figure 7.12 Actual vs. potential evapotranspiration.

∂ ( )∂

= −Θ t

tEr

Et

PET whenri

w i= ( ) ≤ ≤ΘΘ

Θ Θ Θ



The application of Equations 7.17 through 7.19 to compute the change in soilmoisture between two time intervals requires the integration of these equations.Consider the time interval from 0 to a time t. The soil moisture at time t expressedas a function of the soil moisture at time 0 is as follows:

(7.20)

If the soil is wet (moisture content greater than Θi) the moisture content at timet (Θ2) is simply the initial moisture content (Θ1) minus the PET rate multiplied bythe time interval.

Θ2 = Θ1 – PET t (7.21)

If the soil is dry (moisture content less than Θi) the moisture content at time tis the initial moisture content minus the integrated ET rate over the time interval.

(7.22)

Equation 7.22 cannot be solved directly because there is no relationship for soilmoisture as a function of time.

The equation to characterize the change in moisture content when the soil is drycan be derived by substituting Equation 7.19 into Equation 7.17.

(7.23)

Separating variables and integrating Equation 7.23 between limits,

(7.24)

(7.25)

Substituting and exponentiation of both sides of Equation 7.22 yields

(7.26)

Θ Θt E dtr

t

( ) = ( ) − ∫00

Θ Θ ΘΘ

tt

PET dti

t

( ) = ( ) − ( )∫00

∂ ( )∂

= − ( )Θ ΘΘ

t

t

tPET

i

d t

tPET

dti

tΘ

Θ ΘΘ

Θ( )

( )= −∫ ∫

1

2

0

lnΘΘΘ

Θ

1

2

0

== PETt

i

t

ΘΘ Θ

2

1

= −

exp

PETt

i



The moisture content at time t (Θ2) can therefore be computed from the initialmoisture content (Θ1) using the formula:

(7.27)

Equations 7.21 and 7.27 describe the change in moisture content with time whenthe soil is dry or wet, respectively. A third case that needs to be considered is whenconditions change between wet and dry during the time period being evaluated (oneday for this model). Before Equation 7.21 is evaluated, the time necessary forconditions to change from wet to dry at current PET rates is computed with thefollowing equation:

(7.28)

If the time is greater than one day (the time period), Equation 7.21 is applied.If the time is less than one day, Equation 7.21 is applied for the computed time (td),and Equation 7.27 is applied for the remaining portion of the time period (1 – td)

7.8.7 Water Balance Model

Water balance models of the soil profile are based upon fundamental principlesof the behavior of water in the soil. During a storm event, surficial soils are saturatedby precipitation. Initially, water percolates vertically in the profile, redistributingmoisture until the remaining water held by surface tension on the soil particles isin equilibrium with gravitational forces causing drainage. The moisture content atwhich this equilibrium occurs is termed field capacity. Water uptake by plantscontinues to drive the drying process with little soil moisture restrictions untilmoisture contents reach 25 to 80% of field capacity. Transpiration rates decrease asthe soil continues to dry until the wilting point of the plants is reached. Furtherdeclines in soil moisture levels are controlled by the hydraulic conductivity of thesoil and occur as a result of evaporative and diffusion processes.25,27

These principles are used to develop a water balance model of an engineeredphyto-cover system. The water balance for a phyto-cover system begins with pre-cipitation. It is assumed for this analysis that all precipitation is in the form of rain;this assumption causes conservative design considerations related to cover thicknessand total water storage requirements. The soil surface separates precipitation intorunoff and infiltration. Runoff is estimated using the curve number model discussedpreviously. This procedure was selected because it consistently underpredicts runoffvolumes and adds an additional degree of conservatism in that infiltration is over-estimated. The runoff analysis for water balance assumes that daily precipitationtotals correspond to individual storms. After precipitation infiltrates the soil, wateris either stored, removed through ET, or, if moisture content is in excess of fieldcapacity, percolated through root zone and into the waste.

Θ ΘΘ2 1= −

exp

PETt

i

t PETdi=

−( )Θ Θ1



7.9 EXAMPLE APPLICATION

An example application was developed for a hypothetical site in Central Mary-land. Precipitation data were assembled from the Conowingo Dam weather stationto evaluate local climatic conditions. Average annual precipitation for the 37-yearrecord evaluated is 45.2 inches. A more detailed look at long-term precipitationtrends and totals indicates that precipitation can vary widely for a given year, season,or month. From 1960 through 1996 the average annual precipitation varied from aminimum of 26.29 inches in 1965 to a maximum of 62.36 inches in 1996. Similarvariability can also be observed in the monthly precipitation totals. To the extentpractical, this variability must be accounted for in the design of the phyto-coversystem to demonstrate adequate performance under extreme weather conditions.Therefore, the available daily precipitation data for the 37-year period between 1960and 1996 were assembled in order to design the phyto-cover system. Figure 7.13shows the daily precipitation data for a three-year period of the total years analyzed.Daily precipitation records for all 37 years are utilized in this analysis in order toaccount for rainfall runoff and extreme precipitation conditions.

Figure 7.13 Daily precipitation data for a three-year period for the example application; thisanalysis was performed for 37 years of data looked at for the site.

1/1/

60

4/23

/60

8/13

/60

12/3

/60

3/25

/61

7/15

/61

11/4

/61

2/24

/62

6/16

/62

10/6

/62

Date

0.00

1.00

2.00

3.00

Dai

ly P

reci

pita

tion

(in)



Potential ET rates are known for this area of the U.S. from data collected duringa 12-year lysimeter study in Seabrook, NJ.15,28 Actual ET rates are computed basedupon potential ET rates and soil moisture levels. For this example, the actual andpotential ET rates are assumed to be equal until moisture contents fall to less than25% of soil moisture storage between field capacity and wilting point. At lowermoisture contents, ET is assumed to decline linearly to zero at the wilting point.Other soil moisture models support the assumption that ET rates begin to declinewhen the moisture content is 25% of the difference between field capacity andwilting point (according to EPIC27 and CREAMS29) developed by the U.S.D.A.Therefore, the methodology and assumptions of this water balance analysis aretechnically defensible and comparable to the EPIC and CREAMS models developedby the USDA. An example of the expected daily performance of the final design ofthe engineered phyto-cover is shown in Figure 7.14. Figures that simulate daily soilwater storage, daily evaportranspiration, and observed precipitation from 1960through 1996 can be obtained through this analysis. If the total water holdingcapacity of the designed phyto-cover is more than the highest daily storage predicted,then the phyto-cover can be theoretically expected to prevent percolation.

The water mass balance analysis employing the data presented above was usedto determine the site-specific performance of various engineered phyto-cover sys-tems with different water holding capacities. This analysis is summarized in Figure7.15 and shows that the average annual percolate is a function of the total waterholding capacity of the cover design. The greater the total water holding capacityof the phyto-cover the less the average annual percolate.

Table 7.1 shows the designed water holding capacity of a phyto-cover utilizingexisting soil cover, supplemental imported soils, and municipal solid waste. Theexisting soil cover at the site is assumed to have a thickness of 1.0 foot. The water

Figure 7.14 Predicted daily performance of the final engineered phyto-cover design.

1/1/

604/

23/6

08/

13/6

012

/3/6

03/

25/6

17/

15/6

111

/4/6

12/

24/6

26/

16/6

210

/6/6

2

Dai

ly E

vapo

tran

spira

tion

(in)

1.00

0.90

0.80

0.70

0.60

0.50

0.40

0.30

0.20

0.10

0.00

Legend

Daily ET

Daily Storage

Dai

ly P

reci

pita

tion0.00

0.501.001.502.002.503.00

Soi

l Wat

er S

tora

ge (

in)

9.00

8.00

7.00

6.00

5.00

4.00

3.00

2.00

1.00

0.00

Date



holding capacity (WHC) for the soil type at the site (silt clay loam) is 1.3 inchesper foot.15,17 The supplemental soil cover has been assumed to be a silt loam witha water holding capacity of 1.8 inches per foot for this water balance analysis. Figure7.16 shows the water holding capacity of different types of soils.

The water holding capacity of waste in a mature landfill is between 0.5 and 1.7inches of water per foot depending on the percentage of municipal solid waste.3

Water holding capacity as high as 4.85 inches of water per foot of waste has beenreported. The landfill is assumed to consist primarily of municipal solid waste. TheHELP model reports that this type of waste has a water holding capacity of 2.58in/foot. Poplar trees grow vigorously over landfills, as evidenced by conditions at

Figure 7.15 Long-term performance as a function of total water holding capacity.

Table 7.1 Analysis of the Water Holding Capacity (WHC) of the Phyto-Cover “Sponge”

MaterialThickness

(feet)Field

CapacityWilting Point

WHC (in/foot)

Total WHC (inches)

Additional Imported Soil Cover 2.0 0.284 0.135 1.79 3.58Existing Landfill Soil Cover 1.0 0.244 0.136 1.30 1.30Root Growth in Landfill Matrix 5.0 0.292 0.077 2.58 12.90

Total 17.58

Dec

reas

ed L

each

ate

Pro

du

ctio

n

Ave

rage

Ann

ual F

lux

thro

ugh

Cov

er (

in/y

r)

100.00

10.00

1.00

0.10

0.01

Total Water Storage Capacity (in)

Increased Cover Thickness

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20

5.95

inch

es

10.0

0 in

ches

Existing Conditions (8.5 in/yr)

Standard Single-Barrier CoverCompacted Clay (1.09 in/yr)

Phyto-Cover, Total WaterStorage 10 inches (0.24 in/yr)



abandoned landfills, and routinely develop roots deeper than 8 feet below the soilsurface. Accordingly, the engineered phyto-cover system is designed to root into thewaste to capture additional water holding capacity.

According to Table 7.1, 17.78 inches is the total water holding capacity with anadditional 2 feet of soil cover, the existing average 1.0 feet of soil cover, and rootgrowth 5 feet into the landfill matrix. Therefore, the performance efficiency of thedesigned phyto-cover is greater than 98%; this level of efficiency favorably compareswith the performance efficiency of the RCRA cap over a long period of time.

7.10 SUMMARY OF PHYTO-COVER WATER BALANCE

After the phyto-cover is in place and functioning, the moisture levels in the wasteand soils below the root zone and above the water table will fall well below fieldcapacity as gravity and capillary action pull moisture out. These now-drier unsatur-ated materials will provide a water absorption capacity in addition to that within theroot zone to absorb potential infiltration below the root zone during occasionalextremely wet periods. As the overlying soils again fall below field capacity withadditional evapotranspiration, capillary action will act to remove moisture againfrom below the root zone and above the water table. Thus relatively small amountsof water moving through the root zone during infrequent periods of extreme precip-itation will not proceed directly to the water table, but will be held by the waste andunsaturated soils, and acted on by capillary action and osmotic potential to moveback up and into the atmosphere. In contrast, there is no upward gradient formovement of potential leakage from a membrane cap.

Figure 7.16 Water holding characteristics of soils.

Wat

er C

onte

nt in

Inch

es p

er F

oot o

f Soi

l4

3

2

1

0

Gravitational Water

Field Capacit

y

Water Readily Accessible to Plants

Water Accessible Only at R

educed Rate

Wilting Point

Limited Availability

San

d

Fin

eS

and

San

dyLo

am Fin

eS

andy

Loam

Loam S

iltLo

amLt

. Cla

yLo

am

Cla

yLo

am

Hv.

Cla

yLo

am

Cla

y



This design for the phyto-cover system incorporates a margin of safety becausethe water-balance calculations were based on the following conservative input val-ues:

• Underestimation of surface-water runoff by the curve number method (overesti-mation of infiltration)

• Assumption of a consistently short growing season• Consumptive-use coefficient lower than known measured values for the type of

trees in the design• Assumption of root growth depth shallower than that known to occur with the

type of trees in the design• Not accounting for additional storage capacity provided by the very large volume

of landfill matrix which will remain below field capacity most of the time, andthe subsequent capillary removal of moisture which may, on occasion, extend intothis mass

In addition to the use of these conservative design parameter inputs, the engi-neered phyto-cover system design, unlike the design of RCRA barrier caps, caninclude the actual measurement of soil moisture as an element of operations andmaintenance (O&M), thus ensuring continuous evaluation of the system’s hydraulicperformance.

This analysis demonstrates that the designed phyto-cover could perform at leastas effectively as a conventional landfill cover over a long period of time. In addition,once the phyto-cover is installed, the waste below the root zone will continuouslydry up and will encounter a significant loss in moisture content. This continuousdrying will provide additional water holding capacity to handle the flux associatedwith an extreme precipitation event. In summary, the analysis effectively demon-strates that the engineered phyto-cover system presented has the ability to achievethe objectives of conventional landfill covers and to protect human health and theenvironment, by protecting groundwater quality through minimizing the generationof landfill leachate. In addition, if coupled with monitored natural attenuation ofgroundwater, the engineered phyto-cover technology provides the optimum balancefor achieving all of the stated goals for landfill closure.

7.11 GENERAL PHYTO-COVER MAINTENANCE ACTIVITIES

7.11.1 Site Inspections

To ensure the continued proper functioning and integrity of the engineered phyto-cover system, regularly scheduled inspections should be conducted by on-site per-sonnel and the local expert. In addition, during the early part of the growing seasonor upon discovery of any unusual operational conditions, inspections by membersof the maintenance team should be completed. The local expert should have generalknowledge of erosion control, irrigation systems, soil moisture monitoring, andvegetation care. Inspections should be conducted during the growing season on aweekly schedule during the first year of the O&M period, biweekly during the second



year, and monthly or as needed thereafter throughout the O&M period. Detaileddescriptions of the activities specific to each time period are provided in latersections. During the first three years of the maintenance period, additional inspec-tions of the cover system should be conducted following storm events. Inspectionobservations should be recorded on an inspection log and stored on-site in a three-ring log binder. Photographic documentation of the growth and maturation of phyto-cover should also be collected and included with the inspection log.30

The log from each inspection should be reviewed by the maintenance team toevaluate phyto-cover system maturation and recommend any required correctiveaction. Expert personnel should be consulted, as necessary, to assist in the identifi-cation of diseases and pests that may affect the overall maintenance and health ofthe cover system and, in particular, the hybrid popular trees planted at the site.

To determine the condition of the phyto-cover system during routine inspections,the site operator should walk the perimeter and traverse the cover along a significantnumber of randomly selected, noncontiguous tree rows. Different rows should beselected each week, except when reinspection is necessary. Conditions that the siteoperator will inspect include, but are not limited to 1) surface disturbances or rutting;2) gullies, washouts, or other cover disturbances caused by water erosion; 3) settle-ment/subsidence; 4) wear, such as burrows due to animals; 5) vegetation condition;6) evidence of malfunctions within the irrigation system; 7) saturated soils andprecipitation ponding; 8) insects or pests; 9) tensiometer readings; 10) irrigationtotalizer readings; and 11) structural integrity of the perimeter fencing.

Immediate restorative actions are required if any of the following conditions areobserved during a visual inspection:

• Evidence of stressed vegetation, or evidence of animal, insect, disease, or otherdamage to vegetation

• Evidence of saturated soils, ponding water, or excessively dry soils• Erosion that could compromise the integrity of the phyto-cover system• Other conditions that may interfere with the proper performance of the phyto-

cover system

The site operator should document the inspection results and any observeddeficiencies on an inspection log, and corrective action should be undertaken anddocumented.

7.11.2 Soil Moisture Monitoring

As the phyto-cover matures, trees will develop roots through the surficial vege-tative soil zone and into the underlying waste.30 Initially, the majority of root growthwill be laterally along the surficial soil zone (the upper 24 inches). The cover usesroot uptake to “de-water” the root zone and underlying waste. By monitoring thesoil-moisture content in this zone, the site operator can qualitatively estimate waterremoval by plant uptake and moisture replenishment through irrigation and naturalprecipitation. Measurements will be made using two tensiometers placed at selectedlocations within the phyto-cover. Tensiometers will be placed at the beginning of



the growth season between trees in the rows, in areas of contrasting moisture contentbased on visual observation of the site after significant precipitation events.

A tensiometer is a simple device that measures the relative moisture of the soilby comparing uptake or loss of water through a ceramic membrane and recordingthe change in vacuum (Figure 7.17). The tensiometer data will be used along withsite-specific precipitation data to determine the rate and frequency of irrigationapplications as discussed below. Tensiometers must be removed and stored in accor-dance with the manufacturer’s instructions prior to the onset of freezing conditions.

7.11.2.1 Drainage Measurement

While many studies have documented parameters related to phyto-cover perfor-mance (soil-moisture content, precipitation, runoff), these measurements by them-selves do not address the central issue, namely the actual deep percolation throughthe cover. In most cases, the collection of soil moisture, runoff, and precipitationdata has been performed to meet regulatory performance requirements. Methodsutilizing these data to estimate the ability of a cover design to limit the flux of waterhave to rely on predictive methods and thus have inherent uncertainties.11,12

Methods of determining deep percolation include those based on fixed fractionsof annual precipitation, groundwater quality and level changes within the footprintof the landfill, water balance models, environmental tracer models, and lysimetry.11,12

Water-balance lysimetry is the most direct, quantitative method utilized to directlymeasure deep percolation through engineered phyto-covers. It has not been useduniversally due to the high cost of installation.

Figure 7.17 Photograph of a field tensiometer.



Water-balance lysimeters are typically soil-filled containers buried flush with thesoil surface that allow for some method of collecting drainage. Often these lysimeterscan also be used to measure water storage either directly through weighing or byusing independent methods (e.g., neutron probe or TDR) to monitor soil waterstorage. Lysimeters are semipermanent structures that are often expensive to con-struct and must be designed to account for soil physical properties and site-specificenvironmental variables. Despite these concerns, it was concluded that lysimetersprovide the most reliable means of measuring deep percolation, given adequatesurface area and long-term monitoring, with precision in drainage “often better than1 mm/year.”11,12

Measurement of percolation by drainage lysimetry is made possible by thepresence of an impermeable geomembrane forming the bottom boundary. The inter-ruption of downward movement of moisture by the geomembrane causes increasedmoisture content in the soil layer immediately above the membrane. Drainage occurswhen the soil above the membrane liner reaches near-saturation point. This require-ment presents a design problem if roots from surface vegetation are allowed topenetrate to the bottom of the lysimeter. When the downward flux of moisture isimpeded by an impermeable membrane and plants are allowed access to the trappedmoisture, percolation is reduced, which can result in false negatives. This factor canbe addressed if the lysimeter is relatively deep and roots are restrained fromapproaching the bottom liner of the lysimeter. Of all currently available methods,only the use of water-balance lysimeters (over several years, combined with climaticobservations, plant community activities, and soil parameters) can provide the datanecessary to quantify the performance of phyto-covers and validate and calibratenumerical models used for design purposes.19 Many different configurations ofwater-balance lysimeters are shown in Figures 7.18a through h.

Figures 7.18a Detail of proposed lysimeter test facility design (courtesy of Rock and USEPA,1999).

20 m Natural Slope

60-milHDPE Liner

GeosyntheticRoot Barrier

GeocompositeDrainage Layer

Cover Materials:Variable Depth

Surface FlowDiversion

Interim Cover: Variable Depth

3 to 5% Slope

French Drain,Sump Pump

Electronic Measurementof Runoff and Drainage

Manhole10 m



7.11.3 General Irrigation Guidelines

The phyto-cover system will require intermittent irrigation from approximatelyMay through September, depending on observed weather conditions and year ofoperation. An existing site precipitation gauge should be utilized to measure precip-itation received at the site on a weekly basis at a minimum. If precipitation exceeds

Figures 7.18b Lysimeter facility at INEEL EBTF (courtesy of Rock and USEPA, 1999).

Figures 7.18c Live oak (Atlanta) lysimeter (courtesy of Rock and USEPA, 1999).

Measurements

DrainagePrecipitationSoil Moisture ContentSoil Moisture Potential

Lysimeters: 2 rows of 5 eachDimensions:

Construction: concrete walls and floor

3m wide3m long3m deep

Drainage

Access Trench

Berm

Measurements

Soil MoistureDrainagePrecipitationAir TemperatureRelative HumiditySolar RadiationWind SpeedRunoffDew Point

37% slope

RunoffCollection

Drainage

60-mil HDPE Liner

Diversion Berms

Lysimeter Dimension: 12 m x 18



0.25 inches in a 24-hour period, or a total of 1 inch in any 1-week period, theirrigation system should be turned off at the backflow prevention valve to preventoverwatering of the phyto-cover system and saturation of the soils. The precipitationgauge information should be augmented by the tensiometer data collected at thephyto-cover. The tensiometer readings should be used along with the tree indicatorparameters described later to determine if the precipitation events have providedsufficient soil moisture or if augmentation through irrigation is required. The coversystem soils must also be visually inspected prior to reactivation of the irrigationsystem to determine soil moisture conditions. If either saturated soils or standingwater is observed, irrigation of the engineered phyto-cover system should not beconducted until the affected surface is dry, standing water is absent, and thetensiometer readings indicate a reduction in soil moisture to less than the assumed

Figures 7.18d Omega Hills lysimeter facility (courtesy of Rock and USEPA, 1999).

Figures 7.18e Twentynine Palms lysimeter facility (courtesy of Rock and USEPA, 1999).

Measurements

Soil MoistureDrainagePrecipitationSoil TemperatureRunoffVegetationActivities

DrainageCollection

Collection Sump

Sand Bedding

Drainage Grid andHypalon Membrane

33% Slope

RunoffCollection

CoverMaterials

Cover SoilTDR

40-mil HDPE

BermDrainageCollection

Collection Pipe

Measurements

Soil Moisture

3% Slope



field capacity (+/– 30 centibars). Again, indicator parameters should be the primarymethod used to determine when to resume irrigation. The irrigation system zonecapabilities make it possible to irrigate discrete areas of the cover in the event thatsignificant areal variability in soil moisture is observed. An example of an irrigationsystem is shown in Figure 7.19.

Figures 7.18f Rocky Mountain Arsenal lysimeter facility (courtesy of Rock and USEPA, 1999).

Figures 7.18g Hill AFB lysimeter facility (courtesy of Rock and USEPA, 1999).

Measurements

DrainagePrecipitationSoil MoistureIrrigationRunoff

CoverMaterials Drainage Geocomposite

DrainageCollection

60-mil VFPE Geomembrane

3% Slope

Overall Dimensions: 15.2 m long by 9.1 m wide

Measurements

DrainagePrecipitationSoil MoistureSoil Temperature

Air TemperatureSoil Heat FluxNet RadiationSnow Depth

Cover Soils

Instrument Access

20-mil Reinforced Geomembrane

Size: 4.6 m wide x 11.0 m long

DrainageFiberglass _______ 15 cm PVC Pipe



For the first 3 years, hybrid polar trees planted at the site will require approxi-mately 1 inch of water per week to promote rapid growth and development of ahealthy root system. This water can be from natural precipitation or through irriga-tion. The ideal distribution of any required irrigation water would be in daily events,irrigating at night to reduce evaporative losses. The drip irrigation system will beadjusted by the site operator to assure that trees receive the appropriate amount ofwater. This is very important because insufficient water will cause trees to wilt andexcess watering will stunt growth and possibly cause trees to die from lack of oxygenat the root zone.

Should natural precipitation occur coincident with irrigation, the irrigation shouldcontinue until either 1 inch of precipitation is received within 7 days, or if any single

Figures 7.18h Detail of a vadose zone monitoring station (courtesy of Rock and USEPA,1999).

Figure 7.19 An example of an irrigation system for a phyto-cover.

CoverMaterials

MinirhizotronAccess Tube

Interim Cover

Soil Water Content Soil Water Pressure Soil Temperature

Datalogger



daily precipitation event exceeds 0.25 inches.30 In the event that irrigation is stoppeddue to the 0.25-inch standard being met, then irrigation should continue again within2 days of the event, unless soils are visually saturated or tensiometer readings indicateless than 30 centibars vacuum.

The irrigation system must be turned off and winterized during the dormantseason for trees; generally, this occurs in mid- to late September. Soil-moisturemonitoring and inspection of the engineering phyto-cover system should beconducted at the beginning of the next growing year to verify soil moisture contentbefore the drip irrigation system is reactivated.

7.11.4 Tree Evaluation

The most important and effective indication of maturation and health of the treepopulation is the presence of abundant, large leaves and an increase in both stemheight and girth. Most woody plants grow intermittently, and it is not unusual to seegrowth in what are known as “flushes” when environmental factors are favorable,and slower growth when unfavorable conditions exist. The following field evaluationtechniques will assist in the evaluation of tree performance.

7.11.4.1 Stem

The woody stems of one-year-old rooted stock poplar trees at planting have anaverage diameter of less than one half inch at planting. During the first year ofgrowth, it is not unusual for the diameter to double; This increase in girth is referredto as secondary growth. In addition to the increase in girth, the height of the treemay increase by an average of 25% or more. The increase in length is known asprimary growth (Figure 7.20).

In addition to increases in primary and secondary growth, the stem will formbranches to expose as many leaves as possible for the production of energy. It is notunusual for poplar species present at the site to form branches all along the stem duringthe first year of growth. As primary growth increases in later years, these branches aregenerally lost through the process of abscission, and replaced by branches higher upthe stem at the tree crown. Abscission is a change in plant hormonal conditions thatcauses the plant to shed leaves and small limbs, while at the same time generatingprotective materials (cork) to cover the wound or abscission site.

A few simple field tests can be completed during the early years to assess stemcondition. The simplest method involves bending the top of the stem, which shouldbe supple and flexible and not crack or break. A second test involves scratching barkfrom the stem to reveal the cambium. This test should be performed if the bendingtest indicates problems with the meristematic tissue. If the cambium is green, thenthe meristematic tissue is healthy, and the stem is healthy. To assess the cambium,the stem should be scratched beginning at the top and working down to the base ofthe tree, as die-back in the stem generally occurs from the extremities back tothe roots.



7.11.4.2 Leaves

The leaves of trees serve to convert raw materials (sunlight, water, CO2, etc.)into sugars and control the loss of water by evapotranspiration. The leaves arecomposed of several different cell types, but the most important structures affectingperformance of the phyto-cover system are the cuticle-covered epidermis and thestomata located on the underside of the leaves. The cuticle is a waxy coating thatlimits secondary or cuticle transpiration. The stomata are the primary means oftranspiration, and account for more than 90% of transpirative losses. Indicators tocheck include uniformity in leaf color and growth.

Poplar tree leaves form from buds on the stem and branches. Healthy poplarleaves form large delta-shaped leaves with a diameter that can grow in excess of sixinches at the widest part, with lengths in excess of six inches as well (Figure 7.21a).The leaves are a brilliant green color with variations of darker green also present;They should not contain black or brown areas of discoloration. The leaves areattached to the stem by a flat-shaped petiole.

Leaves are the primary indication of stress or disease (Figure 7.21b). Theyindicate stress from a variety of environmental conditions including drought, over-watering, lack of oxygen or excessive CO2 at the roots, and animal foraging andinfestation. Drought conditions are evidenced by withering and, if severe enough,by early abscission, whereas other stresses can cause the leaves to change color andwither on the stem without falling. Infestation and foraging are more readily diag-nosed by physical evidence associated with discoloration or the loss of part of anotherwise healthy leaf. Visible evidence of leaf stress must be immediately addressed,as certain conditions can cause rapid declines in health and survivability of the tree.

Figure 7.20 Healthy stem of a growing poplar tree.



7.11.5 Agronomic Chemistry Sampling

Soil and foliage samples must be collected in August during the first 3 years ofoperations, and in subsequent years as the need may arise to diagnose growthproblems or apparent disease. Nutrient analyses of soils and leaves must be used todetermine the fertilizer addition necessary, including organic content analyses(American Society for Testing and Materials [ASTM] D-2974) and pH (ASTM D-4972), among others (e.g., nitrogen, phosphorous, etc.). Additional testing must beperformed as necessary to diagnose performance problems identified during inspec-tions, and assess the results of any corrective action taken.

Figure 7.21 (a) Healthy and (b) diseased leaves of a poplar tree.



The phyto-cover system must be divided into representative sections by assigningtransects perpendicular to the tree rows. A reasonable number of soil subsamplesacross each transect must be collected and combined into one sample. Soils can becollected within 18 inches of trees, and must be collected from a depth of approx-imately 12 inches below grade. The samples must be analyzed for pH and essentialmicronutrients as discussed above. Results should be reviewed and compared tosuggested concentrations provided in the environmental soil analysis report suppliedby the agronomic testing laboratory to determine necessary fertilizer addition or pHadjustments at the site. Laboratory recommendations for fertilizing deciduous treesshould be used to determine the appropriate fertilization rate.

In addition to soil samples, samples of leaf tissue must be collected from thesame transects. For each transect, a total of 20 leaves (or the minimum numberrequired by the laboratory, if higher) must be collected from 20 trees. It is suggestedthat samples be collected from healthy, recently matured leaves. These samples mustbe composited and analyzed for 13 nutrient elements.

7.11.6 Safety and Preventative Maintenance

During maintenance, all fertilizers, herbicides, and insecticides must be broughton-site in ready-to-apply concentrations. No mixing of undiluted applications shouldbe permitted, and all chemicals brought on-site must be approved by the site operatorand accompanied by appropriate material safety data sheets. All chemicals broughton-site must be documented to have no deleterious effect on the trees. If required bylocal statute, all herbicides and insecticides must be applied by a licensed contractor.

7.11.7 Repairs and Maintenance

Based upon the outcome of the visual observations, the inspection logs mustaddress, as needed, repair and maintenance of security control devices, vegetation,erosion of the phyto-cover system, surficial settlement, and irrigation system. If aproblem is identified during the course of site inspections, the situation should beevaluated by the site operator and necessary responses must be initiated in a timeframe appropriate to the condition. After the repair, a reinspection must be madeand documented on an inspection log. If, upon reinspection, it is determined thatadditional corrective maintenance is necessary, prompt attention must be given tothe deficiency.

7.12 OPERATION AND MAINTENANCE (O&M) SCHEDULE

The O&M associated with the phyto-cover is divided into three distinct phases:year 1 — establishment; years 2 and 3 — active maintenance; and year 4 andsucceeding years — passive maintenance. The site operator or assignee must performmaintenance activities, and record observations on the inspection log form. Eachmaintenance period is discussed in detail in the following sections.



7.12.1 Year 1 — Establishment

Management tasks for year 1— Establishment include the following activities:

1. The inspector(s) will conduct weekly site inspections of the entire cover systemduring the growing season and after major storm events. Biweekly inspectionswill be conducted during the dormant season.

2. Periodic mowing will be conducted to maintain the understory health and reduceunderstory competition with the trees. It is anticipated that monthly mowing betweenthe tree rows during the growing season will be sufficient. Care must be exercisedso that irrigation drip lines are not damaged during mowing. Mowing between treeswithin the rows is not necessary and will result in damage to the drip piping.

3. Herbicides will be applied by an appropriate contractor, as required, to preventthe propagation of undesirable species. The selected herbicide must not have anyeffect on the poplar trees.

4. Insecticide will be applied by an appropriate contractor, as required, based on inspec-tions for infestations of cottonwood beetles, eastern tent caterpillars, or other peststhat pose potential problems. The selected herbicide will have no effect on the trees.

5. Monitoring will include foliar and soil sampling during August of the first year;results will be analyzed for year 2 fertilization requirements. Foliar and soilsampling will be conducted in accordance with the methods of the AgriculturalResearch Service, USDA, to determine soil fertility status as it affects foliationand plant rooting patterns.

6. Repair or replacement of erosion control and security features will be imple-mented, as required, by an appropriate contractor. This will include surface appli-cation of compost mulch, straw, or shredded yard debris.

7. Irrigation system inspection and maintenance will be conducted, including:• inspecting and maintaining the backflow prevention valve, supply manifold,

zone valves, and drip lines; and,• winterization of the drip system, including turning off water supply valving,

blowing out the lines with air, and removal of batteries from the zone valvesat the end of the irrigation season.

7.12.2 Years 2 and 3 — Active Maintenance

The following tasks will be performed for years 2 and 3 — Active Maintenancefor the phyto-cover system:

1. The inspector will conduct biweekly site inspections of the entire cover system,year round and after major storm events.

2. The inspector will contract to apply fertilizer in April, as recommended by agro-nomic analyses from the year 1 August foliar and soil sampling.

3. The inspector will supply weed control on an as-needed basis. Less emphasisshould be required in year 2, and weed control should only be applied if specificstress indicators that can be attributed to weeds are present on the trees.

4. Mowing will take place as required to maintain the understory health.5. Dead or diseased trees must be removed and replaced with a healthy tree.6. Insecticides may be required. Spraying leaf surfaces is especially important for

an infestation of cottonwood beetle or eastern tent caterpillar, should it occur.



Such insect infestations are normally a more significant problem during the first2 years of establishment.

7. Monitoring will include August foliar and soil sampling as in year 1 for nutrientcontent. Results will be analyzed to determine the following year’s spring fertil-ization requirements.

8. Repair or replacement of erosion control structures will be implemented asrequired.

9. Irrigation system inspection and maintenance will be conducted, including:• inspecting and maintaining the backflow prevention valve, supply manifold,

zone valves, and drip lines; and,• winterization of the drip system, including turning off water supply valving,

blowing out the lines with air, and removal of batteries from the zone valvesat the end of the irrigation season.

7.12.3 Year 4 — Passive Maintenance

Passive maintenance of the phyto-cover system will be conducted beginning inyear 4 and continuing for succeeding years. These tasks will include the followingactivities:

1. Monthly site inspections will be conducted. More frequent inspections will becompleted during the spring leafing and fall bud setting to evaluate the tree growth,at the discretion of the site operator.

2. Fertilization will be applied, as required, based on the previous year’s agronomicanalyses.

3. Weed control will not be required unless undesirable species present a threat tothe trees.

4. Dead or diseased trees must be removed and replaced with a healthy tree.5. Insecticides will be applied only if necessary.6. August foliar and soil sampling may not be required, but nutrients may be applied

periodically in following years to maintain the best growth at the discretion ofthe site operator.

7. Repair or replacement of erosion control features will be implemented as required.8. Irrigation system operations should only be required during periods of unusually

dry weather, when precipitation amounts are lower than normal. Inspection andmaintenance of the irrigation system will be conducted on an annual basis, andbefore the system is used after a long period of inactivity, including:• clearing of potential sediment buildup in the drip lines; and,• maintaining the connection of the backflow prevention valve, manifold, zone

valves, and drip lines.

Over the life of the phyto-cover, indigenous species will be allowed to invadein several stages of forest community succession, leading to climax culture of amature forest predominant in the region. This is inherent in the design of the system,leading to a self-sustaining biotic community and providing an alternative to har-vesting and replanting. Furthermore, it allows the slower growing, longer lived treesalready populating the perimeter of the site to colonize into the poplar stand as thepoplar trees reach the end of their life expectancy of 50 or more years.



The protection of human health and the environment provided by the phyto-cover system, as designed, will be equivalent in performance to that of a single-barrier cover system. If it is determined that the invasion of indigenous species doesnot maintain the performance at an equal or better level of protection, these otherspecies will not be allowed to establish themselves and supplant the design species(poplars).

7.13 SPECIFIC OPERATIONAL ISSUES

7.13.1 Irrigation System Requirements

The irrigation system comprises several components that require routine evalu-ation and maintenance. The backflow prevention valve will require seasonal frostproofing. Prior to operating the valve in the spring, the bleed valves must be closed,and the unit checked for leaks. The irrigation manifold piping will also requireseasonal frost proofing. After closing and frost proofing the backflow preventionvalve, the manifold should be evacuated using compressed air.

In addition to frost-proofing requirements, the zone control valves require routinemaintenance. The valves should be monitored during weekly inspections to ensureproper operation. This can be accomplished by opening and closing the valves withthe programmable controller. The valves can be operated by nine-volt batteries. Morefrequent battery changeouts may be required depending upon the programmed irri-gation cycle frequency.

7.13.2 Tree Replacement

Should the replacement of trees be required, the optimal planting time would bethe end of March or beginning of April. Trees can also be planted in the fall; however,this is not recommended for bare-root stock. One-year-old rooted stock of the samespecies and preferably from the same supplier should be obtained and plantedimmediately upon receipt. Trees should be planted in the same location as the treebeing replaced. To plant the trees, dig or drill a 6-inch diameter or larger, 18- to 24-inch-deep hole and place the root mass directly into the hole. Carefully backfill thehole with the soils removed during excavation or with a good quality organic humus.Compact the soils around the tree carefully by hand, and water each tree with atleast ten gallons of water. Place a bamboo stake and protective cover around thetree, and replace the drip irrigation pipe so that the drip emitter is within 6 inchesof the tree stem.

REFERENCES

1. USEPA, Final Covers on Hazardous Waste Landfills and Surface Impoundments,Technical Guidance Document, Office of Solid Waste and Emergency Response,Washington, D.C., 1989.



2. Rock, S. A. and P. G. Soyre, Phytoremediation of hazardous wastes: potential regu-latory acceptability,

Remed. J.

, Autumn, 1998.3. McBean, E. A., F. A. Rovers, and G. J. Farquhar,

Solid Waste Landfill Engineeringand Design

, Prentice Hall, Englewood Cliffs, NJ, 1995.4. Reinhart, D. R. and T. G. Townsend,

Landfill Bioreactor Design and Operation

, LewisPublishers, Boca Raton, FL, 1997.

5. Suter, G. W., R. J. Luxmoore, and E. D. Smith, Compacted soil barriers at abandonedlandfill sites are likely to fail in the long term,

J. Environ. Qual.

, 22, 217–226, 1993.6. Koerner, R. M. and D. E. Daniel,

Final Covers for Solid Waste Landfills and Aban-doned Dumps

, ASCE Press, Reston, VA, 1997.7. Board, M. and D. Laine, Corralling liner nightmares,

MSW Manage.

, 5, 48–51, 1995.8. Crozier, F. and T. Walker, How much does your liner leak?,

Wastes Manage.

, 24–26,1995.

9. Finn, S., Golder Associates, personal communication, 1998.10. Licht, L., Ecolotree, Inc., personal communications, 1996, 1997, 1998.11. USEPA,

Alternative Cover Assessment Project (ACAP), Phase I Report

, prepared byWater Resources Center, Desert Research Institute, August, 1999.

12. Rock, S., personal communications, 1997, 1998, 1999, 2000.13. Gatliff, E., personal communication, 2000.14. Brady, N. C.,

The Nature and Properties of Soils

, MacMillan Publishing Company,New York, 1992.

15. Potter, S., ARCADIS G & M, Inc., personal communications, 1997, 1998, 1999, 2000.16. Bent, Tim, Bridgestone-Firestone, Inc., personal communication, 1997.17. Schroeder, D. R. et al.,

The Hydrologic Evaluation of Landfill Performance (HELP)Model: User’s Guide for Version 3

, EPA/600/9-94/1689, USEPA Risk ReductionEngineering Laboratory, Cincinnati, OH, 1994.

18. Ecolotree, Inc. and CH

2

H-Hill, Inc.,

Ecolotree Cap System; Focused Feasibility Studyfor Selected Virginia Landfills

, April 22, 1996.19. Khire, M. V., C. H. Benson, and P. J. Bosschur, Water balance modeling of earthen

final covers,

J. Geotech. Geoenviron. Eng.

, 123, 744–754, 1997.20. Anderson, J. E.,

Soil Plant Cover Systems for Final Closure of Solid Waste Landfillsin Arid Regions

, in Landfill Capping in the Semi-Arid West: Problems, Perspectives,and Solutions, Environmental Science and Research Foundation, Idaho Falls, ID,1997, 27–38.

21. ARCADIS G & M, Inc., Proprietary Software

PHYTOSOLVE for Phyto Cover Eval-uation, 1998.

22. Soil Conservation Service,

Urban Hydrology for Small Watersheds

, Technical Release55, 1986.

23. Schultz, E. F.,

Problems in Applied Hydrology

, Water Resource Publications, FortCollins, CO, 1973.

24. Stewart, B. A. and D. R. Nielson,

Irrigation of Agricultural Crops

, Agronomy No.30, American Society of Agronomy and Soul Science Society of America, Madison,WI, 1990.

25. Maidment, D. R., Ed.,

Handbook of Hydrology

, McGraw-Hill, Inc., New York, 1993.26. Fedder, R. A., P. J. Kowalrik, and H. Zaradyn,

Simulation of Field Water Use andCrop Yield

, Centre for Agricultural Publishing and Documentation, Wageningen,Netherlands, 1978.

27. Sharpley, A. N. and J. R. Williams, Eds.,

Erosion/Productivity Impact Calculator: 1.Model Documentation

, Technical Bulletin, 1768, U.S. Dept. Agric., Washington,D.C., 1990.



28. American Society of Civil Engineers (ASCE), Evapotranspiration and IrrigationWater Requirements, ASCE Press, New York, 1990.

29. Knisel, W. G. Ed., CREAMS: A Field Scale Model for Chemicals Runoff and Erosionfrom Agricultural Management Systems, U.S. Dept. Agric., Conservation ResearchReport, No. 26, 345, 1980.

30. Hannum, E., ARCADIS G & M, Inc., personal communications, 1999, 2000.


365

APPENDIX

A

Physical Properties of Some CommonEnvironmental Contaminants

L1282/Appendix A/frame Page 365 Monday, June 18, 2001 9:16 AM


Tab

le A

1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3

/mo

lV

apo

r P

ress

ure

m

m H

g.

So

lub

ility

mg

/LL

og

K

oc

A

Ace

naph

then

e15

4.21

7.92

×

10

–5

(25

°C)

0.00

155

(25°

C)

3.47

(25

°C)

1.25

Ace

naph

thyl

ene

152.

202.

8

×

10

–4

0.02

90 (

20°C

) 3.

93 (

25°C

)3.

68A

ceto

ne58

.08

3.97

×

10

–5

(25

°C)

266

(25°

C)

Mis

cibl

e–0

.43

Acr

olei

n56

.06

4.4

×

10

–6

(25

°C)

265

(25°

C)

200,

000

(25°

C)

–0.2

8A

cryl

onitr

ile53

.06

1.10

×

10

–4

(25

°C)

110-

115

(25°

C)

80,0

00 (

25°C

)–1

.13

Ald

rin36

4.92

4.96

×

10

–4

6

×

10

–6

(25

°C)

0.01

1 (2

5°C

)2.

61A

nthr

acen

e17

8.24

6.51

×

10

–5

(25

°C)

1.95

x 1

0

–4

(25

°C)

0.07

5 (2

5°C

)4.

41A

ceta

ldeh

yde

44.0

56.

61

×

10

–5

(25

°C)

760

(20.

2°C

)M

isci

ble

Una

vaila

ble

Ace

tic A

cid

60.0

51.

23

×

10

–3

(25

°C)

11.4

(20

°C)

Mis

cibl

eU

nava

ilabl

eA

cetic

Anh

ydrid

e10

2.09

3.92

×

10

–6

(20

°C)

5 (2

5°C

)12

% b

y w

t. (2

0°C

)U

nava

ilabl

eA

ceto

nitr

ile41

.05

3.46

×

10

–6

(25

°C)

73 (

20°)

Mis

cibl

e0.

342-

Ace

tyla

min

ofluo

rene

223

.27

——

—3.

20A

cryl

amid

e71

.08

3.03

×

10

–3

(20

°C)

7

×

10

–3

(20

°C)

2.15

5 g/

L (3

0°C

)U

nava

ilabl

eA

llyl A

lcoh

ol58

.08

5.00

×

10

–6

(25

°C)

20 (

20°C

)M

isci

ble

0.5

1A

llyl C

hlor

ide

76.5

31.

08

×

10

–2

(25

°C)

360

(25°

C)

—1.

68A

llyl G

lyci

dyl E

the

114.

143.

83

×

10

–6

(20

°C)

3.6

(20°

C)

141

g/L

Una

vaila

ble

4-A

min

obip

heny

l16

9.23

3.89

×

10

–10

(25

°C)

6 x

10

–5

(20

–30°

C)

842

(20–

30°C

)2.

032-

Am

inop

yrid

ine

94

.12

—Lo

w10

0 w

t. %

at

(20°

C)

Una

vaila

ble

Am

mon

ia17

.04

2.91

×

10

–4

(20

°C)

10 a

tm

(25.

7°C

)53

1g/L

(20

°C)

0.49

n

-Am

yl A

ceta

te13

0.19

3.88

×

10

–4

(25

°C)

4.1

(25°

C)

1.8

g/L

(20°

C)

Una

vaila

ble

sec

-Am

yl A

ceta

te13

0.19

4.87

×

10

–4

(20

°C)

10 (

35.2

°C)

0.2

wt.

% (

20°C

Una

vaila

ble

Ani

line

93.1

30.

136

(25°

C)

0.6

(20°

C)

—1.

41

o

-Ani

sidi

ne12

3.15

1.25

×

10

–6

(25

°C)

0.1

(30°

C)

1.3

wt.

% (

20°C

)U

nava

ilabl

e

p

-Ani

sidi

ne23

.15

——

3.3

(Roo

m T

emp)

Una

vaila

ble

Ant

u20

2.27

—

≈

0 (

20°C

)60

0 (2

0°C

)—


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 367

B

Ben

zene

78.1

10.

0054

8 (2

5°C

)95

.2 (

25°C

)18

00 (

25°C

)1.

92B

enzi

dine

84.2

43.

88

×

10

–11

(25

°C)

0.83

(20

°C)

500

(25°

C)

.60

Ben

zo[a

]ant

hrac

ene

228.

308.

0

×

10

–6

1.1

×

10

–7

(25

°C)

0.01

4 (2

5°C

)6.

14B

enzo

[b]fl

uora

nthe

ne25

2.32

1.2

×

10

–5

(20

-25°

C)

5

×

10

–7

(20

°C)

0.00

12 (

25°C

)5.

74B

enzo

[k]fl

uora

nthe

ne25

2.32

0.00

104

9.59

×

10

–11

(25

°C)

0.00

055

(25°

C)

6.64

Ben

zoic

Aci

d12

2.12

7.02

×

10

–8

0.00

45 (

25°C

)34

00 (

25°C

)1.

48–2

.70

Ben

zo[g

hi]p

erle

ne27

6.34

1.4

×

10

–7

(25

°C)

1.01

×

10

–10

(25

°C)

0.00

026

(25°

C)

6.89

Ben

zo[a

]pyr

ene

252.

32<

2.4

×

10

–6

5.6

×

10

–9

(25

°C)

0.00

38 (

25°C

)5.

60–6

.29

Ben

zyl A

lcoh

ol10

8.14

Insu

ffici

ent

vapo

r pr

essu

re d

ata

for

calc

ulat

ion

at 2

5°C

1 (5

8°C

)42

,900

(25

°C)

1.98

Ben

zyl B

utyl

Pht

hala

te31

2.37

1.3

×

10

–6

(25

°C)

8.6

×

10

–6

(20

°C)

42.2

(25

°C)

.83–

2.54

α

-Bhc

290.

835.

3

×

10

–6

(20

°C)

2.5

×

10

–5

(20

°C)

2.0

(25°

C)

3.27

9

β

-Bhc

290.

832.

3

×

10

–7

(20

°C)

2.8

×

10

–7

(20

°C)

0.24

(25

°C)

3.55

3

δ

-Bhc

290.

832.

5

×

10

–7

(20

-25°

C)

.7

×

10

–5

(20

°C)

31.4

(25

°C)

3.27

9B

is(2

-chl

oroe

thox

y) M

etha

ne17

3.04

3.78

×

10

–7

1 (5

3°C

)81

,000

(25

°C)

2.06

Bis

(2-c

hlor

oeth

yl)

Eth

er14

3.01

1.3

×

10

–5

1.55

(25

°C)

10,2

00 (

25°C

)1.

15B

is(2

-chl

oroi

sopr

opyl

) E

ther

171.

071.

1

×

10

–4

0.85

(20

°C)

1700

(20

°C)

1.79

Bis

(2-E

thyl

hexy

l) P

htha

late

390.

571.

1

×

10

–5

(25

°C)

6.2

×

10

–8

(25

°C)

0.4

(25°

C)

5.0

Bro

mod

ichl

orom

etha

ne16

3.83

2.12

×

10

–4

50 (

20°C

)45

00 (

0°C

)1.

79B

rom

ofor

m25

2.73

5.6

×

10

–4

5.6

(25°

C)

3.13

0 (2

5°C

)2.

454-

Bro

mop

heny

l Phe

nyl E

ther

249.

201.

0

×

10

–4

0.00

15 (

20°C

)N

o da

ta f

ound

4.94

2-B

utan

one

72.1

14.

66

×

10

–5

(25

°C)

100

(25.

0°C

)25

.57

wt.

% (

25°C

)0.

09B

enzo

[e]p

yren

e25

2.32

4.84

×

10

–7

(25

°C)

5.54

×

10

–9

(25

°C)

0.00

38 (

25°C

)5.

6B

enzy

l Chl

orid

e12

6.59

3.04

×

10

–4

(20

°C)

1 (2

2.0°

C)

493

(20°

C)

2.28

Bip

heny

l15

4.21

4.15

×

10

–4

(25

°C)

10

–2

(25

°C)

7.5

(25°

C)

3.71

Bro

mob

enze

ne15

7.01

2.4

×

10

–3

(25

°C)

4.14

(25

°C)

409

(25°

C)

2.33

Bro

moc

hlor

omet

hane

129.

391.

44

×

10

–3

(24

–25°

C)

141.

07 (

24.0

5°C

)0.

129

M (

25.0

°C)

1.43

Bro

mot

rifluo

rom

etha

ne14

8.91

5.00

×

10

–1

(25

°C)

149

(20°

C)

0.03

wt.

% (

20°C

) 2

.44

1, 3

-But

adie

ne

54.0

96.

3

×

10

–2

(25

°C)

2105

(25

°C)

735

(20°

C)

2.08

n

-But

ane

58

.12

9.30

×

10

–1

(25

°C)

1820

(25

°C)

61 (

20°C

)U

nava

ilabl

e



Tab

le A

1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3

/mo

lV

apo

r P

ress

ure

m

m H

g.

So

lub

ility

mg

/LL

og

K

oc

1-B

uten

e56

.11

2.5

×

10

–1

(25

°C)

2230

(25

°C)

222

(25°

C)

Una

vaila

ble

But

oxye

than

o11

8.18

2.36

×

10

–6

0.76

(20

°C)

Mis

cibl

eU

nava

ilabl

e

n

-but

yl A

ceta

te11

6.16

3.3

×

10

–4

(25

°C)

15 (

25°C

)50

00 (

25°C

)U

nava

ilabl

e

sec

-But

yl A

ceta

te11

6.16

1.91

×

10

–4

(20

°C)

10 (

20°C

)0.

8 w

t. %

(20

°C)

Una

vaila

ble

tert

-But

yl A

ceta

te11

6.16

——

—U

nava

ilabl

e

n

-But

yl A

lcoh

ol74

.12

8.81

×

10

–6

(25

°C)

7.0

(25°

C)

74,7

00 (

25°C

)U

nava

ilabl

e

sec

-But

yl A

lcoh

ol74

.12

1.02

×

10

–5

(25

°C)

13 (

20°C

)20

1,00

0 (2

0°C

)U

nava

ilabl

e

tert

-But

yl A

lcoh

ol74

.12

1.20

×

10

–5

(25

°C)

42 (

25°C

)M

isci

ble

Una

vaila

ble

n

-But

ylbe

nzen

e13

4.22

1.25

×

10

–2

(25

°C)

1.03

(25

°C)

1.26

(25

.0°C

)3.

40

sec

-But

ylbe

nzen

e13

4.22

1.14

×

10

–2

(25

°C)

1.81

(25

°C)

309

(25.

0°C

)2.

95

tert

-But

ylbe

nzen

e13

4.22

1.17

×

10

–2

(25

°C)

2.14

(25

°C)

34 (

25.0

°C)

2.83

n

-But

yl M

erca

ptan

90.1

87.

04

×

10–3

(20

–22°

C)

55.5

(25

°C)

590

(22°

C)

Una

vaila

ble

C

Car

bon

Dis

sulfi

de76

.13

0.0

133

360

(25°

C)

2300

(22

°C)

2.38

–2.5

5C

arbo

n Te

trac

hlor

ide

153.

820.

024

(20°

C)

113

(25°

C)

1.16

0 (2

5°C

)2.

35C

hlor

dane

409.

784.

8 ×

10–5

1 ×

10–5

(25

°C)

1.85

(25

°C)

5.57

cis-

Chl

orda

ne40

9.78

Insu

ffici

ent

vapo

r pr

essu

re d

ata

for

calc

ulat

ion

at 2

5°C

No

data

fou

nd0.

051

(20–

25°C

)6.

0

tran

s-C

hlor

dane

409.

78In

suffi

cien

t va

por

pres

sure

dat

a fo

r ca

lcul

atio

n at

25°

C

No

data

fou

ndN

o da

ta f

ound

6.0

4-C

hlor

oani

line

127.

571.

07 ×

10–5

(25

°C)

0.02

5(25

°C)

3.9

g/L

(20–

25°C

)2.

42C

hlor

oben

zene

112.

560.

0044

5 (2

5°C

)11

.8 (

25°C

)50

2 (2

5°C

)1.

68p-

Chl

oro-

m-c

reso

l14

2.59

1.78

× 1

0–6N

o da

ta f

ound

3850

(25

°C)

2.89

Chl

oroe

than

e

64.5

20.

0085

(25

°C)

1064

(20

°C)

5,74

0 (2

0°C

)0.

512-

Chl

oroe

thyl

Vin

yl E

ther

106.

552.

5 ×

10–4

26.7

5 (2

0°C

)15

,000

(20

°C)

0.82

Chl

orof

orm

119.

380.

0032

(25

°C)

198

(25°

C)

9300

(25

°C)

1.64


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 3692-

Chl

oron

apht

hale

ne16

2.62

6.12

× 1

0–40.

017

(25°

C)

6.74

(25

°C)

3.93

2-C

hlor

ophe

nol

128.

565.

6 ×

10–7

(25

°C)

1.42

(25

°C)

28,0

00 (

25°C

)2.

564-

Chl

orop

heny

l Phe

nyl E

ther

204.

662.

2 ×

10–4

0.00

27 (

25°C

)3.

3 (2

5°C

)3.

6C

hrys

ene

228.

307.

26 ×

10–2

06.

3 x

10-9

(25

°C)

0.00

6 (2

5°C

)5.

39C

amph

or15

2.24

3.00

× 1

0–5 (

20°C

)0.

18 (

20°C

)0.

12%

(20

°C)

Una

vaila

ble

Car

bary

l 2

01.2

21.

27 ×

10–5

(20

°C)

6.57

8 ×

10–6

(25

°C)

0.41

05 (

25°C

)2.

42C

arbo

fura

n22

1.26

3.88

× 1

0–8 (

30–3

3°C

)2

× 10

–5 (

33°C

)70

0 (2

5°C

)2.

2C

hlor

oace

tald

ehyd

e78

.50

—10

0 (2

0°C

)ab

out

50 w

t. %

, fo

rms

a he

mih

ydra

teU

nava

ilabl

e

α-C

hlor

oace

toph

enon

e15

4.60

—0.

012

(20°

C)

Mis

cibl

eU

nava

ilabl

eo-

Chl

orob

enzy

liden

emal

onitr

ile18

8.61

Not

app

licab

le re

acts

with

w

ater

3.4

x 10

–5 (

20°C

)N

ot a

pplic

able

reac

ts w

ith

wat

erN

ot a

pplic

able

re

acts

with

w

ater

p-C

hlor

onitr

oben

zene

157.

56<

6.91

× 1

0–3 (

20°C

)<

1 (2

0°C

)0.

003

wt.

% (

20°C

)2.

68l-C

hlor

o-1-

nitr

opro

pane

123.

541.

57 ×

10–1

(20

–25°

C)

5.8

(25°

C)

<0.

8 w

t. %

(20

°C)

3.34

Chl

orop

icrin

164.

388.

4 ×

10–2

23.8

(25

°C)

1.62

1 g/

L (2

5°C

)0.

82C

hlor

opre

ne88

.54

3.20

× 1

0–220

0 (2

0°C

)—

—C

hlor

opyr

ifos

350.

594.

16 ×

10–6

(25

°C)

1.87

× 1

0–5 (

25°C

)2

(25°

C)

3.86

Cro

tona

ldeh

yde

70.0

91.

96 ×

10–5

30 (

20°C

)18

.1 w

t. %

(20

°C)

Una

vaila

ble

Cyc

lohe

ptan

e

98.1

9—

—30

(25

°C)

Una

vaila

ble

Cyc

lohe

xane

84.1

61.

94 ×

10–1

(25

°C)

95 (

20°C

)58

.4 (

25°C

)U

nava

ilabl

eC

yclo

hexa

nol

100.

165.

74 ×

10–6

(25

°C)

1 (2

0°C

)36

,000

(20

°C)

Una

vaila

ble

Cyc

lohe

xano

ne

98.1

41.

2 ×

10–5

(25

°C)

4 (2

0°C

)23

,000

(20

°C)

Una

vaila

ble

Cyc

lohe

xene

82

.15

4.6

× 10

–2 (

25°C

)67

(20

°C)

213

(25°

C)

Una

vaila

ble

Cyc

lope

ntad

iene

66.1

0—

—0.

0103

mol

/L a

t ro

om

tem

pera

ture

Una

vaila

ble

Cyc

lope

ntan

e70

.13

1.86

× 1

0–1 (

25°C

)40

0 (3

1.0°

C)

164

(25°

C)

Una

vaila

ble

Cyc

lope

nten

e68

.12

6.3

× 10

–2 (

25°C

)—

535

(25°

C)

Una

vaila

ble

D

p, p

′ DD

D32

0.05

2.16

× 1

0–51.

02 ×

10–6

(30

°C)

0.16

0 (2

4°C

)4.

64p,

p′ D

DE

319.

032.

34 ×

10–5

6.49

× 1

0–6 (

30°C

)0.

0013

(25

°C)

6p,

p′ D

DT

354.

495.

2 ×

10–5

1.9

x 10

–7 (

25°C

)0.

0004

(25

°C)

6.26


370 NATURAL AND ENHANCED REMEDIATION SYSTEMSTa

ble

A1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3 /m

ol

Vap

or

Pre

ssu

re

mm

Hg

.S

olu

bili

ty m

g/L

Lo

g K

oc

Dib

enz[

a,h]

anth

race

ne27

8.36

7.33

× 1

0–9≈

10–1

0 (2

0°C

)0.

0024

9 (2

5°C

)6.

22D

iben

zofu

ran

168.

20In

suffi

cien

t va

por

pres

sure

dat

a fo

r ca

lcul

atio

n at

(25

°C)

No

data

fou

nd10

(25

°C)

3.91

–4.1

0

Dib

rom

ochl

orom

etha

ne20

8.28

9.9

× 10

–476

(20

°C)

4000

(20

°C)

1.92

Di-n

-but

yl P

htha

late

278.

356.

3 ×

10–5

1.4

× 10

-5 (

25°C

)40

0 (2

5°C

)3.

141,

2-D

ichl

orob

enze

ne14

7.00

0.00

24 (

25°C

)1.

5 (2

5°C

)14

5 (2

5°C

)3.

231,

3-D

ichl

orob

enze

ne14

7.00

0.00

47 (

25°C

)2.

3 (2

5°C

)14

3 (2

5°C

)3.

231,

4-D

ichc

loro

benz

ene

147.

000.

0044

5 (2

5°C

)0.

4 (2

5°C

)74

(25

°C)

2.2

3, 3

′-Dic

hlor

oben

zidi

ne25

3.13

4.5

× 10

–8 (

25°C

)1

× 10

–5 m

/L (

22°C

)3.

11 (

25°C

)3.

3D

ichl

orod

ifluo

rom

etha

ne12

0.91

0.42

5 (2

5°C

)4.

887

(25°

C)

280

(25°

C)

2.56

1, 1

-Dic

hlor

oeth

ane

98.9

60.

0058

7 (2

5°C

)23

4 (2

5°C

)50

60 (

25°C

)1.

481,

2-D

ichl

oroe

than

e98

.96

9.8

× 10

–4 (

25°C

)87

(25

°C)

8300

(25

°C)

1.15

1, 1

-Dic

hlor

oeth

ylen

e96

.94

0.02

159

1 (2

5°C

)50

00 (

25°C

)1.

81tr

ans-

1, 2

-Dic

hlor

oeth

ylen

e96

.94

0.00

674

(25°

C)

410

(30°

C)

6300

(25

°C)

1.77

2, 4

-Dic

hlor

ophe

no16

3.00

6.66

× 1

0–60.

089

(25°

C)

4500

(25

°C)

2.94

1, 2

-Dic

hlor

ophe

nol

112.

990.

0029

4 (2

5°C

)50

(25

°C)

2800

(25

°C)

1.71

cis-

1, 3

-Dic

hlor

opro

pyle

ne11

0.97

0.00

355

43 (

25°C

)27

00 (

25°C

)1.

68tr

ans-

1, 3

-Dic

hlor

opro

pyle

ne11

0.97

0.00

355

34 (

25°C

)28

00 (

25°C

)1.

68D

ield

rin38

0.91

2 ×

10–7

1.8

× 10

–7 (

25°C

)0.

20 (

25°C

)4.

55D

ieth

yl P

htha

late

222.

248.

46 ×

10–7

0.22

(±

0.7)

Pa

(25°

C)

1000

(25

°C)

1.84

2, 4

-Dim

ethy

lphe

nol

122.

176.

55 ×

10–6

(25

°C)

0.09

8 (2

5°C

)78

68 (

25°C

)2.

07D

imet

hyl P

htha

late

194.

194.

2 ×

10–7

0.22

± 0

.7 P

a (2

5°C

)43

20 (

25°C

)1.

634,

6-D

initr

o-o-

cres

ol19

8.14

1.4

× 10

–65.

2 ×

10–5

(25

°C)

250

(25°

C)

2.64

2, 4

-Din

itrop

heno

l18

4.11

1.57

× 1

0–8 (

18–2

0°C

)0.

0003

9 (2

0°C

)60

00 (

25°C

)1.

252,

4-D

initr

otol

uene

182.

148.

67 ×

10–7

1.1

× 10

–4 (

20°C

)27

0 (2

2°C

)1.

792,

6-D

initr

otol

uene

182.

142.

17 ×

10–7

3.5

× 10

–4 (

20°C

) ≈

300

1.79

Di-n

-oct

yl P

htha

late

390.

571.

41 ×

10–1

2 (2

5°C

)0.

0014

mm

(25

°C)

3 (2

5°C

)8.

99


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 3711,

2-D

iphe

nylh

ydra

zine

18

4.24

4.11

× 1

0–11

(25°

C)

2.6

× 10

–5 (

25°C

)22

1 (2

5°C

)2.

822,

4-D

2

21.0

41.

95 ×

10–2

(20

°C)

0.00

47 (

20°C

)89

0 pp

m (

25°C

)1.

68D

ecah

ydro

naph

thal

ene

138.

2539

.2 (

25°C

)1

(22.

5°C

)0.

889

ppm

(25

°C)

Una

vaila

ble

n-D

ecan

e14

2.28

1.87

× 1

0–1 (

25°C

)1.

35 (

25°C

)0.

022

(25°

C)

Una

vaila

ble

Dia

ceto

ne A

lcoh

ol11

6.16

—1

(22.

0°C

)M

isci

ble

Una

vaila

ble

1, 4

-Dib

rom

oben

zene

235.

915.

0 ×

10–4

(25

°C)

0.16

1 (2

5°C

)16

.5 (

25°C

)3.

21,

2-D

ibro

mo-

3-ch

loro

prop

ane

236.

36

2.49

× 1

0–4 (

20°C

)0.

8 (2

1°C

)10

00 a

t ro

om

tem

pera

ture

2.11

Dib

rom

odifl

uoro

met

hane

209.

82—

688

(20°

C)

——

1-3-

Dic

hlor

o-5,

5

Dim

ethy

lhyd

anto

in19

7.03

Not

app

licab

le re

acts

with

w

ater

—0.

21 w

t. %

(25

°C)

Not

app

licab

le

reac

ts w

ith

wat

erD

ichl

orofl

uoro

met

hane

120.

91≈

2.42

× 1

0–2 (

20–3

0°C

)76

0 (8

.9°C

)1

wt.

% (

20°C

)1.

57sy

m-D

ichl

orom

ethy

l Eth

er11

4.96

Not

app

licab

le re

acts

with

w

ater

—D

ecom

pose

sN

ot a

pplic

able

re

acts

with

w

ater

Dic

hlor

vos

220.

985.

0 ×

10–3

0.05

27 (

25°C

)≈

1 w

t. %

(20

°C)

9.57

Die

thyl

amin

e73

.14

2.56

× 1

0–5 (

25°C

)19

5 (2

0°C

)81

5,00

0 (1

4°C

)U

nava

ilabl

e2-

Die

thyl

amin

oeth

anol

117.

19—

1 (2

0°C

)M

isci

ble

Una

vaila

ble

1, 1

-Difl

uoro

tetr

achl

oroe

than

e20

3.83

—40

(19

.8°C

)—

—1,

2-D

ifluo

rote

trac

hlor

oeth

ane

203.

831.

07 ×

10–1

(20

°C)

40 (

19.8

°C)

0.01

wt.

% (

20°C

)2.

78D

iisob

utyl

Ket

one

142.

246.

36 ×

10–4

(20

°C)

1.7

(20°

C)

0.05

wt.

% (

20°C

)U

nava

ilabl

eD

iisop

ropy

lam

ine

101.

19—

60 (

20°C

)M

isci

ble

Una

vaila

ble

N,

N-D

imet

hyla

ceta

mid

e11

5.18

—1.

3 (2

5°C

)M

isci

ble

Una

vaila

ble

Dim

ethy

lam

ine

45.0

81.

77 ×

10–5

(25

°C)

1520

(10

°C)

Mis

cibl

eU

nava

ilabl

ep-

Dim

ethy

lam

inoa

zobe

nzen

e22

5.30

——

13.6

(20

–30°

C)

3D

imet

hyla

nilin

e12

1.18

4.98

× 1

0–6 (

20°C

)1

(29.

5°C

)11

05.2

(25

°C)

Una

vaila

ble

2, 2

-Dim

ethy

lbut

ane

86.1

81.

943

(25°

C)

319.

1 (2

5°C

)21

.2 (

25°C

)U

nava

ilabl

e2,

3 -

Dim

ethy

lbut

ane

86.1

81.

18 (

25°C

)23

4.6

(25°

C)

19.1

(25

°C)

Una

vaila

ble

cis-

1, 2

-Dim

ethy

lcyc

lohe

xane

112.

223.

54 ×

10–1

(25

°C)

14.5

(25

°C)

6.0

(25°

C)

Una

vaila

ble

tran

s-1,

4-D

imet

hylc

yclo

hexa

ne11

2.22

8.70

× 1

0–1 (

25°C

)22

.65

(25°

C)

3.84

ppm

(25

°C)

Una

vaila

ble

Dim

ethy

lform

amid

e73

.09

—3.

7 m

m (

25°C

)M

isci

ble

Una

vaila

ble

1, 1

-Dim

ethy

lhyd

razi

ne60

.10

2.45

× 1

0–9 (

25°C

)15

7 (2

5°C

)M

isci

ble

–0.7



ble

A1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3 /m

ol

Vap

or

Pre

ssu

re

mm

Hg

.S

olu

bili

ty m

g/L

Lo

g K

oc

2, 3

-Dim

ethy

lpen

tane

100.

201.

73 (

25°C

)10

0 (3

3.3°

C)

5.25

(25

°C)

Una

vaila

ble

2, 4

Dim

ethy

lpen

tane

100.

203.

152

(25°

C)

98.4

(25

°C)

5.50

(25

°C)

Una

vaila

ble

3, 3

-Dim

ethy

lpen

tane

100.

201.

84 (

25°C

)82

.8 (

25°C

)5.

94 (

25°C

)U

nava

ilabl

e2,

2-D

imet

hylp

ropa

ne72

.15

2.18

(25

°C)

1.28

7 (2

5°C

)33

.2 (

25°C

)U

nava

ilabl

e2,

7-D

imet

hylq

uino

line

157.

22—

—1.

795

(25°

C)

—D

imet

hyl S

ulfa

te12

6.13

2.96

× 1

0–6 (

20°C

)0.

5 (2

0°C

)2.

8 w

t. %

(20

°C)

0.61

1, 2

-Din

itrob

enze

ne16

8.11

<1.

47 ×

10–3

(20

°C)

<1

(20°

C)

0.01

5 w

t. %

(20

°C)

Una

vaila

ble

1, 3

-Din

itrob

enze

ne16

8.11

2.75

× 1

0–7 (

35°C

)8.

15 ×

10–4

(35

°C)

0.05

wt.

% (

20°C

)2.

181,

4-D

initr

oben

zene

168.

114.

79 ×

10–7

(35

°C)

2.25

× 1

0–4 (

35°C

)0.

01 w

t. %

(20

°C)

Una

vaila

ble

Dio

xane

88.

114.

88 ×

10–6

(25

°C)

37 (

25°C

)M

isci

ble

0.54

Diu

ron

233.

111.

46 ×

10–9

(25

–30°

C)

2 ×

10–7

(30

°C)

42 (

25°C

)2.

51n-

Dod

ecan

e17

4.34

24.2

(25

°C)

0.05

7 (2

5°C

)0.

008

(25°

C)

Una

vaila

ble

E

α-E

ndos

ulfa

n 4

06.9

21.

01 ×

10–4

(25

°C)

10–5

(25

°C)

0.53

0 (2

5°C

)3.

31β-

End

osul

fan

406.

921.

91 ×

10–5

(25

°C)

10–5

(25

°C)

0.28

0 (2

5°C

)3.

37E

ndos

ulfa

n S

ulfa

te42

2.92

Insu

ffici

ent

vapo

r pr

essu

re d

ata

for

calc

ulat

ion

No

data

fou

nd0.

117

3.37

End

rin38

0.92

5.0

× 10

–77

× 10

–7 (

25°C

)0.

26 (

25°C

)3.

92E

ndrin

Ald

ehyd

e38

0.92

3.86

× 1

0–7 (

25°C

)2

× 10

–7 (

25°C

)0.

26 (

25°C

)4.

43E

thyl

benz

ene

06.1

70.

0086

8 (2

5°C

)10

(25

.9°C

)15

2 (2

5°C

)1.

98E

pich

loro

hydr

in92

.53

2.38

–2.5

4 ×

10–5

(20

°C)

13 (

20°C

)60

,000

(2

0°C

)1

EP

N32

3.31

—0.

0003

(10

0°C

)—

3.12

Eth

anol

amin

e

61.0

8—

<1

(20°

C)

Mis

cibl

eU

nava

ilabl

e2-

Eth

oxye

than

ol90

.12

—4

(20°

C)

Mis

cibl

eU

nava

ilabl

e2-

Eth

oxye

thyl

Ace

tate

132.

189.

07 ×

10–7

(20

°C)

2 (2

0°C

)23

wt.

% (

20°C

)U

nava

ilabl

eE

thyl

Ace

tate

88

.11

1.34

× 1

0–4 (

25°C

)94

.5 (

25°C

)10

0 m

l/L (

25°C

)U

nava

ilabl

eE

thyl

Acr

ylat

e10

0.12

1.94

–2.5

9 ×

10–3

(20

°C)

29.5

(20

°C)

.5 w

t. %

(20

°C)

Una

vaila

ble


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 373E

thyl

amin

e45

.08

1.07

× 1

0–5 2

5°C

)40

0 (2

.0°C

)M

isci

ble

Una

vaila

ble

Eth

yl B

rom

ide

108.

977.

56 ×

10–3

(25

°C)

386

(20°

C)

0.9

wt.

% (

20°C

) 2

.67

Eth

ylcy

clop

enta

ne98

.19

2.10

× 1

0–2 (

25°C

)40

(25

.0°C

)24

5 (2

5°C

)U

nava

ilabl

eE

thyl

ene

Chl

oroh

ydrin

80.5

1—

8 (2

5°C

)M

isci

ble

Una

vaila

ble

Eth

ylen

edia

min

e 6

0.10

1.73

× 1

0–9 (

25°C

)10

(21

.5°C

)M

isci

ble

Una

vaila

ble

Eth

ylen

e D

ibro

mid

e18

7.86

7.06

× 1

0–4 (

25°C

)11

(25

°C)

3370

1.64

Eth

ylen

imin

e43

.07

1.33

× 1

0–7 (

25°C

) 2

50 (

30°C

) M

isci

ble

0.11

Eth

yl E

ther

74.1

2 1.

28 ×

10–3

(25

°C)

442

(20°

C)

6.05

wt.

% (

25°C

)U

nava

ilabl

eE

thyl

For

mat

e74

.08

2.23

× 1

0–4 (

25°C

)19

4 (2

0°C

)11

8,00

0 (2

5°C

)U

nava

ilabl

eE

thyl

Mec

apta

n62

.13

2.74

× 1

0–3 (

25°C

)52

7.2

(25°

C)

1.3

wt.

% (

20°C

)U

nava

ilabl

e4-

Eth

ylm

orph

olin

e11

5.18

—6.

1 (2

0°C

)M

isci

ble

Una

vaila

ble

2-E

thyl

thio

phen

e11

2.19

—60

.9 (

60.3

°C)

292

(25°

C)

Una

vaila

ble

F

Flu

oran

then

e20

2.26

0.01

69 (

25°C

) 5

.0 ×

10-

6 (2

5°C

)0.

265

(25°

C)

4.62

Flu

oren

e16

6.22

2.1

× 10

–410

(14

6°C

)1.

98 (

25°C

)3.

7F

orm

alde

hyde

30.0

33.

27 ×

10–7

400

(–33

°C)

Mis

cibl

e0.

56F

orm

ic A

cid

46.0

31.

67 ×

10–7

at

pH 4

35 (

20°C

)M

isci

ble

Una

vaila

ble

Fur

fura

l96

.09

1.52

–3.0

5 ×

10–6

(20

°C)

2 (2

0°C

)8.

3 w

t. %

(20

°C)

Una

vaila

ble

Fur

fury

l Alc

ohol

98.1

0—

0.4

(20°

C)

Mis

cibl

e U

nava

ilabl

e

G

Gly

cido

l74

.08

—0.

9 (2

5°C

)M

isci

ble

Una

vaila

ble

H

Hep

tach

lor

373.

320.

0023

4 ×

10–4

(25

°C)

180

ppb

(25°

C)

4.34

Hep

tach

lor

Epo

xide

389.

323.

2 ×

10–5

2.6

× 1

0–6 (

20°C

)0.

350

(25°

C)

4.32

Hex

achl

orob

enze

ne28

4.78

0.00

171.

089

× 10

–5 (

20°C

)0.

006

(25°

C)

3.59

Hex

achc

hlor

obut

adie

ne26

0.76

0.02

60.

15 (

20°C

)3.

23 (

25°C

)3.

67H

exac

hlor

ocyc

lope

ntad

iene

272.

77 0

.016

0.08

1 (2

5°C

)1.

8 (2

5°C

)3.

63H

exac

hlor

oeth

ane

236.

740.

0025

0.8

(30°

C)

27.2

(25

°C)

3.34

2-H

exan

one

100.

160.

0017

5 (2

5°C

)3.

8 (2

5°C

)35

,000

(25

°C)

2.13



ble

A1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3 /m

ol

Vap

or

Pre

ssu

re

mm

Hg

.S

olu

bili

ty m

g/L

Lo

g K

oc

n-H

epta

ne10

0.20

2.03

5 (2

5°C

)45

.85

(25°

C)

2.24

(25

°C)

Una

vaila

ble

2-H

epta

none

114.

191.

44 ×

10–4

(25

°C)

2.6

(20°

C)

0.43

wt.

% (

25°C

)U

nava

ilabl

e3-

Hep

tano

ne11

4.19

4.20

× 1

0–5 (

20°C

)1.

4 (2

5°C

)14

,300

(20

°C)

Una

vaila

ble

cis-

2-H

epte

ne98

.19

4.13

× 1

0–1 (

20°C

)48

(25

°C)

15 (

25°C

)U

nava

ilabl

etr

ans-

2-H

epte

ne98

.19

4.22

× 1

0–1 (

25°C

)49

(25

°C)

15 (

25°C

)U

nava

ilabl

en-

Hex

ane

86.1

81.

184

(25°

C)

151.

5 (2

5°C

)9.

47 (

25°C

)U

nava

ilabl

el-H

exen

e84

.16

4.35

× 1

0–1 (

25°C

)18

6.0

(25°

C)

50 (

25°C

) U

nava

ilabl

ese

c-H

exyl

Ace

tate

144.

214.

38-5

.84

× 10

–3 (

20°C

)4

(20°

C)

0.01

3 w

t. %

(20

°C)

Una

vaila

ble

Hyd

rogu

inon

e11

0.11

<2.

07 ×

10–9

(20

–25°

C)

1 (1

32.4

)70

,000

(25

°C)

0.98

I

Inde

no[1

, 2,

3-c

d]py

rene

276.

342.

96 ×

10–2

0 (2

5°C

)0–1

0 (2

5°C

)0.

062

7.49

Isop

horo

ne13

8.21

5.8

× 10

–60.

38 (

20°C

)12

,000

(25

°C)

1.49

Inda

n11

8.18

——

88.9

(25

°C)

2.48

Indo

le11

7.15

——

3558

(25

°C)

1.69

Indo

line

——

—10

,800

(25

°C)

1.42

1-Io

dopr

opan

e16

9.99

9.09

× 1

0–343

.1 (

25°C

)0.

1065

wt.

% (

23.5

°C)

2.16

Isoa

myl

Ace

tate

130.

195.

87 ×

10–2

(25

°C)

4 (2

0°C

)0.

2 w

t. %

(20

°C)

1.95

Isoa

myl

Alc

ohol

88.1

58.

89 ×

10–6

(20

°C)

2.3

(20°

C)

26,7

20 (

22°C

)U

nava

ilabl

eIs

obut

yl A

ceta

te11

6.16

4.85

× 1

0–4 (

25°C

)20

(25

°C)

6300

(2

5°C

)U

nava

ilabl

eIs

obut

yl A

lcoh

ol74

.12

9.25

× 1

0–6 (

20°C

)10

.0 (

20°C

)8.

7 w

t. %

(20

°C)

Una

vaila

ble

Isob

utyl

Ben

zene

134.

221.

09 ×

10–2

(25

°C)

2.06

(25

°C)

33.7

1 (2

5°C

)3.

9Is

opro

pyl A

ceta

te10

2.13

2.81

× 1

0–4 (

25°C

)73

(25

°C)

18,0

00 (

20°C

)U

nava

ilabl

eIs

opro

pyla

min

e59

.11

—47

8 (2

0°C

)M

isci

ble

Una

vaila

ble

Isop

ropy

lben

zene

120.

191.

47 ×

10–2

(25

°C)

4.6

(25°

C)

48.3

(25

°C)

3.45

Isop

ropy

l Eth

er10

2.18

9.97

× 1

0–3 (

25°C

)15

0 (2

5°C

)0.

65 w

t. %

(25

°C)

Una

vaila

ble


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 375

K

Kep

one

490.

683.

11

×

10

–2

(25

°C)

2.25

(25

°C)

2.7

(20

–25°

C)

4.74

L

Lind

ane

290.

834.

8

×

10

–7

6.7

×

10

–5

(25

°C)

7.52

(25

°C)

3.0

3

M

Met

hoxy

chlo

r34

5.66

Insu

ffici

ent

vapo

r pr

essu

re d

ata

for

calc

ulat

ion

at (

25°C

)

No

data

fou

nd0.

1 (2

5°C

)4.

9

Met

hyl B

rom

ide

94.9

4 0

.216

33 (

25°C

)13

,000

(25

°C)

1.92

Met

hyl C

hlor

ide

50.4

80.

010

(25°

C)

3789

(20

°C)

7400

(25

°C)

1.4

Met

hyle

ne C

hlor

ide

84.9

3 0

.002

69 (

25°C

)45

5 (2

5°C

)13

,000

(25

°C)

0.94

2-M

ethy

lnap

htha

lene

142.

20In

suffi

cien

t va

por

pres

sure

dat

a fo

r ca

lcul

atio

n

No

data

fou

nd25

.4 (

25°C

)3.

93

4-M

ethy

l-2-P

enta

none

100.

161.

49

×

10

–5

(25

°C)

15 (

20°C

)1.

91 w

t. %

(25

°C)

0.79

2-M

ethy

lphe

nol

108.

141.

23

×

10

–6

(25

°C)

0.24

(25

°C)

25,0

00 (

25°C

)1.

344-

Met

hylp

heno

l10

8.14

7.92

×

10

–7

(25

°C)

0.10

8 (2

5°C

)23

,000

(25

°C)

1.69

Mal

athi

on33

0.36

4.89

x 1

0

–9

(25

°C)

7.95

×

10

–6

(25

°C)

330

(30°

C)

2.46

Mal

eic

Anh

ydrid

e98

.06

Not

app

licab

le re

acts

with

w

ater

5

×

10

–5

(20

°C)

—N

ot a

pplic

able

re

acts

with

w

ater

Mes

ityl O

xide

98.1

44.

01

×

10

–6

(20

°C)

8.7

(20°

C)

3 w

t. %

(20

°C)

Una

vaila

ble

Met

hyl A

ceta

te74

.08

9.09

×

10

–5

(25

°C)

235

(25°

C)

240,

000

(20°

C)

Una

vaila

ble

Met

hyl A

cryl

ate

86.0

91.

23–1

.44

×

10

–4

(20

°C)

70 (

20°C

)52

,000

Una

vaila

ble

Met

hyla

l76

.10

1.73

x 1

0

–4

(25

°C)

400

(25°

C)

33 w

t. %

(20

°C)

Una

vaila

ble

Met

hyl A

lcoh

ol32

.04

4.66

×

10

–6

(25

°C)

127.

2 (2

5°C

)M

isci

ble

Una

vaila

ble

Met

hyla

min

e31

.06

1.81

×

10

–2

(25

°C)

3.1

atm

(20

°C)

9.59

0 (2

5°C

)U

nava

ilabl

eM

ethy

lani

line

107.

161.

19

×

10

–5

(25

°C)

<1.

0 (2

0°C

)5.

624

g/L

(25°

C)

Una

vaila

ble

2-M

ethy

lant

hrac

ene

192.

96—

—0.

039

(25°

C)

5.12

L1282/Appendix A/frame Page 375 Tuesday, June 19, 2001 1:13 PM


ble

A1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3 /m

ol

Vap

or

Pre

ssu

re

mm

Hg

.S

olu

bili

ty m

g/L

Lo

g K

oc

2-M

ethy

l-1,

3-B

utad

iene

68.1

27.

7 ×

10–2

(25

°C)

550.

1 (2

5°C

)64

2 (2

5°C

)U

nava

ilabl

e2-

Met

hylb

utan

e72

.15

1.35

(25

°C)

687.

4 (2

5°C

)49

.6 (

25°C

)U

nava

ilabl

e3-

Met

hyl-1

-But

ene

70.1

35.

35 ×

10–1

(25

°C)

902.

1 (2

5°C

)13

0 (2

5°C

)U

nava

ilabl

eM

ethy

l Cel

loso

lve

76.1

0—

6 (2

0°C

)M

isci

ble

Una

vaila

ble

Met

hyl C

ello

solv

e A

ceta

te11

8.13

—7

(20°

C)

Mis

cibl

eU

nava

ilabl

eM

ethy

lcyc

lohe

xane

98.1

94.

35 ×

10–1

(25

°C)

46.3

(25

°C)

16.0

(25

°C)

Una

vaila

ble

o-M

ethy

lcyc

lohe

xano

ne11

2.17

—≈1

(20

°C)

——

l-M

ethy

lcyc

lohe

xene

96.1

7—

—52

(2°

C)

Una

vaila

ble

Met

hylc

yclo

pent

ane

84.1

63.

62 ×

10–1

(25

°C)

137.

5 (2

5°C

)41

.8 (

25°C

)U

nava

ilabl

eM

ethy

l For

mat

e60

.05

2.23

× 1

0–4 (

25°C

)62

5 (2

5°C

)30

wt.

% (

20°C

)U

nava

ilabl

e3-

Met

hylh

epta

ne11

4.23

3.70

(25

°C)

19.5

(25

°C)

0.79

2 (2

5°C

)U

nava

ilabl

e5-

Met

hyl-3

-Hep

tano

ne12

8.21

1.30

× 1

0–4 (

20°C

)2

(25°

C)

0.26

wt.

% (

20°C

)U

nava

ilabl

e2-

Met

hylh

exan

e10

0.20

3.42

(25

°C)

65.9

(25

°C)

2.54

(25

°C)

Una

vaila

ble

3-M

ethy

lhex

ane

100.

201.

55–1

.64

(25°

C)

61.6

(25

°C)

4.95

(25

°C)

Una

vaila

ble

Met

hylh

ydra

zine

46.0

7—

49.6

(25

°C)

Mis

cibl

eU

nava

ilabl

eM

ethy

l Iod

ide

141.

945.

87 ×

10–3

(25

°C)

405

(25°

C)

2 w

t. %

(20

°C)

1.36

Met

hyl I

socy

anat

e57

.05

3.89

× 1

0–4 (

20°C

)34

8 (2

0°C

)6.

7 w

t. %

(20

°C)

Una

vaila

ble

Met

hyl M

erca

ptan

48.1

03.

01 ×

10–3

(25

°C)

1516

(25

°C)

23.3

0 g/

L (2

0°C

)U

nava

ilabl

eM

ethy

l Met

hacr

ylat

e10

0.12

2.46

× 1

0–4 (

20°C

)40

(26

°C)

1.5

wt.

% (

20°C

)U

nava

ilabl

e4-

Met

hylo

ctan

e12

8.26

10.2

7 (2

5°C

)7

(25°

C)

0.11

5 (2

5°C

)U

nava

ilabl

e2-

Met

hylp

enta

ne86

.18

1.73

2 (2

5°C

)21

1.8

(25°

C)

13.8

(25

°C)

Una

vaila

ble

3-M

ethy

lpen

tane

86.1

81.

693

(25°

C)

189.

8 (2

5°C

)17

.9 (

25°C

) U

nava

ilabl

e2-

Met

hyl-1

-Pen

tene

84.1

62.

77 ×

10–1

(25

°C)

195.

4 (2

5°C

)78

(25

°C)

Una

vaila

ble

4-M

ethy

l-1-P

ente

ne84

.16

6.15

× 1

0–1 (

25°C

)27

0.8

(25°

C)

48 (

25°C

)U

nava

ilabl

e1-

Met

hylp

hena

nthr

ene

192.

26—

—26

9 pp

b (2

5°C

)4.

562-

Met

hylp

ropa

ne58

.12

1.17

1 (2

5°C

)10

atm

(66

.8°C

)48

.9 (

25°C

)U

nava

ilabl

e2-

Met

hylp

rope

ne56

.11

2.1

× 10

–1 (

25°C

)2.

270

(25°

C)

263

(25°

C)

Una

vaila

ble

α-M

ethy

lsty

rene

118.

18—

1.9

(20°

C)

——


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 377M

evin

phos

224.

16—

0.00

3 (2

0°C

)M

isci

ble

Una

vaila

ble

Mor

phol

ine

87.1

2—

13.4

(25

°C)

Mis

cibl

eU

nava

ilabl

e

N

Nap

htha

lene

128.

184.

6 ×

10–4

0.23

(25

°C)

30 (

25°C

)2.

742-

Nitr

oani

line

138.

139.

72 ×

10–5

(25

°C)

8.1

(25°

C)

1260

(2

5°C

)1.

23–1

.62

3-N

itroa

nilin

e13

8.13

Insu

ffici

ent

vapo

r pr

essu

re d

ata

for

calc

ulat

ion

1 (1

19.3

°C)

890

(25°

C)

1.26

4-N

itroa

nilin

e13

8.13

1.14

× 1

0–8 (

25°C

)0.

0015

(20

°C)

800

(18.

5°C

)1.

08N

itrob

enze

ne12

3.11

2.45

× 1

0–50.

28 (

25°C

)2,

000

(25°

C)

2.36

2-N

itrop

heno

l13

9.11

3.5

× 10

–60.

20 (

25°C

)2,

000

(25°

C)

1.57

4-N

itrop

heno

l13

9.11

3.0

× 10

–5 (

20°C

)10

-4 (

20°C

)16

,000

(25

°C)

2.33

N-N

itros

odim

ethy

lam

ine

74.0

90.

143

(25°

C)

8.1

(25°

C)

Mis

cibl

e1.

41N

-Nitr

osod

iphe

nyla

min

e19

8.22

2.33

× 1

0–8 (

25°C

)N

o da

ta f

ound

35.1

(25

°C)

2.76

N-N

itros

odi-n

-Pro

pyla

min

e13

0.19

Insu

ffici

ent

vapo

r pr

essu

re d

ata

to

calc

ulat

e

No

data

fou

nd99

00 (

25°C

)1.

01

Nal

ed38

0.79

—2

× 10

–4 (

20°C

)—

Not

app

licab

le

reac

ts w

ith

wat

er1-

Nap

hthy

lam

ine

143.

191.

27 ×

10–1

0 (2

5°C

)6.

5 ×

10–5

(2

0–30

°C)

1700

3.5

1

2-N

apht

hyla

min

e14

3.19

2.0

1 ×

10–9

(25

°C)

2.56

× 1

0–4

(20–

30°C

)58

6 (2

0–30

°C)

2.11

Nitr

apyr

in23

0.90

2.13

× 1

0–30.

0028

(20

°C)

40 2

.64

4-N

itrob

iphe

nyl

199.

21—

——

—N

itroe

than

e75

.07

4.66

× 1

0–5 (

25°C

)15

.6 (

20°C

)45

ml/L

(20

°C)

Una

vaila

ble

Nitr

omet

hane

61.0

42.

86 ×

10–5

27.8

(20

°C)

22 m

l/L (

20°C

)U

nava

ilabl

e1-

Nitr

opro

pane

89.0

98.

68 ×

10–5

(25

°C)

7.5

(20°

C)

1.4

wt.

% (

20°C

)U

nava

ilabl

e2-

Nitr

opro

pane

89.0

91.

23 ×

10–4

(25

°C)

12.9

(20

°C)

1.7

wt.

% (

20°C

)U

nava

ilabl

e2-

Nitr

otol

uene

137.

144.

51 ×

10–5

(20

°C)

0.15

(20

°C)

0.06

wt.

% (

20°C

)U

nava

ilabl

e3-

Nitr

otol

uene

13

7.14

5.41

× 1

0–5 (

20°C

)0.

25 (

25°C

)0.

05 w

t. %

(20

°C)

Una

vaila

ble



ble

A1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3 /m

ol

Vap

or

Pre

ssu

re

mm

Hg

.S

olu

bili

ty m

g/L

Lo

g K

oc

4-N

itrot

olue

ne13

7.14

5.0

× 10

–5 (

25°C

)5.

484

(26.

0°C

)0.

005

wt.

% (

20°C

)U

nava

ilabl

en-

Non

ane

128.

265.

95 (

25°C

)4.

3 (2

5°C

)0.

122

(25°

C)

Una

vaila

ble

O

Oct

achl

oron

apht

hale

ne40

3.73

—<

1 (2

0°C

)—

—n-

Oct

ane

114.

233.

225

(25°

C14

.14

(25°

C)

0.43

1 (2

5°C

)U

nava

ilabl

el-O

cten

e11

2.22

9.52

× 1

0–1 (

25°C

)17

.4 (

25°C

)2.

7 (2

5°C

)U

nava

ilabl

eO

xalic

Aci

d90

.04

1.43

× 1

0–10

(pH

4)

<0.

001

(20°

C)

9.81

wt.

% (

25°C

) 0

.89

P

PC

B-1

016

257.

9075

04

× 10

–4 (

25°C

)0.

22–0

.25

4.7

PC

B-1

221

192.

003.

24 ×

10–4

0.00

67 (

25°C

)1.

5 (2

5°C

)2.

44P

CB

-123

222

1.00

8.64

× 1

0–40.

0046

(25

°C)

1.45

(25

°C)

2.83

PC

B-1

242

154–

358

with

an

aver

age

valu

e of

26

1

5.6

× 10

–44.

06 ×

10–4

(25

°C)

0.24

(25

°C)

3.71

PC

B-1

248

222–

358

with

an

aver

age

valu

e of

28

8

0.00

354.

94 ×

10–4

(25

°C)

0.05

45.

64

PC

B-1

254

327

(ave

rage

)0.

0027

7.71

× 1

0–5 (

25°C

)0.

012

(25°

C)

5.61

PC

B-1

260

324–

460

370

(ave

rage

)0.

0071

4.05

× 1

0–5 (

25°C

)0.

080

(24°

C)

6.42

Pen

tach

loro

phen

ol26

6.34

3.4

× 1

0–61.

7 ×

10–4

(20

°C)

20–2

5 (2

5°C

)2.

96P

hena

nthr

ene

178.

242.

56 ×

10–5

(25

°C)

6.80

× 1

0–4 (

25°C

)1.

18(2

5°C

)3.

72P

heno

l94

.11

3.97

× 1

0–7 (

25°C

)0.

34 (

25°C

)93

,000

(25

°C)

1.43

Pyr

ene

202.

261.

87 ×

10–5

6.85

× 1

0–7 (

25°C

)0.

148

(25°

C)

4.66

Par

athi

on 2

91.2

7 8

.56

× 10

–8 (

25°C

)9.

8 ×

10–6

(25

°C)

24 (

25°C

)3.

68P

enta

chlo

robe

nzen

e25

0.34

0.00

71 (

20°C

)6.

0 ×

10–3

(2

0–30

°C)

2.24

× 1

0–6 M

(25

°C)

6.3


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 379P

enta

chlo

roet

hane

202.

282.

45 ×

10–3

(25

°C)

4.5

(25°

C)

7.69

and

500

wer

e re

port

ed a

t 25

°C a

nd

20°C

3.28

l, 4-

Pen

tadi

ene

68.1

21.

20 ×

10–1

(25

°C)

734.

6 (2

5°C

)55

8 (2

5°C

)U

nava

ilabl

en-

Pen

tane

72

.15

1.25

5 (2

5°C

)51

2.8

(25°

C)

39.5

(25

°C)

Una

vaila

ble

2-P

enta

none

86

.13

6.44

× 1

0–5 (

25°C

)16

(25

°C)

5.51

wt.

% (

25°C

) U

nava

ilabl

el-P

ente

ne70

.13

4.06

× 1

0–1 (

25°C

)63

7.7

(25°

C)

148

(25°

C)

Una

vaila

ble

cis-

2-P

ente

ne70

.13

2.25

× 1

0–1 (

25°C

)49

4.6

(25°

C)

203

(25°

C)

Una

vaila

ble

tran

s-2-

Pen

tene

70.1

32.

34 ×

10–1

(25

°C)

505.

5 (2

5°C

)20

3 (2

5°C

)U

nava

ilabl

eP

enty

cycl

open

tane

140.

28—

—0.

115

(25°

C)

Una

vaila

ble

p-P

heny

lene

diam

ine

108.

14—

—38

,000

(24

°C)

Una

vaila

ble

Phe

nyl E

ther

170.

212.

13 ×

10–4

(20

°C)

0.12

(30

°C)

21 (

25°C

)U

nava

ilabl

eP

heny

lhyd

razi

ne10

8.14

—<

0.1

(20°

C)

—U

nava

ilabl

eP

htha

lic A

nhyd

ride

148.

126.

29 ×

10–9

(20

°C)

2 ×

10–4

(20

°C)

0.62

wt.

% (

20°C

)1.

9P

icric

Aci

d 2

29.1

1<

2.15

× 1

0–5 (

20°C

)<

1 (2

0°C

)1.

4 w

t. %

(20

°C)

—P

indo

ne23

0.25

——

18 (

25°C

)2.

95P

ropa

ne44

.10

7.06

× 1

0–1 (

25°C

)8.

6 at

m (

20°C

)62

.4 (

25°C

)U

nava

ilabl

eβ-

Pro

piol

acto

ne72

.06

7.63

× 1

0–7 (

25°C

)3.

4 (2

5°C

)37

vol

. % (

25°C

)U

nava

ilabl

en-

Pro

pyl A

ceta

te10

2.12

1.99

× 1

0–4 (

25°C

)35

(25

°C)

18,9

00 (

20°C

)U

nava

ilabl

en-

Pro

pyl A

lcoh

ol60

.10

6.74

× 1

0–6 (

25°C

)20

.8 (

25°C

)M

isci

ble

Una

vaila

ble

n-P

ropy

lben

zene

120.

191.

0 ×

10–2

(25

°C)

3.43

(25

°C)

55 (

25°C

)2.

87P

ropy

lcyc

lope

ntan

e11

2.22

8.90

× 1

0–1 (

25°C

)12

.3 (

25°C

)2.

04 (

25°C

)U

nava

ilabl

eP

ropy

lene

Oxi

de58

.08

8.34

× 1

0–5 (

20°C

)44

5 (2

0°C

)41

wt.

% (

20°C

)N

ot a

pplic

able

re

acts

with

w

ater

n-P

ropy

l Nitr

ate

105.

09—

18 (

20°C

)—

Una

vaila

ble

Pro

pyne

40.0

61.

1 ×

10–1

(25

°C)

4310

(25

°C)

3640

(20

°C)

Una

vaila

ble

Pyr

idin

e79

.10

8.88

× 1

0–6 (

25°C

)20

(25

°C)

Mis

cibl

eU

nava

ilabl

e

Q

p-Q

uino

ne10

8.10

9.48

× 1

0–7 (

20°C

) 0

.1 (

20°C

)1.

5 w

t. %

(20

°C)

Una

vaila

ble



ble

A1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3 /m

ol

Vap

or

Pre

ssu

re

mm

Hg

.S

olu

bili

ty m

g/L

Lo

g K

oc

R

Ron

nel

321.

578.

46 ×

10–6

(25

°C)

8 ×

10–4

(25

°C)

40 (

25°C

)2.

76

S

Sty

rene

104.

150.

0026

16.

45 (

25°C

)0.

031

wt.

% (

25°C

)2.

87S

tryc

hnin

e33

4.42

——

0.02

wt.

% (

20°C

)2.

45S

ulfo

tepp

322.

302.

88 ×

10–6

(20

°C)

0.00

017

(20°

C)

25 2

.87

T

TC

DD

321.

985.

40 ×

10–2

3 (1

8–22

°C)

7.2

× 10

–10

(25°

C)

0.01

93 p

pb (

22°C

)6.

661,

1,

2, 2

-Tet

rach

loro

etha

ne16

7.85

4.56

× 1

0–4 (

25°C

)6

(25°

C)

2970

(25

°C)

2.07

Tetr

achl

oroe

thyl

ene

165.

830.

0153

20 (

25°C

)15

0 (2

5°C

)2.

42To

luen

e92

.14

0.00

674

(25°

C)

22 (

20°C

)49

0 (2

5°C

)2.

06To

xaph

ene

413.

820.

063

0.2–

0.4

(25°

C)

0.2–

0.4

(25°

C)

3.18

1, 2

, 4-

Tric

hlor

oben

zene

181.

450.

0023

20.

29 (

25°C

)31

.3 (

25°C

)2.

71,

1,

1-Tr

ichl

oroe

than

e13

3.40

0.01

62 (

25°C

)12

4 (2

5°C

)95

0 (2

5°C

)2.

181,

1,

2-Tr

ichl

oroe

than

e13

3.40

9.09

× 1

0–4 (

25°C

)19

(20

°C)

4,50

0 (2

0°C

)1.

75Tr

ichl

oroe

thyl

ene

131.

390.

0091

72.6

(25

°C)

1100

(25

°C)

1.81

Tric

hlor

ofluo

rom

etha

ne13

7.37

1.73

(25

°C)

792

(25°

C)

1240

(25

°C)

2.2

2, 4

, 5-

Tric

hlor

ophe

nol

197.

451.

76 ×

10–7

(25

°C)

0.02

2 (2

5°C

)1.

2 g/

L (2

5°C

)2.

852,

4,

6-Tr

ichl

orop

heno

197.

459.

07 ×

10–8

(25

°C)

0.01

7 (2

5°C

)80

0 (2

5°C

)3.

032,

4,

5-T

255.

484.

87 ×

10–8

(25

°C)

6.46

× 1

0–6 (

25°C

)27

8 (2

5°C

)1.

721,

2,

4, 5

-Tet

rabr

omob

enze

ne39

3.70

——

0.04

04.

821,

1,

2, 2

-Tet

rabr

omoe

than

e34

5.65

6.40

× 1

0–5 (

20°C

)0.

1 (2

0°C

)0.

07 w

t. %

(20

°C)

2.45

1, 2

, 3,

4-T

etra

chlo

robe

nzen

e21

5.89

6.9

× 10

–3 (

20°C

)2.

6 ×

10–2

(25

°C)

5.92

(25

°C)

5.4

aver

age

valu

e1,

2,

3, 5

-Tet

rach

loro

benz

ene

215.

891.

58 ×

10–3

(25

°C)

1 (5

8.2°

C)

5.19

(25

°C)

6.0

aver

age


PHYSICAL PROPERTIES OF SOME COMMON ENVIRONMENTAL CONTAMINANTS 3811,

2,

4, 5

-Tet

rach

loro

benz

ene

215.

891.

0 ×

10–2

(20

°C)

<0.

1 (2

5°C

) 0

.465

(25

°C)

6.1

aver

age

Tetr

aeth

yl P

yrop

hosp

hate

290.

20—

1.55

× 1

0–4 (

20°C

)M

isci

ble

Not

app

licab

le

reac

ts w

ith

wat

erTe

trah

ydro

fura

n72

.11

7.06

× 1

0–5 (

25°C

)14

5 (2

0°C

)M

isci

ble

Una

vaila

ble

1, 2

, 4,

5-T

etra

met

hylb

enze

ne13

4.22

2.49

× 1

0–2 (

25°C

)0.

49 (

25°C

)3.

48 (

25°C

)3.

79Te

tran

itrom

etha

ne19

6.03

—13

(25

°C)

——

Tetr

yl28

7.15

<1.

89 ×

10–3

(20

°C)

<1

(20°

C)

0.02

wt.

% (

20°C

)2.

37T

hiop

hene

84

.14

2.93

× 1

0–3 (

25°C

)79

.7 (

25°C

)30

15 (

25°C

)1.

73T

hira

m26

9.35

——

30—

2, 4

-Tol

uene

Dis

ocya

nate

174.

15—

0.0

1 (2

0°C

)N

ot a

pplic

able

reac

ts w

ith

wat

erN

ot a

pplic

able

re

acts

with

w

ater

o-To

luid

ine

107.

161.

88 ×

10–6

(25

°C)

0.1

(20°

C)

15,0

000

(25°

C)

2.61

1, 3

, 5-

Trib

rom

oben

zene

314.

80—

—2.

51 ×

10–6

(25

°C)

4.0

5Tr

ibut

yl P

hosp

hate

266.

32—

—0.

1 w

t. %

(20

°C)

2.29

1, 2

, 3-

Tric

hlor

oben

zene

181.

458.

9 ×

10–3

(20

°C)

1 (4

0°C

)18

.0 (

25°C

)3.

871,

3,

5 Tr

ichl

orob

enze

ne18

1.45

1.9

× 10

–3 (

20°C

)0.

58 (

25°C

)6.

01 (

25°C

)5.

7 (a

vera

ge)

1, 2

, 3-

Tric

hlor

opro

pane

147.

433.

18 ×

10–4

(25

°C)

3.4

(20°

C)

——

1, 1

, 2-

Tric

hlor

otrifl

uoro

etha

ne18

7.38

3.33

× 1

0–1 (

20°C

)27

0 (2

0°C

) 0

.02

wt.

% (

20°C

)2.

59Tr

i-o-C

resy

l Pho

spha

te36

8.37

——

3.1

(25°

C)

3.37

Trie

thyl

amin

e10

1.19

4.79

× 1

0–4 (

20°C

)54

(20

°C)

15,0

00 (

20°C

)U

nava

ilabl

eTr

iflur

alin

335.

294.

84 x

10–5

(23

°C)

1.1

× 10

–4 (

25°C

) 2

40 3

.73

1, 2

, 3-

Trim

ethy

lben

zene

120.

193.

18 x

10–3

(25

°C)

.51

(25°

C)

75.2

(25

°C)

3.34

1, 2

, 4-

Trim

ethy

lben

zene

120.

195.

7 ×

10–3

(25

°C)

2.03

(25

°C)

51.9

(25

°C)

3.57

1, 3

, 5-

Trim

ethy

lben

zene

120.

193.

93 ×

10–3

(25

°C)

2.42

(25

°C)

48.2

(25

°C)

3.21

1, 1

, 3-

Trim

ethy

lcyc

lohe

xane

126.

24—

—1.

77 (

25°C

)U

nava

ilabl

e1,

1,

3-Tr

imet

hylc

yclo

pent

ane

112.

221.

57 (

25°C

)39

.7 (

25°C

)3.

73 (

25°C

)U

nava

ilabl

e2,

2,

5-Tr

imet

hylh

exan

e12

8.26

2.42

(25

°C)

16.5

(25

°C)

1.15

(25

°C)

Una

vaila

ble

2, 2

, 4-

Trim

ethy

lpen

tane

114.

233.

01 (

25°C

)49

.3 (

25°C

)2.

05 (

25°C

)U

nava

ilabl

e2,

3,

4-Tr

imet

hylp

enta

ne11

4.23

2.98

(25

°C)

27.0

(25

°C)

1.36

(25

°C)

Una

vaila

ble

2, 4

, 6-

Trin

itrot

olue

ne22

7.13

—4.

26 ×

10–3

(54

.8°C

)0.

013

wt.

% (

20°C

)2.

48Tr

iphe

nyl P

hosp

hate

326.

295.

88 ×

10–2

(20

–25°

C)

<0.

1 (2

0°C

)0.

001

wt.

% (

20°C

)3.

72



ble

A1

Phy

sica

l P

rop

erti

es o

f S

om

e C

om

mo

n E

nvir

on

men

tal

Co

nta

min

ants

(C

on

tin

ued

)

Co

mp

ou

nd

Mo

lecu

lar

Wei

gh

tH

enry

’s L

aw C

on

stan

t at

m·m

3 /m

ol

Vap

or

Pre

ssu

re

mm

Hg

.S

olu

bili

ty m

g/L

Lo

g K

oc

V

Vin

yl A

ceta

te86

.09

4.81

× 1

0–411

5 (2

5°C

)25

,000

(25

°C)

0.45

Vin

yl C

hlor

ide

62.5

02.

7826

60 (

25°C

)11

00 (

25°C

)0.

39

W

War

farin

308.

33—

—17

(20

°C)

2.9

6

X

o-X

ylen

e10

6.17

0.00

535

(25°

C)

6.6

(25°

C)

213

(25

°C)

2.11

m-X

ylen

e10

6.17

0.0

063

(25°

C)

8.28

7 (2

5°C

)17

3 (2

5°C

)3.

2p-

Xyl

ene

106.

170.

0063

(25

°C)

8.76

3 (2

5°C

)20

0 (2

5°C

)2.

31

Sou

rces

:M

ontg

omer

y, J

. H. a

nd L

. M. W

elko

m,

1990

, G

roun

dwat

er C

hem

ical

s D

esk

Ref

eren

ce,

Lew

is P

ublis

hers

, C

hels

ea,

MI.

Mon

tgom

ery,

J. H

., 19

91,

Gro

undw

ater

Che

mic

als

Des

k R

efer

ence

, Vol

. 2,

Lew

is P

ublis

hers

, C

hels

ea,

MI.


383

APPENDIX

B

Useful Information forBiogeochemical Sampling

Table B1 Factors Used to Correct Atmospheric Pressures

Adjusted to Sea Level

Elevation of Weather Station (feet above sea level)

Value to Subtract (millimeters of mercury)

0 01000 272000 533000 794000 1045000 1286000 151

Source: USGS Groundwater Sampling Field Manual (Chapter 6).

L1282/Appendix B/frame Page 383 Monday, June 18, 2001 11:29 AM


Tab

le B

2S

olu

bili

ty o

f O

xyg

en in

Wat

er a

t Var

iou

s Te

mp

erat

ure

s an

d P

ress

ure

s (T

emp

˚C

, tem

per

atu

re in

deg

rees

Cel

siu

s;

atm

osp

her

ic p

ress

ure

s fr

om

695

to

600

mill

imet

ers

mer

cury

beg

in a

fter

40˚

C)

Tem

p

Atm

osp

her

ic p

ress

ure

, in

mill

imet

ers

of

mer

cury

˚C79

579

078

578

077

577

076

576

075

575

074

574

073

573

072

572

071

571

070

570

0

0.0

15.3

15.2

15.1

15.0

14.9

14.8

14.7

14.6

14.5

14.4

14.3

14.2

14.1

14.0

13.9

13.8

13.7

13.6

13.5

13.4

0.5

15.1

15.0

14.9

14.8

14.7

14.6

14.5

14.4

14.3

14.2

14.1

14.0

13.9

13.8

13.7

13.6

13.5

13.4

13.3

13.2

1.0

14.8

14.7

14.7

14.6

14.5

14.4

14.3

14.2

14.1

14.0

13.9

13.8

13.7

13.6

13.5

13.4

13.3

13.2

13.2

13.1

1.5

14.6

14.5

14.5

14.4

14.3

14.2

14.1

14.0

13.9

13.8

13.7

13.6

13.5

13.4

13.3

13.2

13.2

13.1

13.0

12.9

2.0

14.4

14.3

14.3

14.2

14.1

14.0

13.9

13.8

13.7

13.6

13.5

13.4

13.3

13.3

13.2

13.1

13.0

12.9

12.8

12.7

2.5

14.2

14.2

14.1

14.0

13.9

13.8

13.7

13.6

13.5

13.4

13.3

13.3

13.2

13.1

13.0

12.9

12.8

12.7

12.6

12.5

3.0

14.1

14.0

13.9

13.8

13.4

13.6

13.5

13.4

13.3

13.3

13.2

13.1

13.0

12.9

12.8

12.7

12.6

12.5

12.5

12.4

3.5

13.9

13.8

13.7

13.6

13.5

13.4

13.3

13.3

13.2

13.1

13.0

12.9

12.8

12.7

12.6

12.6

12.5

12.4

12.3

12.2

4.0

13.7

13.6

13.5

13.4

13.3

13.3

13.2

13.1

13.0

12.9

12.8

12.7

12.6

12.6

12.5

12.4

12.3

12.2

12.1

12.0

4.5

13.5

13.4

13.3

13.3

13.2

13.1

13.0

12.9

12.8

12.7

12.7

12.6

12.5

12.4

12.3

12.2

12.1

12.1

12.0

11.9

5.0

13.3

13.3

13.2

13.1

13.0

12.9

12.8

12.7

12.7

12.6

12.5

12.4

12.3

12.2

12.2

12.1

12.0

11.9

11.8

11.7

5.5

13.2

13.1

13.0

12.9

12.8

12.7

12.7

12.6

12.5

12.4

12.3

12.2

12.2

12.1

12.0

11.9

11.8

11.7

11.7

11.6

6.0

13.0

12.9

12.8

12.8

12.7

12.6

12.5

12.4

12.3

12.3

12.2

12.1

12.0

11.9

11.8

11.8

11.7

11.6

11.5

11.4

6.5

12.8

12.8

12.7

12.6

12.5

12.4

12.3

12.3

12.2

12.1

12.0

11.9

11.9

11.8

11.7

11.6

11.5

11.5

11.4

11.3

7.0

12.7

12.6

12.5

12.4

12.4

12.3

12.2

12.1

12.0

12.0

11.9

11.8

11.7

11.6

11.6

11.5

11.4

11.3

11.2

11.1

7.5

12.5

12.4

12.4

12.3

12.2

12.1

12.0

12.0

11.9

11.8

11.7

11.6

11.6

11.5

11.4

11.3

11.3

11.2

11.1

11.0

8.0

12.4

12.3

12.2

12.1

12.1

12.0

11.9

11.8

11.7

11.7

11.6

11.5

11.4

11.3

11.3

11.2

11.1

11.0

11.0

10.9

8.5

12.2

12.1

12.1

12.0

11.9

11.8

11.8

11.7

11.6

11.5

11.4

11.4

11.3

11.2

11.1

11.1

11.0

10.9

10.8

10.7

9.0

12.1

12.0

11.9

11.8

11.8

11.7

11.6

11.5

11.5

11.4

11.3

11.2

11.2

11.1

11.0

10.9

10.8

10.8

10.7

10.6

9.5

11.9

11.9

11.8

11.7

11.6

11.6

11.5

11.4

11.3

11.2

11.2

11.1

11.0

10.9

10.9

10.8

10.7

10.6

10.6

10.5


USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 38510

.011

.811

.711

.611

.611

.511

.411

.311

.311

.211

.111

.011

.010

.910

.810

.710

.710

.610

.510

.410

.410

.511

.711

.611

.511

.411

.411

.311

.211

.111

.111

.010

.910

.810

.810

.710

.610

.510

.510

.410

.310

.211

.011

.511

.411

.411

.311

.211

.211

.111

.010

.910

.910

.810

.710

.610

.610

.510

.410

.310

.310

.210

.111

.511

.411

.311

.211

.211

.111

.011

.010

.910

.810

.710

.710

.610

.510

.410

.410

.310

.210

.210

.110

.012

.011

.311

.211

.111

.011

.010

.910

.810

.810

.710

.610

.510

.510

.410

.310

.310

.210

.110

.010

.09.

9

12.5

11.1

11.1

11.0

10.9

10.8

10.8

10.7

10.6

10.6

10.5

10.4

10.4

10.3

10.2

10.1

10.1

10.0

9.9

9.9

9.8

13.0

11.0

10.9

10.9

10.8

10.7

10.7

10.6

10.5

10.4

10.4

10.3

10.2

10.2

10.1

10.0

10.0

9.9

9.8

9.7

9.7

13.5

10.9

10.8

10.7

10.7

10.6

10.5

10.5

10.4

10.3

10.3

10.2

10.1

10.1

10.0

9.9

9.8

9.8

9.7

9.6

9.6

14.0

10.8

10.7

10.6

10.6

10.5

10.4

10.4

10.3

10.2

10.1

10.1

10.0

9.9

9.9

9.8

9.7

9.7

9.6

9.5

9.5

14.5

10.6

10.6

10.5

10.4

10.4

10.3

10.2

10.2

10.1

10.0

10.0

9.9

9.8

9.8

9.7

9.6

9.6

9.5

9.4

9.4

15.0

10.5

10.5

10.4

10.3

10.3

10.2

10.1

10.1

10.0

9.9

9.9

9.8

9.7

9.7

9.6

9.5

9.5

9.4

9.3

9.3

15.5

10.4

10.4

10.3

10.2

10.2

10.1

10.0

10.0

9.9

9.8

9.8

9.7

9.6

9.6

9.5

9.4

9.4

9.3

9.2

9.2

16.0

10.3

10.2

10.2

10.1

10.0

10.0

9.9

9.8

9.8

9.7

9.7

9.6

9.5

9.5

9.4

9.3

9.3

9.2

9.1

9.1

16.5

10.2

10.1

10.1

10.0

9.9

9.9

9.8

9.7

9.7

9.6

9.5

9.5

9.4

9.4

9.3

9.2

9.2

9.1

9.0

9.0

17.0

10.1

10.0

10.0

9.9

9.8

9.8

9.7

9.6

9.6

9.5

9.4

9.4

9.3

9.3

9.2

9.1

9.1

9.0

8.9

8.9

17.5

10.0

9.9

9.9

9.8

9.7

9.7

9.6

9.5

9.5

9.4

9.3

9.3

9.2

9.2

9.1

9.0

9.0

8.9

8.8

8.8

18.0

9.9

9.8

9.8

9.7

9.6

9.6

9.5

9.4

9.4

9.3

9.3

9.2

9.1

9.1

9.0

8.9

8.9

8.8

8.7

8.7

18.5

9.8

9.7

9.7

9.6

9.5

9.5

9.4

9.3

9.3

9.2

9.2

9.1

9.0

9.0

8.9

8.8

8.8

8.7

8.7

8.6

19.0

9.7

9.6

9.6

9.5

9.4

9.4

9.3

9.3

9.2

9.1

9.1

9.0

8.9

8.9

8.8

8.8

8.7

8.6

8.6

8.5

19.5

9.6

9.5

9.5

9.4

9.3

9.3

9.2

9.2

9.1

9.0

9.0

8.9

8.9

8.8

8.7

8.7

8.6

8.5

8.5

8.4

20.0

9.5

9.4

9.4

9.3

9.3

9.2

9.1

9.1

9.0

8.9

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.5

8.4

8.3

20.5

9.4

9.3

9.3

9.2

9.2

9.1

9.0

9.0

8.9

8.9

8.8

8.7

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.3

21.0

9.3

9.2

9.2

9.1

9.1

9.0

8.9

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.5

8.4

8.4

8.3

8.2

8.2

21.5

9.2

9.2

9.1

9.0

9.0

8.9

8.9

8.8

8.7

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.3

8.2

8.1

8.1

22.0

9.1

9.1

9.0

9.0

8.9

8.8

8.8

8.7

8.7

8.6

8.5

8.5

8.4

8.4

8.3

8.2

8.2

8.1

8.1

8.0



Tab

le B

2S

olu

bili

ty o

f O

xyg

en in

Wat

er a

t Var

iou

s Te

mp

erat

ure

s an

d P

ress

ure

s (T

emp

˚C

, tem

per

atu

re in

deg

rees

Cel

siu

s;

atm

osp

her

ic p

ress

ure

s fr

om

695

to

600

mill

imet

ers

mer

cury

beg

in a

fter

40˚

C)

(Co

nti

nu

ed)

Tem

p

Atm

osp

her

ic p

ress

ure

, in

mill

imet

ers

of

mer

cury

˚C79

579

078

578

077

577

076

576

075

575

074

574

073

573

072

572

071

571

070

570

0

22.5

9.0

9.0

8.9

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.5

8.4

8.3

8.3

8.2

8.2

8.1

8.0

8.0

7.9

23.0

9.0

8.9

8.8

8.8

8.7

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.3

8.2

8.1

8.1

8.0

8.0

7.9

7.9

23.5

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.5

8.4

8.4

8.3

8.2

8.2

8.1

8.1

8.0

8.0

7.9

7.8

7.8

24.0

8.8

8.7

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.9

7.8

7.8

7.7

24.5

8.7

8.7

8.6

8.5

8.5

8.4

8.4

8.3

8.3

8.2

8.1

8.1

8.0

8.0

7.9

7.9

7.8

7.7

7.7

7.6

25.0

8.6

8.6

8.5

8.5

8.4

8.3

8.3

8.2

8.2

8.1

8.1

8.0

8.0

7.9

7.8

7.8

7.7

7.7

7.6

7.6

25.5

8.5

8.5

8.4

8.4

8.3

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.9

7.8

7.8

7.7

7.7

7.6

7.6

7.5

26.0

8.5

8.4

8.4

8.3

8.3

8.2

8.1

8.1

8.0

8.0

7.9

7.9

7.8

7.8

7.7

7.6

7.6

7.5

7.5

7.4

26.5

8.4

8.3

8.3

8.2

8.2

8.1

8.1

8.0

8.0

7.9

7.8

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.4

27.0

8.3

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.9

7.8

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.3

7.3

27.5

8.2

8.2

8.1

8.1

8.0

8.0

7.9

7.9

7.8

7.8

7.7

7.7

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

28.0

8.2

8.1

8.1

8.0

8.0

7.9

7.9

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.2

28.5

8.1

8.0

8.0

7.9

7.9

7.8

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.1

7.1

29.0

8.0

8.0

7.9

7.9

7.8

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

29.5

8.0

7.9

7.9

7.8

7.8

7.7

7.6

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

7.0

30.0

7.9

7.8

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

7.0

6.9

30.5

7.8

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.9

31.0

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.8

31.5

7.7

7.6

7.6

7.5

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

32.0

7.6

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7



.57.

67.

57.

57.

47.

47.

37.

37.

27.

27.

17.

17.

07.

06.

96.

96.

86.

86.

76.

76.

633

.07.

57.

57.

47.

47.

37.

37.

27.

27.

17.

17.

07.

06.

96.

96.

86.

86.

76.

76.

66.

633

.57.

47.

47.

37.

37.

27.

27.

17.

17.

17.

07.

06.

96.

96.

86.

86.

76.

76.

66.

66.

534

.07.

47.

37.

37.

27.

27.

17.

17.

07.

06.

96.

96.

86.

86.

76.

76.

76.

66.

66.

56.

534

.57.

37.

37.

27.

27.

17.

17.

07.

06.

96.

96.

86.

86.

76.

76.

66.

66.

56.

56.

56.

4

35.0

7.3

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

35.5

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

36.0

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

36.5

7.1

7.0

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

37.0

7.0

7.0

6.9

6.9

6.9

6.8

6.8

6.8

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

37.5

7.0

6.9

6.9

6.8

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

38.0

6.9

6.9

6.8

6.8

6.7

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

38.5

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

39.0

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

6.0

39.5

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

6.0

5.9

40.0

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

5.9



Tab

le B

2S

olu

bili

ty o

f O

xyg

en in

Wat

er a

t Var

iou

s Te

mp

erat

ure

s an

d P

ress

ure

s (T

emp

˚C

, tem

per

atu

re in

deg

rees

Cel

siu

s;

atm

osp

her

ic p

ress

ure

s fr

om

695

to

600

mill

imet

ers

mer

cury

beg

in a

fter

40˚

C)

(Co

nti

nu

ed)

Tem

p

Atm

osp

her

ic p

ress

ure

, in

mill

imet

ers

of

mer

cury

˚C69

569

068

568

067

567

066

566

065

565

064

564

063

563

062

562

061

561

060

560

0

0.0

13.3

13.2

13.1

13.0

12.9

12.8

12.8

12.7

12.6

12.5

12.4

12.3

12.2

12.1

12.0

11.9

11.8

11.7

11.6

11.5

0.5

13.1

13.1

13.0

12.9

12.8

12.7

12.6

12.5

12.4

12.3

12.2

12.1

12.0

11.9

11.8

11.7

11.6

11.5

11.4

11.3

1.0

13.0

12.9

12.8

12.7

12.6

12.5

12.4

12.3

12.2

12.1

12.0

11.9

11.8

11.7

11.6

11.6

11.5

11.4

11.3

11.2

1.5

12.8

12.7

12.6

12.5

12.4

12.3

12.2

12.1

12.0

12.0

11.9

11.8

11.7

11.6

11.5

11.4

11.3

11.2

11.1

11.0

2.0

12.6

12.5

12.4

12.3

12.2

12.2

12.1

12.0

11.9

11.8

11.7

11.6

11.5

11.4

11.3

11.2

11.1

11.1

11.0

10.9

2.5

12.4

12.4

12.3

12.2

12.1

12.0

11.9

11.8

11.7

11.6

11.5

11.4

11.4

11.3

11.2

11.1

11.0

10.9

10.8

10.7

3.0

12.3

12.2

12.1

12.0

11.9

11.8

11.7

11.7

11.6

11.5

11.4

11.3

11.2

11.1

11.0

10.9

10.9

10.8

10.7

10.6

3.5

12.1

12.0

11.9

11.8

11.8

11.7

11.6

11.5

11.4

11.3

11.2

11.1

11.1

11.0

10.9

10.8

10.7

10.6

10.5

10.4

4.0

12.0

11.9

11.8

11.7

11.6

11.5

11.4

11.3

11.3

11.2

11.1

11.0

10.9

10.8

10.7

10.7

10.6

10.5

10.4

10.3

4.5

11.8

11.7

11.9

11.5

11.5

11.4

11.3

11.2

11.1

11.0

10.9

10.9

10.8

10.7

10.6

10.5

10.4

10.3

10.3

10.2

5.0

11.6

11.6

11.5

11.4

11.3

11.2

11.1

11.1

11.0

10.9

10.8

10.7

10.6

10.5

10.5

10.4

10.3

10.2

10.1

10.0

5.5

11.5

11.4

11.3

11.2

11.2

11.1

11.0

10.9

10.8

10.7

10.7

10.6

10.5

10.4

10.3

10.2

10.2

10.1

10.0

9.9

6.0

11.4

11.3

11.2

11.1

11.0

10.9

10.9

10.8

10.7

10.6

10.5

10.4

10.4

10.3

10.2

10.1

10.0

9.9

9.9

9.8

6.5

11.2

11.1

11.0

11.0

10.9

10.8

10.7

10.6

10.6

10.5

10.4

10.3

10.2

10.1

10.1

10.0

9.9

9.8

9.7

9.7

7.0

11.1

11.0

10.9

10.8

10.7

10.7

10.6

10.5

10.4

10.3

10.3

10.2

10.1

10.0

9.9

9.9

9.8

9.7

9.6

9.5

7.5

10.9

10.9

10.8

10.7

10.6

10.5

10.5

10.4

10.3

10.2

10.1

10.1

10.0

9.9

9.8

9.7

9.7

9.6

9.5

9.4

8.0

10.8

10.7

10.6

10.6

10.5

10.4

10.3

10.2

10.2

10.1

10.0

9.9

9.9

9.8

9.7

9.6

9.5

9.5

9.4

9.3

8.5

10.7

10.6

10.5

10.4

10.4

10.3

10.2

10.1

10.0

10.0

9.9

9.8

9.7

9.7

9.6

9.5

9.4

9.3

9.3

9.2

9.0

10.5

10.5

10.4

10.3

10.2

10.2

10.1

10.0

9.9

9.8

9.8

9.7

9.6

9.5

9.5

9.4

9.3

9.2

9.2

9.1

9.5

10.4

10.3

10.3

10.2

10.1

10.0

10.0

9.9

9.8

9.7

9.7

9.6

9.5

9.4

9.4

9.3

9.2

9.1

9.0

9.0



.010

.310

.210

.110

.110

.09.

99.

89.

89.

79.

69.

59.

59.

49.

39.

29.

29.

19.

08.

98.

910

.510

.210

.110

.09.

99.

99.

89.

79.

79.

69.

59.

49.

49.

39.

29.

19.

19.

08.

98.

88.

811

.010

.110

.09.

99.

89.

89.

79.

69.

59.

59.

49.

39.

29.

29.

19.

09.

08.

98.

88.

78.

711

.59.

99.

99.

89.

79.

69.

69.

59.

49.

49.

39.

29.

19.

19.

08.

98.

88.

88.

78.

68.

612

.09.

89.

89.

79.

69.

59.

59.

49.

39.

29.

29.

19.

09.

08.

98.

88.

78.

78.

68.

58.

5

12.5

9.7

9.6

9.6

9.5

9.4

9.4

9.3

9.2

9.1

9.1

9.0

8.9

8.9

8.8

8.7

8.6

8.6

8.5

8.4

8.4

13.0

9.6

9.5

9.5

9.4

9.3

9.3

9.2

9.1

9.0

9.0

8.9

8.8

8.8

8.7

8.6

8.5

8.5

8.4

8.3

8.3

13.5

9.5

9.4

9.4

9.3

9.2

9.1

9.1

9.0

8.9

8.9

8.8

8.7

8.7

8.6

8.5

8.5

8.4

8.3

8.2

8.2

14.0

9.4

9.3

9.3

9.2

9.1

9.0

9.0

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.2

8.2

8.1

14.5

9.3

9.2

9.2

9.1

9.0

8.9

8.9

8.8

8.7

8.7

8.6

8.5

8.5

8.4

8.3

8.3

8.2

8.1

8.1

8.0

15.0

9.2

9.1

9.1

9.0

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.2

8.2

8.1

8.0

8.0

7.9

15.5

9.1

9.0

9.0

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.8

16.0

9.0

8.9

8.9

8.8

8.7

8.7

8.6

8.5

8.5

8.4

8.3

8.3

8.2

8.1

8.1

8.0

7.9

7.9

7.8

7.7

16.5

8.9

8.8

8.8

8.7

8.6

8.6

8.5

8.4

8.4

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.8

7.8

7.7

7.7

17.0

8.8

8.7

8.7

8.6

8.5

8.5

8.4

8.3

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.8

7.8

7.7

7.6

7.6

17.5

8.7

8.6

8.6

8.5

8.5

8.4

8.3

8.3

8.2

8.1

8.1

8.0

7.9

7.9

7.8

7.7

7.7

7.6

7.6

7.5

18.0

8.6

8.6

8.5

8.4

8.4

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.9

7.8

7.7

7.7

7.6

7.5

7.5

7.4

18.5

8.5

8.5

8.4

8.3

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.8

7.8

7.7

7.7

7.6

7.5

7.5

7.4

7.3

19.0

8.4

8.4

8.3

8.3

8.2

8.1

8.1

8.0

7.9

7.9

7.8

7.8

7.7

7.6

7.6

7.5

7.4

7.4

7.3

7.3

19.5

8.4

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.9

7.8

7.7

7.7

7.6

7.6

7.5

7.4

7.4

7.3

7.2

7.2

20.0

8.3

8.2

8.2

8.1

8.0

8.0

7.9

7.8

7.8

7.7

7.7

7.6

7.5

7.5

7.4

7.4

7.3

7.2

7.2

7.1

20.5

8.2

8.1

8.1

8.0

7.9

7.9

7.8

7.8

7.7

7.6

7.6

7.5

7.5

7.4

7.3

7.3

7.2

7.2

7.1

7.0

21.0

8.1

8.1

8.0

7.9

7.9

7.8

7.8

7.7

7.6

7.6

7.5

7.5

7.4

7.3

7.3

7.2

7.2

7.1

7.0

7.0

21.5

8.0

8.0

7.9

7.9

7.8

7.7

7.7

7.6

7.6

7.5

7.4

7.4

7.3

7.3

7.2

7.1

7.1

7.0

7.0

6.9

22.0

8.0

7.9

7.8

7.8

7.7

7.7

7.6

7.5

7.5

7.4

7.4

7.3

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.8



Tab

le B

2S

olu

bili

ty o

f O

xyg

en in

Wat

er a

t Var

iou

s Te

mp

erat

ure

s an

d P

ress

ure

s (T

emp

˚C

, tem

per

atu

re in

deg

rees

Cel

siu

s;

atm

osp

her

ic p

ress

ure

s fr

om

695

to

600

mill

imet

ers

mer

cury

beg

in a

fter

40˚

C)

(Co

nti

nu

ed)

Tem

p

Atm

osp

her

ic p

ress

ure

, in

mill

imet

ers

of

mer

cury

˚C69

569

068

568

067

567

066

566

065

565

064

564

063

563

062

562

061

561

060

560

0

22.5

7.9

7.8

7.8

7.7

7.6

7.6

7.5

7.5

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

6.9

6.9

6.8

6.8

23.0

7.8

7.7

7.7

7.6

7.6

7.5

7.5

7.4

7.3

7.3

7.2

7.2

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

23.5

7.7

7.7

7.6

7.6

7.5

7.4

7.4

7.3

7.3

7.2

7.2

7.1

7.0

7.0

6.9

6.9

6.8

6.7

6.7

6.6

24.0

7.7

7.6

7.5

7.5

7.4

7.4

7.3

7.3

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.7

6.7

6.6

6.6

24.5

7.6

7.5

7.5

7.4

7.4

7.3

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

25.0

7.5

7.5

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.4

25.5

7.4

7.4

7.3

7.3

7.2

7.2

7.1

7.1

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.4

6.4

26.0

7.4

7.3

7.3

7.2

7.2

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.5

6.5

6.4

6.4

6.3

26.5

7.3

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

27.0

7.2

7.2

7.1

7.1

7.0

7.0

6.9

6.9

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

27.5

7.2

7.1

7.1

7.0

7.0

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

28.0

7.1

7.1

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.1

6.1

28.5

7.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.2

6.2

6.1

6.1

6.0

29.0

7.0

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.2

6.2

6.1

6.1

6.0

6.0

29.5

6.9

6.9

6.8

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

30.0

6.9

6.8

6.8

6.7

6.7

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

30.5

6.8

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

5.8

31.0

6.7

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

31.5

6.7

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

5.7

32.0

6.6

6.6

6.5

6.5

6.4

6.4

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7



.56.

66.

56.

56.

46.

46.

36.

36.

26.

26.

16.

16.

06.

05.

95.

95.

85.

85.

75.

75.

633

.06.

56.

56.

46.

46.

36.

36.

26.

26.

16.

16.

06.

05.

95.

95.

85.

85.

75.

75.

65.

633

.56.

56.

46.

46.

36.

36.

26.

26.

16.

16.

06.

05.

95.

95.

85.

85.

75.

75.

65.

65.

534

.06.

46.

46.

36.

36.

26.

26.

16.

16.

06.

05.

95.

95.

85.

85.

75.

75.

65.

65.

55.

534

.56.

46.

36.

36.

26.

26.

16.

16.

06.

05.

95.

95.

85.

85.

75.

75.

65.

65.

55.

55.

4

35.0

6.3

6.3

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

35.5

6.2

6.2

6.2

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.3

36.0

6.2

6.1

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.3

5.3

36.5

6.1

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.3

5.3

5.2

37.0

6.1

6.1

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.3

5.3

5.3

5.2

37.5

6.0

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.3

5.3

5.3

5.2

5.2

38.0

6.0

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.3

5.3

5.3

5.2

5.2

5.1

38.5

6.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.4

5.3

5.3

5.2

5.2

5.1

5.1

39.0

5.9

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.4

5.3

5.3

5.2

5.2

5.1

5.1

5.0

39.5

5.9

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.4

5.3

5.3

5.2

5.2

5.1

5.1

5.0

5.0

40.0

5.8

5.8

5.7

5.7

5.6

5.6

5.5

5.5

5.4

5.4

5.4

5.3

5.3

5.2

5.2

5.1

5.1

5.0

5.0

5.0

Sou

rce:

US

GS

Gro

undw

ater

Sam

plin

g F

ield

Man

ual (

Cha

pter

6).



Tab

le B

3S

alin

ity

Co

rrec

tio

n F

acto

rs f

or

Dis

solv

ed O

xyg

en i

n W

ater

(b

ased

on

co

nd

uct

ivit

y) (

Tem

p °

C, t

emp

erat

ure

in

deg

rees

Cel

siu

s;

salin

ity

corr

ecti

on

fac

tors

at

30 t

o 3

5°C

are

sh

ow

n a

t th

e en

d o

f th

is t

able

)

Tem

pC

on

du

ctiv

ity,

in

mic

rosi

emen

s p

er c

enti

met

er a

t 25

°C˚C

010

0020

0030

0040

0050

0060

0070

0080

0090

0010

000

1100

012

000

1300

014

000

1500

016

000

0.0

1.00

00.

996

0.99

20.

989

0.98

50.

981

0.97

70.

973

0.96

90.

965

0.96

10.

957

0.95

30.

950

0.94

60.

942

0.93

81.

01.

000

0.99

60.

992

0.98

90.

985

0.98

10.

977

0.97

30.

969

0.96

50.

962

0.95

80.

954

0.95

00.

946

0.94

20.

938

2.0

1.00

00.

996

0.99

20.

989

0.98

50.

981

0.97

70.

973

0.97

00.

966

0.96

20.

958

0.95

40.

950

0.94

60.

942

0.93

83.

01.

000

0.99

60.

993

0.98

90.

985

0.98

10.

977

0.97

40.

970

0.96

60.

962

0.95

80.

954

0.95

10.

947

0.94

30.

939

4.0

1.00

00.

996

0.99

30.

989

0.98

50.

981

0.97

80.

974

0.97

00.

966

0.96

20.

959

0.95

50.

951

0.94

70.

943

0.93

9

5.0

1.00

00.

996

0.99

30.

989

0.98

50.

981

0.97

80.

974

0.97

00.

966

0.96

30.

959

0.95

50.

951

0.94

70.

944

0.94

06.

01.

000

0.99

60.

993

0.98

90.

985

0.98

20.

978

0.97

40.

970

0.96

70.

963

0.95

90.

955

0.95

20.

948

0.94

40.

940

7.0

1.00

00.

996

0.99

30.

989

0.98

50.

982

0.97

80.

974

0.97

10.

967

0.96

30.

959

0.95

60.

952

0.94

80.

944

0.94

18.

01.

000

0.99

60.

993

0.98

90.

986

0.98

20.

978

0.97

50.

971

0.96

70.

963

0.96

00.

956

0.95

20.

949

0.94

50.

941

9.0

1.00

00.

996

0.99

30.

989

0.98

60.

982

0.97

80.

975

0.97

10.

967

0.96

40.

960

0.95

60.

953

0.94

90.

945

0.94

1

10.0

1.00

00.

996

0.99

30.

989

0.98

60.

982

0.97

90.

975

0.97

10.

968

0.96

40.

960

0.95

70.

953

0.94

90.

946

0.94

211

.01.

000

0.99

60.

993

0.98

90.

986

0.98

20.

979

0.97

50.

971

0.96

80.

964

0.96

10.

957

0.95

30.

950

0.94

60.

942

12.0

1.00

00.

997

0.99

30.

989

0.98

60.

982

0.97

90.

975

0.97

20.

968

0.96

50.

961

0.95

70.

954

0.95

00.

946

0.94

313

.01.

000

0.99

70.

993

0.99

00.

986

0.98

30.

979

0.97

50.

972

0.96

80.

965

0.96

10.

958

0.95

40.

950

0.94

70.

943

14.0

1.00

00.

997

0.99

30.

990

0.98

60.

983

0.97

90.

976

0.97

20.

969

0.96

50.

961

0.95

80.

954

0.95

10.

947

0.94

3

15.0

1.00

00.

997

0.99

30.

990

0.98

60.

983

0.97

90.

976

0.97

20.

969

0.96

50.

962

0.95

80.

955

0.95

10.

947

0.94

416

.01.

000

0.99

70.

993

0.99

00.

986

0.98

30.

979

0.97

60.

972

0.96

90.

966

0.96

20.

958

0.95

50.

951

0.94

80.

944

17.0

1.00

00.

997

0.99

30.

990

0.98

60.

983

0.98

00.

976

0.97

30.

969

0.96

60.

962

0.95

90.

955

0.95

20.

948

0.94

518

.01.

000

0.99

70.

993

0.99

00.

987

0.98

30.

980

0.97

60.

973

0.96

90.

966

0.96

30.

959

0.95

60.

952

0.94

90.

945

19.0

1.00

00.

997

0.99

30.

990

0.98

70.

983

0.98

00.

976

0.97

30.

970

0.96

60.

963

0.95

90.

956

0.95

20.

949

0.94

5



.01.

000

0.99

70.

993

0.99

00.

987

0.98

30.

980

0.97

70.

973

0.97

00.

966

0.96

30.

960

0.95

60.

953

0.94

90.

946

21.0

1.00

00.

997

0.99

30.

990

0.98

70.

984

0.98

00.

977

0.97

30.

970

0.96

70.

963

0.96

00.

957

0.95

30.

950

0.94

622

.01.

000

0.99

70.

993

0.99

00.

987

0.98

40.

980

0.97

70.

974

0.97

00.

967

0.96

40.

960

0.95

70.

953

0.95

00.

947

23.0

1.00

00.

997

0.99

40.

990

0.98

70.

984

0.98

00.

977

0.97

40.

971

0.96

70.

964

0.96

00.

957

0.95

40.

950

0.94

724

.01.

000

0.99

70.

994

0.99

00.

987

0.98

40.

981

0.97

70.

974

0.97

10.

967

0.96

40.

961

0.95

70.

954

0.95

10.

947

25.0

1.00

00.

997

0.99

40.

990

0.98

70.

984

0.98

10.

977

0.97

40.

971

0.96

80.

964

0.96

10.

958

0.95

40.

951

0.94

826

.01.

000

0.99

70.

994

0.99

00.

987

0.98

40.

981

0.97

80.

974

0.97

10.

968

0.96

50.

961

0.95

80.

955

0.95

10.

948

27.0

1.00

00.

997

0.99

40.

991

0.98

70.

984

0.98

10.

978

0.97

50.

971

0.96

80.

965

0.96

20.

958

0.95

50.

952

0.94

828

.01.

000

0.99

70.

994

0.99

10.

987

0.98

40.

981

0.97

80.

975

0.97

20.

968

0.96

50.

962

0.95

90.

955

0.95

20.

949

29.0

1.00

00.

997

0.99

40.

991

0.98

80.

984

0.98

10.

978

0.97

50.

972

0.96

90.

965

0.96

20.

959

0.95

60.

952

0.94

9

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C17

000

1800

019

000

2000

021

000

2200

023

000

2400

025

000

2600

027

000

2800

029

000

3000

031

000

3200

033

000

0.0

0.93

40.

930

0.92

60.

922

0.91

80.

914

0.91

00.

905

0.90

10.

897

0.89

30.

889

0.88

50.

881

0.87

70.

873

0.86

91.

00.

934

0.93

00.

926

0.92

20.

918

0.91

40.

910

0.90

60.

902

0.89

80.

894

0.89

00.

886

0.88

20.

878

0.87

40.

870

2.0

0.93

50.

931

0.92

70.

923

0.91

90.

915

0.91

10.

907

0.90

30.

899

0.89

50.

891

0.88

70.

883

0.87

90.

875

0.87

13.

00.

935

0.93

10.

927

0.92

30.

919

0.91

50.

911

0.90

70.

903

0.89

90.

895

0.89

10.

887

0.88

30.

879

0.87

50.

871

4.0

0.93

50.

932

0.92

80.

924

0.92

00.

916

0.91

20.

908

0.90

40.

900

0.89

60.

892

0.88

80.

884

0.88

00.

876

0.87

2

5.0

0.93

60.

932

0.92

80.

924

0.92

00.

917

0.91

30.

909

0.90

50.

901

0.89

70.

893

0.88

90.

885

0.88

10.

877

0.87

36.

00.

936

0.93

30.

929

0.92

50.

921

0.91

70.

913

0.90

90.

905

0.90

20.

898

0.89

40.

890

0.88

60.

882

0.87

80.

874

7.0

0.93

70.

933

0.92

90.

925

0.92

20.

918

0.91

40.

910

0.90

60.

902

0.89

80.

894

0.89

10.

887

0.88

30.

879

0.87

58.

00.

937

0.93

30.

930

0.92

60.

922

0.91

80.

914

0.91

10.

907

0.90

30.

899

0.89

50.

891

0.88

70.

884

0.88

00.

876

9.0

0.93

70.

934

0.93

00.

926

0.92

20.

919

0.91

50.

911

0.90

70.

904

0.90

00.

896

0.89

20.

888

0.88

40.

880

0.87

7

10.0

0.93

80.

934

0.93

10.

927

0.92

30.

919

0.91

60.

912

0.90

80.

904

0.90

00.

897

0.89

30.

889

0.88

50.

881

0.87

711

.00.

939

0.93

50.

931

0.92

70.

924

0.92

00.

916

0.91

20.

909

0.90

50.

901

0.89

70.

894

0.89

00.

886

0.88

20.

878

12.0

0.93

90.

935

0.93

20.

928

0.92

40.

920

0.91

70.

913

0.90

90.

906

0.90

30.

898

0.89

40.

890

0.88

70.

883

0.87

913

.00.

939

0.93

60.

932

0.92

80.

925

0.92

10.

917

0.91

40.

910

0.90

60.

902

0.89

90.

895

0.89

10.

887

0.88

40.

880

14.0

0.94

00.

936

0.93

30.

929

0.92

50.

922

0.91

80.

914

0.91

10.

907

0.90

30.

899

0.89

60.

892

0.88

80.

884

0.88

1



Tab

le B

3S

alin

ity

Co

rrec

tio

n F

acto

rs f

or

Dis

solv

ed O

xyg

en i

n W

ater

(b

ased

on

co

nd

uct

ivit

y) (

Tem

p °

C, t

emp

erat

ure

in

deg

rees

Cel

siu

s;

salin

ity

corr

ecti

on

fac

tors

at

30 t

o 3

5°C

are

sh

ow

n a

t th

e en

d o

f th

is t

able

) (C

on

tin

ued

)

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C17

000

1800

019

000

2000

021

000

2200

023

000

2400

025

000

2600

027

000

2800

029

000

3000

031

000

3200

033

000

15.0

0.94

00.

937

0.93

30.

929

0.92

60.

922

0.91

80.

915

0.91

10.

908

0.90

40.

900

0.89

60.

893

0.88

90.

885

0.88

216

.00.

941

0.93

70.

934

0.93

00.

926

0.92

30.

919

0.91

50.

912

0.90

90.

904

0.90

10.

897

0.89

30.

890

0.88

60.

882

17.0

0.94

10.

938

0.93

40.

930

0.92

70.

923

0.92

00.

916

0.91

20.

909

0.90

50.

901

0.89

80.

894

0.89

10.

887

0.88

318

.00.

942

0.93

80.

934

0.93

10.

927

0.92

40.

920

0.91

70.

913

0.91

00.

906

0.90

20.

899

0.89

50.

891

0.88

80.

884

19.0

0.94

20.

938

0.93

50.

931

0.92

80.

924

0.92

10.

917

0.91

40.

911

0.90

60.

903

0.89

90.

896

0.89

20.

888

0.88

5

20.0

0.94

20.

939

0.93

50.

932

0.92

80.

925

0.92

10.

918

0.91

40.

911

0.90

70.

903

0.90

00.

896

0.89

30.

889

0.88

621

.00.

943

0.93

90.

936

0.93

20.

929

0.92

50.

922

0.91

80.

915

0.91

10.

908

0.90

40.

901

0.89

70.

893

0.89

00.

886

22.0

0.94

30.

940

0.93

60.

933

0.92

90.

926

0.92

20.

919

0.91

50.

912

0.90

80.

905

0.90

10.

898

0.89

40.

891

0.88

723

.00.

944

0.94

00.

937

0.93

30.

930

0.92

60.

923

0.91

90.

916

0.91

20.

909

0.90

50.

902

0.89

80.

895

0.89

10.

888

24.0

0.94

40.

941

0.93

70.

934

0.93

00.

927

0.92

30.

920

0.91

70.

913

0.91

00.

906

0.90

30.

899

0.89

60.

892

0.88

9

25.0

0.94

40.

941

0.93

80.

934

0.93

10.

927

0.92

40.

921

0.91

70.

914

0.91

00.

907

0.90

30.

900

0.89

60.

893

0.89

026

.00.

945

0.94

10.

938

0.93

50.

931

0.92

80.

925

0.92

10.

918

0.91

40.

911

0.90

70.

904

0.90

10.

897

0.89

40.

890

27.0

0.94

50.

942

0.93

80.

935

0.93

20.

928

0.92

50.

922

0.91

80.

915

0.91

10.

908

0.90

50.

901

0.89

80.

894

0.89

128

.00.

946

0.94

20.

939

0.93

60.

932

0.92

90.

926

0.92

20.

919

0.91

50.

912

0.90

90.

905

0.90

20.

898

0.89

50.

892

29.0

0.94

60.

943

0.93

90.

936

0.93

30.

929

0.92

60.

923

0.91

90.

916

0.91

30.

909

0.90

60.

903

0.89

90.

896

0.89

2

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C34

000

3500

036

000

3700

038

000

3900

040

000

4100

042

000

4300

044

000

4500

046

000

4700

048

000

4900

050

000

0.0

0.86

50.

861

0.85

60.

852

0.84

80.

844

0.84

00.

836

0.83

20.

828

0.82

30.

819

0.81

50.

811

0.80

70.

803

0.79

91.

00.

866

0.86

20.

857

0.85

30.

849

0.84

50.

841

0.83

70.

833

0.82

90.

825

0.82

10.

816

0.81

20.

808

0.80

40.

800

2.0

0.86

70.

862

0.85

80.

854

0.85

00.

846

0.84

20.

838

0.83

40.

830

0.82

60.

822

0.81

80.

814

0.80

90.

805

0.80

13.

00.

867

0.86

30.

859

0.85

50.

851

0.84

70.

843

0.83

90.

835

0.83

10.

827

0.82

30.

819

0.81

50.

811

0.80

70.

803

4.0

0.86

80.

864

0.86

00.

856

0.85

20.

848

0.84

40.

840

0.83

60.

832

0.82

80.

824

0.82

00.

816

0.81

20.

808

0.80

4


USEFUL INFORMATION FOR BIOGEOCHEMICAL SAMPLING 3955.

00.

869

0.86

50.

861

0.85

70.

853

0.84

90.

845

0.84

10.

837

0.83

30.

829

0.82

50.

821

0.81

70.

813

0.80

90.

805

6.0

0.87

00.

866

0.86

20.

858

0.85

40.

850

0.84

60.

842

0.83

80.

834

0.83

00.

826

0.82

20.

818

0.81

40.

810

0.80

67.

00.

871

0.86

70.

863

0.85

90.

855

0.85

10.

847

0.84

30.

839

0.83

50.

831

0.82

80.

824

0.82

00.

816

0.81

20.

808

8.0

0.87

20.

868

0.86

40.

860

0.85

60.

852

0.84

80.

844

0.84

00.

837

0.83

30.

829

0.82

50.

821

0.81

70.

813

0.80

99.

00.

873

0.86

90.

865

0.86

10.

857

0.85

30.

849

0.84

50.

842

0.83

80.

834

0.83

00.

826

0.82

20.

818

0.81

40.

810

10.0

0.87

40.

870

0.86

60.

862

0.85

80.

854

0.85

00.

846

0.84

30.

839

0.83

50.

831

0.82

70.

823

0.81

90.

815

0.81

111

.00.

874

0.87

10.

867

0.86

30.

859

0.85

50.

851

0.84

80.

844

0.84

00.

836

0.83

20.

828

0.82

40.

820

0.81

70.

813

12.0

0.87

50.

871

0.86

80.

864

0.86

00.

856

0.85

20.

849

0.84

50.

841

0.83

70.

833

0.82

90.

825

0.82

20.

818

0.81

413

.00.

876

0.87

20.

869

0.86

50.

861

0.85

70.

853

0.85

00.

846

0.84

20.

838

0.83

40.

830

0.82

70.

823

0.81

90.

815

14.0

0.87

70.

873

0.86

90.

866

0.86

20.

858

0.85

40.

851

0.84

70.

843

0.83

90.

835

0.83

20.

828

0.82

40.

820

0.81

6

15.0

0.87

80.

874

0.87

00.

867

0.86

30.

859

0.85

50.

852

0.84

80.

844

0.84

00.

836

0.83

30.

829

0.82

50.

821

0.81

716

.00.

879

0.87

50.

871

0.86

70.

864

0.86

00.

856

0.85

30.

849

0.84

50.

841

0.83

80.

834

0.83

00.

826

0.82

20.

819

17.0

0.87

90.

876

0.87

20.

868

0.86

50.

861

0.85

70.

854

0.85

00.

846

0.84

20.

839

0.83

50.

831

0.82

70.

824

0.82

018

.00.

880

0.87

70.

873

0.86

90.

866

0.86

20.

858

0.85

50.

851

0.84

70.

843

0.84

00.

836

0.83

20.

829

0.82

50.

821

19.0

0.88

10.

877

0.87

40.

870

0.86

70.

863

0.85

90.

855

0.85

20.

848

0.84

40.

841

0.83

70.

833

0.83

00.

826

0.82

2

20.0

0.88

20.

878

0.87

50.

871

0.86

70.

864

0.86

00.

856

0.85

30.

849

0.84

50.

842

0.83

80.

834

0.83

10.

827

0.82

321

.00.

883

0.87

90.

876

0.87

20.

868

0.86

50.

861

0.85

70.

854

0.85

00.

846

0.84

30.

839

0.83

60.

832

0.82

80.

825

22.0

0.88

40.

880

0.87

60.

873

0.86

90.

866

0.86

20.

858

0.85

50.

851

0.84

80.

844

0.84

00.

837

0.83

30.

829

0.82

623

.00.

884

0.88

10.

877

0.87

40.

870

0.86

60.

863

0.85

90.

856

0.85

20.

849

0.84

50.

841

0.83

80.

834

0.83

00.

827

24.0

0.88

50.

882

0.87

80.

874

0.87

10.

867

0.86

40.

860

0.85

70.

853

0.85

00.

846

0.84

20.

839

0.83

50.

832

0.82

8

25.0

0.88

60.

882

0.87

90.

875

0.87

20.

868

0.86

50.

861

0.85

80.

854

0.85

10.

847

0.84

30.

840

0.83

60.

833

0.82

926

.00.

887

0.88

30.

880

0.87

60.

873

0.86

90.

866

0.86

20.

859

0.85

50.

852

0.84

80.

844

0.84

10.

837

0.83

40.

830

27.0

0.88

70.

884

0.88

00.

877

0.87

40.

870

0.86

70.

863

0.86

00.

856

0.85

30.

849

0.84

50.

842

0.83

80.

835

0.83

128

.00.

888

0.88

50.

881

0.87

80.

874

0.87

10.

867

0.86

40.

860

0.85

70.

853

0.85

00.

846

0.84

30.

839

0.83

60.

832

29.0

0.88

90.

886

0.88

20.

879

0.87

50.

872

0.86

80.

865

0.86

10.

858

0.85

40.

851

0.84

80.

844

0.84

10.

837

0.83

4



Tab

le B

3S

alin

ity

Co

rrec

tio

n F

acto

rs f

or

Dis

solv

ed O

xyg

en i

n W

ater

(b

ased

on

co

nd

uct

ivit

y) (

Tem

p °

C, t

emp

erat

ure

in

deg

rees

Cel

siu

s;

salin

ity

corr

ecti

on

fac

tors

at

30 t

o 3

5°C

are

sh

ow

n a

t th

e en

d o

f th

is t

able

) (C

on

tin

ued

)

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C51

000

5200

053

000

5400

055

000

5600

057

000

5800

059

000

6000

061

000

6200

063

000

6400

065

000

6600

067

000

0.0

0.79

50.

790

0.78

60.

782

0.77

80.

774

0.77

00.

767

0.76

10.

757

0.75

30.

749

0.74

50.

741

0.73

70.

732

0.72

81.

00.

796

0.79

20.

788

0.78

30.

779

0.77

50.

771

0.76

70.

763

0.75

90.

755

0.75

10.

746

0.74

20.

738

0.73

40.

730

2.0

0.79

70.

793

0.78

90.

785

0.78

10.

777

0.77

30.

768

0.76

40.

760

0.75

60.

752

0.74

80.

744

0.74

00.

736

0.73

23.

00.

798

0.79

40.

790

0.78

60.

782

0.77

80.

774

0.77

00.

766

0.76

20.

758

0.75

40.

750

0.74

60.

741

0.73

70.

733

4.0

0.80

00.

796

0.79

20.

788

0.79

40.

780

0.77

50.

771

0.76

70.

763

0.75

90.

755

0.75

10.

747

0.74

30.

739

0.73

5

5.0

0.80

10.

797

0.79

30.

789

0.78

50.

781

0.77

70.

773

0.76

90.

765

0.76

10.

757

0.75

30.

749

0.74

50.

741

0.73

76.

00.

802

0.79

80.

794

0.79

00.

786

0.78

20.

778

0.77

40.

770

0.76

60.

762

0.75

80.

754

0.75

00.

746

0.74

20.

738

7.0

0.80

40.

800

0.79

60.

792

0.78

80.

784

0.78

00.

776

0.77

20.

768

0.76

40.

760

0.75

60.

752

0.74

80.

744

0.74

08.

00.

805

0.80

10.

797

0.79

30.

789

0.78

50.

781

0.77

70.

773

0.76

90.

765

0.76

10.

757

0.75

30.

749

0.74

50.

742

9.0

0.80

60.

802

0.79

80.

794

0.79

00.

787

0.78

30.

779

0.77

50.

771

0.76

70.

763

0.75

90.

755

0.75

10.

747

0.74

3

10.0

0.80

70.

804

0.80

00.

796

0.79

20.

788

0.78

40.

780

0.77

60.

772

0.76

80.

764

0.76

00.

757

0.75

30.

749

0.74

511

.00.

809

0.80

50.

801

0.79

70.

793

0.78

90.

785

0.78

10.

778

0.77

40.

770

0.76

60.

762

0.75

80.

754

0.75

00.

746

12.0

0.81

00.

806

0.80

20.

798

0.79

40.

791

0.78

70.

783

0.77

90.

775

0.77

10.

767

0.76

30.

760

0.75

60.

752

0.74

813

.00.

811

0.80

70.

804

0.80

00.

796

0.79

20.

788

0.78

40.

780

0.77

70.

773

0.76

90.

765

0.76

10.

757

0.75

30.

750

14.0

0.81

20.

809

0.80

50.

801

0.79

70.

793

0.78

90.

786

0.78

20.

778

0.77

40.

770

0.76

60.

763

0.75

90.

755

0.75

1

15.0

0.81

40.

810

0.80

60.

802

0.79

80.

795

0.79

10.

787

0.78

30.

779

0.77

60.

772

0.76

80.

764

0.76

00.

756

0.75

316

.00.

815

0.81

10.

907

0.80

40.

800

0.79

60.

792

0.78

80.

785

0.78

40.

777

0.77

30.

769

0.76

60.

762

0.75

80.

754

17.0

0.81

60.

812

0.80

90.

805

0.80

10.

797

0.79

40.

790

0.78

60.

782

0.77

80.

775

0.77

10.

767

0.76

30.

760

0.75

618

.00.

817

0.81

40.

810

0.80

60.

802

0.79

90.

795

0.79

10.

787

0.78

40.

780

0.77

60.

772

0.76

90.

765

0.76

10.

757

19.0

0.81

90.

815

0.81

10.

807

0.80

40.

800

0.79

60.

792

0.78

90.

785

0.78

10.

777

0.77

40.

770

0.76

60.

763

0.75

9



.00.

820

0.81

60.

812

0.80

90.

805

0.90

10.

797

0.79

40.

790

0.78

60.

783

0.77

90.

775

0.77

10.

768

0.76

40.

760

21.0

0.82

10.

817

0.81

40.

810

0.80

60.

802

0.79

90.

795

0.79

10.

788

0.78

40.

780

0.77

70.

773

0.76

90.

766

0.76

222

.00.

822

0.81

80.

815

0.81

10.

807

0.80

40.

800

0.79

60.

783

0.78

90.

785

0.78

20.

778

0.77

40.

771

0.76

70.

763

23.0

0.82

30.

820

0.81

60.

812

0.80

90.

805

0.80

10.

798

0.79

40.

790

0.78

70.

783

0.77

90.

776

0.77

20.

768

0.76

524

.00.

824

0.82

10.

817

0.81

40.

810

0.80

60.

803

0.79

90.

795

0.79

20.

788

0.78

50.

781

0.77

70.

774

0.77

00.

766

25.0

0.82

60.

822

0.81

80.

815

0.81

10.

808

0.80

40.

800

0.79

70.

793

0.78

90.

786

0.78

20.

779

0.77

50.

771

0.76

826

.00.

827

0.82

30.

820

0.81

60.

812

0.80

90.

805

0.80

20.

798

0.79

40.

791

0.78

70.

784

0.78

00.

776

0.77

30.

769

27.0

0.82

80.

824

0.82

10.

817

0.81

40.

810

0.80

60.

803

0.79

90.

792

0.78

90.

785

0.78

10.

778

0.77

40.

771

28.0

0.82

90.

825

0.82

20.

818

0.81

50.

811

0.80

80.

804

0.80

10.

797

0.79

40.

790

0.78

60.

783

0.77

90.

776

0.77

229

.00.

830

0.82

70.

823

0.82

00.

816

0.81

20.

809

0.80

50.

802

0.79

80.

795

0.79

10.

788

0.78

40.

781

0.77

70.

774

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C0

1000

2000

3000

4000

5000

6000

7000

8000

9000

1000

011

000

1200

013

000

1400

015

000

1600

0

30.0

1.00

00.

997

0.99

40.

991

0.98

80.

985

0.98

10.

978

0.97

50.

972

0.96

90.

966

0.96

20.

959

0.95

60.

953

0.95

031

.01.

000

0.99

70.

994

0.99

10.

988

0.98

50.

982

0.97

80.

975

0.97

20.

969

0.96

60.

963

0.95

90.

956

0.95

30.

950

32.0

1.00

00.

997

0.99

40.

991

0.98

80.

985

0.98

20.

979

0.97

50.

972

0.96

90.

966

0.96

30.

960

0.95

70.

953

0.95

033

.01.

000

0.99

70.

994

0.99

10.

988

0.98

50.

982

0.97

90.

976

0.97

30.

969

0.96

60.

963

0.96

00.

957

0.95

40.

951

34.0

1.00

00.

997

0.99

40.

991

0.98

80.

985

0.98

20.

979

0.97

60.

973

0.97

00.

967

0.96

30.

960

0.95

70.

954

0.95

135

.01.

000

0.99

70.

994

0.99

10.

988

0.98

50.

982

0.97

90.

976

0.97

30.

970

0.96

70.

964

0.96

10.

957

0.95

40.

951

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C17

000

1800

019

000

2000

021

000

2200

023

000

2400

025

000

2600

027

000

2800

029

000

3000

031

000

3200

033

00

30.0

0.94

60.

943

0.94

00.

936

0.93

30.

930

0.92

70.

923

0.92

00.

917

0.91

30.

910

0.90

70.

903

0.90

00.

896

0.89

331

.00.

947

0.94

30.

940

0.93

70.

934

0.93

00.

927

0.92

40.

920

0.91

70.

914

0.91

10.

907

0.90

40.

901

0.89

70.

894

32.0

0.94

70.

944

0.94

10.

937

0.93

40.

931

0.92

80.

924

0.92

10.

918

0.91

40.

911

0.90

80.

905

0.90

10.

898

0.89

533

.00.

947

0.94

40.

941

0.93

80.

935

0.93

10.

928

0.92

50.

922

0.91

80.

915

0.91

20.

908

0.90

50.

902

0.89

90.

895

34.0

0.94

80.

945

0.94

10.

938

0.93

50.

932

0.92

90.

925

0.92

20.

919

0.91

60.

912

0.90

90.

906

0.90

30.

899

0.89

635

.00.

948

0.94

50.

942

0.93

90.

935

0.93

20.

929

0.92

60.

923

0.91

90.

916

0.91

30.

910

0.90

60.

903

0.90

00.

897



Tab

le B

3S

alin

ity

Co

rrec

tio

n F

acto

rs f

or

Dis

solv

ed O

xyg

en i

n W

ater

(b

ased

on

co

nd

uct

ivit

y) (

Tem

p °

C, t

emp

erat

ure

in

deg

rees

Cel

siu

s;

salin

ity

corr

ecti

on

fac

tors

at

30 t

o 3

5°C

are

sh

ow

n a

t th

e en

d o

f th

is t

able

) (C

on

tin

ued

)

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C34

000

3500

036

000

3700

038

000

3900

040

000

4100

042

000

4300

044

000

4500

046

000

4700

048

000

4900

050

000

30.0

0.89

00.

886

0.88

30.

879

0.87

60.

873

0.86

90.

866

0.86

20.

859

0.85

50.

852

0.84

90.

845

0.84

20.

838

0.83

531

.00.

890

0.88

70.

884

0.88

00.

877

0.87

30.

870

0.86

70.

863

0.86

00.

856

0.85

30.

850

0.84

60.

843

0.83

90.

836

32.0

0.89

10.

888

0.88

40.

881

0.87

80.

874

0.87

10.

868

0.86

40.

861

0.85

70.

854

0.85

10.

847

0.84

40.

840

0.83

733

.00.

892

0.88

90.

885

0.88

20.

879

0.87

50.

872

0.86

80.

865

0.86

20.

858

0.85

50.

851

0.84

80.

845

0.84

10.

838

34.0

0.89

30.

889

0.88

60.

883

0.87

90.

876

0.87

30.

869

0.86

60.

863

0.85

90.

856

0.85

20.

849

0.84

60.

842

0.83

935

.00.

893

0.89

00.

887

0.88

30.

880

0.87

70.

874

0.87

00.

867

0.86

30.

860

0.85

70.

853

0.85

00.

847

0.84

38.

400

Tem

p

Co

nd

uct

ivit

y, i

n m

icro

siem

ens

per

cen

tim

eter

at

25°C

˚C51

000

5200

053

000

5400

055

000

5600

057

000

5800

059

000

6000

061

000

6200

063

000

6400

065

000

6600

067

000

30.0

0.83

10.

828

0.82

40.

821

0.81

70.

814

0.81

00.

807

0.80

30.

800

0.79

60.

793

0.78

90.

786

0.78

20.

779

0.77

531

.00.

832

0.82

90.

825

0.82

20.

818

0.81

50.

811

0.80

80.

804

0.80

10.

797

0.79

40.

790

0.78

70.

783

0.78

00.

776

32.0

0.83

30.

830

0.82

60.

823

0.82

00.

816

0.81

30.

809

0.80

60.

802

0.79

90.

795

0.79

20.

788

0.78

50.

781

0.77

833

.00.

834

0.83

10.

828

0.82

40.

821

0.81

70.

814

0.81

00.

807

0.80

30.

800

0.79

70.

793

0.79

00.

786

0.78

30.

779

34.0

0.83

60.

832

0.82

90.

825

0.82

20.

818

0.81

50.

812

0.80

80.

805

0.80

10.

798

0.79

50.

791

0.78

80.

784

0.78

135

.00.

837

0.83

30.

830

0.82

60.

823

0.82

00.

816

0.81

30.

809

0.80

60.

803

0.79

90.

796

0.79

20.

789

0.78

50.

782

Sou

rce:

US

GS

Gro

undw

ater

Sam

plin

g F

ield

Man

ual (

Cha

pter

6).



Table B4 Standard Half-Cell Potentials of Selected Reference Electrodes as a Function of Temperature and Potassium Chloride Reference-Solution Concentration, in Volts (Liquid-junction potential included — multiply volts by 1000 to convert to millivolts; KCl, potassium chloride; Temp °C, temperature in degrees Celsius; M, molar; —, value not provided in

reference)

Silver:silver chloride

Calomel

Temp °C

3M KCI

3.5M KCI

Saturated KCI

3M KCI

3.5M KCI

4M KCI

Saturated KCI

Orion™ 96–78 Combination

Electrode

10 0.220 0.215 0.214 0.260 0.256 — 0.254 0.25615 0.216 0.212 0.209 — — — 0.251 0.25320 0.213 0.208 0.204 0.257 0.252 — 0.248 0.24925 0.209 0.205 0.199 0.255 0.250 0.246 0.244 0.24630 0.205 0.201 0.194 0.253 0.248 0.244 0.241 0.24235 0.202 0.197 0.189 — — — 0.238 0.23840 0.198 0.193 0.184 0.249 0.244 0.239 0.234 0.234


Table B5 Eh of ZoBell’s Solution as a Function of Temperature (°C, degrees Celsius; mV,

millivolts)

Temperature °C Eh (mV)

10 46712 46214 45716 45318 44820 44322 43824 43325 43026 42828 42330 41832 41634 40736 40238 39740 393

Source: USGS Groundwater SamplingField Manual (Chapter 6).



Table B6 Troubleshooting Guide for Eh Measurement (±, plus or minus; mV,

millivolts;

emf

, electromotive force)

Symptom Possible Corrective Action

Eh of ZoBell’s solution exceeds theoretical by ± 5 mV •

Check meter operation:

Use shorting lead to establish meter reading at zero mV.

Excessive drift • Check/replace batteries.Erratic performance • Check against backup meter.Poor response when using paired electrodes

Check electrode operation:

• Check that electrode reference solution level is to the fill hole.

• Plug questionable reference electrode into reference electrode jack and another reference electrode in good working order of the same type into the indicator electrode jack of the meter; immerse electrodes in a potassium chloride solution, record mV, rinse off and immerse electrodes in ZoBell’s solution. The two mV readings should be 0 ± 5 mV. If using different electrodes (Ag:AgCl and Hg:HgCl

2

), reading should be 44 ± 5 mV for a good reference electrode.

• Polish platinum tip with mild abrasive (crocus cloth, hard eraser, or a 400–600-grit wet/dry Carborundum™ paper), rinse thoroughly with deionized water. Use a Kimwipe™ if these abrasives are not available.

• Drain and refill reference electrolyte chamber.• Disconnect reference electrode. Drain and refill

electrolyte chamber with correct filling solution. Wipe off connectors on electrode and meter. Use backup electrode to check the

emf

.• Read

emf

with fresh aliquot of ZoBell’s solution; prepare fresh ZoBell’s solution if possible.

• Recondition electrode by cleaning with aqua regia and renewing filling solution —

this is a last resort

.




Figure B1

Field-measurement procedures using downhole and flowthrough-chamber sys-tems for groundwater. Source: USGS Groundwater Sampling Field Manual(Chapter 6).

Test and calibrate field measurements. Selectdownhole or flowthrough-chamber system.

DOWNHOLE FLOWTHROUGH CHAMBER

Lower sensors and pumpto selected depth.

Install pump in monitoring well or plumbingfor use of existing pump in a supply well.Install sensors in flowthrough chamber.

Turn on pump and adjust flow todesired purge rate, record rate andtime purging began. Allow sensors

to equilibrate at purge rate.

Adjust flow to desired purge rate andrecord rate and time purging began.

• Divert initial water to waste.

• Correct chamber in-line from pump.

• Adjust flow to chamber andeliminate backpressure: allowsensors to equilibrate

Record and monitor sequential sets of field measurementreadings during withdrawal of trial well volume.

• After two or more well volumes are purged and beforefinal five or more readings are made, adjust flow rate tobe used for sampling flow; flow must be sufficient fordissolved oxygen measurements.

• For pH: divert flow from chamber and recordmeasurement when water is quiescent. Redirect flow tochamber for next set of readings.

Are stabilization criteria being met? YESNO

Extend purge time. Documentdifficulty in field notes.

Record at least 5 measurements atintervals of 3 to 5 minutes or more.

Report the median of the last 5 or morereadings and the time of measurement.



Figure B2

Subsample field-measurement procedures for conductivity, pH, and alkalinity ofgroundwater. Source: USGS Groundwater Sampling Field Manual (Chapter 6).

Test and calibrate field instruments.

Purge well (see text for exceptions and instructions).

Field rinse precleaned sampler. Use clean/dirty hands technique.Lower sampler smoothly, without splashing, to desired depth of screened or open interval.

(If using bailer, double check-valve type is recommended.)Raise sampler smoothly at a constant rate, keeping lines clean and off the ground.

Place sampler in holding stand.

Withdraw subsamples from sampler.• If using bailer, insert clean bottom-emptying device with groved hands; device should fit

snugly over collection bottles and (or) measurement vessels.• If a filtered sample is needed, filter in-line from sampler to bottle/vessel.• Drain subsample without turbulence into collection bottles or measurement vessel.• Rinse collection bottle(s) three times with sample (filtrate for filtered samples), fill to brim,

cap tightly, and maintain at temperature of water source until measurement.• Rinse sensors, stir bar, and measurement vessel three times with sample.• For alkalinity, rinse with deionized water.

Insert sensor(s) in measurement vessel.• Wait for sensors to equilibrate to sample temperature.• Don’t let sensors touch bottom or sides of vessel.

Swirl or stir gently to mix sample.• Minimize streaming potential or vortex; keep sensor out of vortex.• For pH, do not stir samples with conductivity less than 100 S/cm.• When using magnetic stirrer, use smallest stir bar.

Record field measurement and method used on field form.• Record median value of stabilized readings.• If readings do not stabilize, extend number of measurements and record median of at

least 5 or more sequential readings.• If there is a constant trend toward lower or higher values, record the first value, the range

of values, and the time period observed.

Repeat process from steps (4) through (7) on two or more subsamples from the samesample volume to document precision.

Rinse sensors and equipment thoroughly with deionized water.Discard sample to waste, in accordance with applicable regulations.

Step (1)

(2)

(3)

(4)

(5)

(6)

(7)

(8)

L1282/Appendix B/frame Page 402 Tuesday, June 19, 2001 1:15 PM


Figure B3

Factors used to correct atmospheric pressures adjusted to sea level. Source:USGS Groundwater Sampling Field Manual (Chapter 6).

-1000

-500

0

500

1000

1500

2000

2500

3000

3500

4000

4500

5000

5500

6000

6500

7000

7500

8000

VALUE TO SUBTRACT FROM ATMOSPHERIC PRESSURE, IN MILLIMETERS OFMERCURY

EL

EVA

TIO

N, N

GV

D O

F 1

929,

IN F

EE

T

0 6040 80 100 100 120 160140 180 20020



405

APPENDIX

C

Common and Scientific Names ofVarious Plants

Common Name Scientific Name

Alfalfa

Medicago sativa

Algae stonewort

Nitella

Alyssum

Alyssum wulfenianum

Bean

Phaseolus coccineus

L.Bean, bush

Phaseolus vulgaris

cv. Tender GreenBermuda grass

Cynodon dactylon

Birch, river

Betula nigra

Black locust

Robinia pseudoacacia

Black willow tree

Salix nigra

Bladder campion

Silene vulgaris

Bluestem, big (prairie grass)

Andropogon gerardi

Vit.Bluestem, little

Andropogon scoparius

Bluestem, little


Boxwood

Buxaceae

Buffalo grass

Buchloe dactyloides

Canada wild rye (prairie grass)

Elymus canadensis

Canola

Brassica napus

Cattail

Typha latifolia

Cherry bark oak

Quercus falcata

Clover

Genus trifolium

Colonial bentgrass

Agrostis tenuis

cv. GoginanColonial bentgrass

Agrostis tenuis

cv. ParysCottonwood, eastern

Populus deltoides

Cottonwood (poplar)

Populus

Crab apple

Malus fusca

Raf. Schneid

L1282/Appendix C/frame Page 405 Tuesday, June 19, 2001 1:16 PM


Crested wheatgrass (Hycrest)


(Fisher ex Link)Schultes

Cypress, bald

Taxodium distichum

Duckweed

Lemna minor

Eastern Cottonwood

Populus deltoides

European milfoil/yarrow

Achillea millefolium

Felt leaf willow

Salix alaxensis

Fescue, hard

Festuca ovina

var.

duriuscula

Fescue, red

Festuca rubra

cv. MerlinFescue, tall

Festuca arundinacea

Schreb.Four-wing saltbrush

Aicanescens

Grama, side oats (prairie grass)

Bouteloua curtipendula

Grama, blue

Bouteloua gracilis

Grass, cool season (colonial bentgrass)

Agrotis tenuis

Grass, warm season (Japanese lawngrass)

Zoysia japonica

Horseradish (roots)

Armoracia rusticana

Hybrid poplar

Populus deltoides x nigra

DN-34,Imperial California;

Populus charkowiieensis

x incrassata

;

Populus tricocarpa

x

deltoides

Hycrest, crested wheatgrass


Indiangrass (prairie grass)

Sorghastrum nutans

Indian mustard

Brassica juncea

Indian ricegrass

Oryza sativa

subsp.

indica

Japanese lawngrass

Zoysia japonica

Jimson weed (thornapple)

Datura innoxia

Kenaf

Hibiscus cannabinus

L. cv. IndianKoa haole

Leucaena leuccephala

Kudzu

Pueraria lobata

Lambsquarter

Chenopodium

Legume

Lespedeza cuneata

(Dumont)Little bluestem (prairie grass)


Loblolly pine

Pinus taeda

L.Mesquite

Prosopis

Millet, Proso

Panicum miliaceum

L.Mulberry, red

Morus rubra

L.Mustard, Indian

Brassica juncea

Mustard weed

Arabidopsis thaliana

Oak, cherry bark

Quercus falcata

Oak, live

Quercus virginiana

Osage, orange

Maclura pomifera

(Raf.) SchneidParrot feather

Myriophyllum aquaticum thlaspi rotundifolium

Pennycress

Thlaspi rotundifolium

Pennycress, alpine

Thlaspi caerulescens

Pennyworth

Hydrocotyle umbellata

L1282/Appendix C/frame Page 406 Tuesday, June 19, 2001 1:17 PM

COMMON AND SCIENTIFIC NAMES OF VARIOUS PLANTS 407

Poplar, cottonwood

Populus

Poplar, hybrid

Populus deltoides x

nigra

DN-34, Imperial California;

Populus charkowiieensis x incrassata;Populus tricocarpa x deltoides;Populus charkowiieensis x incrassata

Poplar, yellow

Liriodendron tulipifera

Red fescue

Festuca rubra

cv. MerlinReeds

Phragmites

Rice

Oryza sativa

L.Sacaton, alkali

Sporobolus wrightii

Sea pink; wild thrift

Armeria maritima

Salt marsh plant

Spartina alterniflora

Sand dropseed

Sporobolu cryptandrus

Soybean

Glycine max

L. Merr, cv. DavisSpearmint

Mentha spicata

Sugarcane

Saccharum officinarum

Sudangrass

Sorghum vulgare

L.Sunflower

Helianthus annuus

Switchgrass (prairie grass)

Panicum virgatum

Tall fescue

Festuca arundinacea

Schreb.Thornapple (jimson weed)

Datura innoxia

Thrift (wild); sea pink

Armeria maritima

Tobacco

Nicotiana tabacum

Water hyacinth


Water milfoil

Myriophyllum spicatum

Water velvet

Azolla pinnata

Wheat grass, slender

Agropyron trachycaulum

Wheat grass, western (prairie grass)

Agropyron smithii

Willow tree, black

Salix nigra

Willow tree, felt leaf

Salix alaxensis

L1282/Appendix C/frame Page 407 Monday, June 18, 2001 9:14 AM

L1282/Appendix C/frame Page 408 Monday, June 18, 2001 9:14 AM

409

Index

A

Absorption, 39Acceptable biodegradation, 26Adenosine triphosphate, 90–91Adsorption, 39Advection dispersion equation, 84–85Aerobic oxidation

cometabolism, 202–204description of, 200–201indications for, 200MTBE degradation, 204–205rates of, 201

Aging, 33Alkane hydroxylase, 30Alkane monooxygenase, 30Alkylbenzene sulfonate, 34Alkylhalides, aerobic oxidation of, 201American Society of Testing and Materials,

monitored natural attenuation support by, 64

Ammonia monooxygenase, 30Ammonification, 197–198Anaerobic organisms, 5Analogue enrichment, 25Anoxic environments, oxidants in, 57Aquifer matrix, 47Aquifer solids, 51Arsenic

biodegradation of, 107

in situ

precipitation, 194–195Atmospheric pressures, 383, 403

B

Benzene, toluene, ethylbenzene, and xylenebiodegradation of, 95–97description of, 6

fermentation, 151natural attenuation of, 97sources of, 95

Bioaugmentation, 136–138Biodegradation

acceptable, 26arsenic, 107benzene, toluene, ethylbenzene, and xylene,

95–97chemical oxidation pretreatment effects, 223chemical structure and, 36chemical substrate concentration effects, 27chlorate, 108chlorinated aliphatics, 99chlorinated aromatics, 100–103chlorobenzoates, 100–101chlorophenols, 100–101chromium, 106–107cometabolism effects, 32conditions necessary for, 26definition of, 26electron-acceptor-limited model of, 94–95environmental impacts of, 32–34first-order decay model, 92–94lindane, 106mercury, 107metals, 105–106natural attenuation effects, 90–109nitrate, 107–108nitrate esters, 104nitroaromatics, 103–104organic contaminants, 95oxyanions, 107–108oxygenated hydrocarbons, 98–99pathways of, 35pentachlorophenol, 101–102perchlorate, 108phyto-cover effects, 326–327polychlorinated biphenyls, 101polycyclic aromatic hydrocarbons, 98, 308

L1282_frame_IDX Page 409 Monday, June 18, 2001 9:13 AM


primary, 26rates of, 35–38selenium, 108structural effects, 34–35treatment wetlands for, 306–309ultimate, 26

Bioemulsifiers, 165Biogeochemical

definition of, 89reactions, 89–90

Bioremediation, 6Biostimulation, 136–138Biosurfactants, 165–166Biphenyl dioxygenase, 30Boiling point, 18Butane-utilizing bacteria, 202–203

C

Carbon tetrachloride, reductive dechlorination of, 159

Chelates, 247Chemical oxidation

advantages of, 206–207applications of

description of, 218–2191,4-dioxane, 222–223

biodegradation effects, 223chemistry of, 208–218concerns regarding, 207description of, 205–206mechanism of action, 206oxidants used

description of, 208–209hydrogen peroxide, 208–209, 211–213potassium permanganate, 209, 213–216

predictions, 219–220Chlorate biodegradation, 108Chlorinated aliphatics

aerobic oxidation of, 201biodegradation of, 99, 138–139mixture of, 156natural attenuation of, 75, 99

Chlorinated alkanesaerobic oxidation of, 202treatment wetland removal of, 306

Chlorinated aromaticsbiodegradation of, 100–103, 138–139natural attenuation of, 75, 100–103

Chlorinated ethenes, reductive dechlorination using, 145, 170–177

Chlorobenzene biodegradation, 100Chlorobenzoate biodegradation, 100–101

Chloroethene reductive dehalogenases, 147Chloroform, reductive dechlorination of, 159Chlorophenol biodegradation, 100–101Chromium

biodegradation of, 106–107

in situ

precipitation, 192–194Colloidal organic matter, 49Colloidal particles, 196–197Cometabolism

aerobic, 202–204cautions associated, 28–29chemicals subject to, 28chlorinated solvents, 139definition of, 27, 140environmental consequences of, 28polycyclic aromatic hydrocarbons, 25reactions

co-oxidation, 140mechanisms of, 29–32reactive dechlorination,

See

Reductive dechlorination

types of, 28Comprehensive Environmental Response,

Compensation, and Liability Act, 314

Constructed treatment wetlandsalgae in, 279–280bacteria in, 278–279benefits of, 272components of, 270–271contaminant removal mechanisms

hydraulic retention time, 297–299partitioning and storage, 295–296volatilization, 294–295

description of, 270dissolved oxygen concentration, 294in Europe, 272–273fish in, 310free water surface, 276fungi in, 278–279groundwater remediation use of, 299–310horizontal flow systems

characteristics of, 276–277emergent macrophyte-based systems with,

285hydrology of, 309–310hydroperiods, 292inorganics removal using, 309issues regarding, 273–274metals removal, 300–305morphology of, 309–310popularity of, 271regulation of, 272, 275–276research of, 274


INDEX 411

soilsbiological influences, 292cation exchange capacity, 289–290characteristics of, 287–289clays, 287hydric, 287microbial processes, 292–293mineral, 287organic, 287–288oxidation reactions, 290–291peats, 287–288pH, 292reduction reactions, 290–291

subsurface flow, 276–277toxic organics removal using, 306–309in United States, 272–273vegetation in

description of, 279, 281–282duckweeds, 283–284emergent macrophyte-based systems

description of, 284with horizontal subsurface flow, 285with vertical subsurface flow, 285

free-floating macrophyte-based systems, 282–284

multistage macrophyte-based systems, 287

submerged macrophyte-based systems, 285–287

water hyacinths, 282–283vertical flow systems


285wildlife considerations, 274–275, 310

Contaminantsavailability of, 14biodegradation of

acceptable, 26chemical structure and, 36chemical substrate concentration effects,

27cometabolism effects, 32conditions necessary for, 26definition of, 26environmental impacts of, 32–34pathways of, 35primary, 26rates of, 35–38structural effects, 34–35ultimate, 26

biological characteristics of, 26–38boiling point of, 18characteristics of, 16–17cometabolism of, 27–32

dense nonaqueous-phase liquiddefinition of, 73identification methods, 88–89permanganate oxidation of, 220–221transport of, 88

Henry's law constant, 19–20hydrolysis of, 22–24immobilization of, 196inorganic

biodegradation of, 105treatment wetland removal of, 309

light nonaqueous-phase liquid, 73light nonaqueous-phase liquids, 73natural attenuation removal of, 72–77nonaqueous-phase liquid

definition of, 73–74transfer of, 88

nonaqueous-phase liquids, 73–74octanol/water partition coefficients, 20photolytic reactions, 24–25physical properties of, 365–382sequestration of, 33–34in soil-plant system, 242–243solubility, 20–22sorption coefficient of

competitive, 49description of, 38–39factors that affect, 39, 48–51kinetic considerations, 50pH effects, 49soil, 43–48temperature effects, 48

source areas of, 8sources zones for, 72–73subsurface movement methods

advection, 81–82dispersion, 83–87

treatment wetland removal ofhydraulic retention time, 297–299partitioning and storage, 295–296volatilization, 294–295

vapor pressure of, 18Cooxidation, 28Cosolvents, 49

D

Dechlorination,

See

Reductive dechlorinationDehalobacgter restrictus, 145, 147Dehalococcoides ethenogenes, 146Dehalospirillum multivorans, 145, 147Dehydrohalogenation, 23Denitrification,

in situ

, 197–198



Dense nonaqueous-phase liquid contaminantsdefinition of, 73identification methods, 88–89permanganate oxidation of, 220–221transport of, 88

Desulfitobacterium sp., 145Desulfuromonas chloroethenica, 1463,5-Dichlorobenzoate, reductive dechlorination

of, 156Dilution, natural attenuation effects, 79–811,4-Dioxane, ozone oxidation of, 222–223Dispersion

advection dispersion equation, 84–85definition of, 83mechanical, 84molecular, 83–84

Dissolved organic matter, 49Dissolved oxygen

measurements ofdescription of, 113electrodes, 113–115in field, 116field calibrations, 115–116

salinity correction factors, 392–398Distribution coefficient

definition of, 45–46estimating of, 87

Duckweed-based wetland systems, 283–284

E

Earthanaerobic organisms, 5history of, 5–6

Electron-acceptor-limited model, 94–95Enhanced reductive dechlorination

anaerobic oxidation, 158biostimulation vs. bioaugmentation, 136–138cometabolic, 142–144electron acceptors, 159–160electron donor selection for, 147–158field studies of, 160–164groundwater solute transport of chlorinated

ethenes, 170–177halorespiring microorganisms, 144–147high constituent concentration areas, 169–170low constituent concentration areas, 169mechanisms of, 138–142microorganisms for

electron donors for, 147–158halorespiring, 144–147hydrogen production, 149–151nutrients, 158

mixture of compounds effect, 155–158

oxidation-reduction potential effects, 142

in situ

reactive zones forbiofilm, 168–169development considerations, 160–163fermentation, 167–168natural surfactant effect, 165–167performance data, 177–183reagent delivery, 164–165

studies of, 135–136temperature effects, 158

Enterobacter agglomerans, 146Environment, biodegradation effects on, 32–34Environmental Protection Agency, monitored

natural attenuation support by, 64Estuaries, natural organic material in, 42Evapotranspiration

description of, 321effective, 340–343potential, 337–340

F

Fermentationbenzene, toluene, ethylbenzene, and xylene,

151definition of, 150hydrogen produced during, 149–151, 153–154primary, 151reductive dechlorination, 167–168secondary, 151

First-order decay model, 92–94Flocculation, 196Fortuitous metabolism, 27

G

Groundwaterfield-measurement procedures, 401–402mobile water phase of, 121phytoremediation of, 259–260, 263rock/soil phase of, 121treatment wetland remediation of, 299–310velocity of, 82

Groundwater capture zone, 267Growth rate, 37

H

Half-velocity coefficient, 37Halorespiration, 144–145


INDEX 413

Halorespiring microorganismshydrogen consumption, 154reductive dechlorination using, 144–147types of, 144

Hazarddefinition of, 2predicting of, 4subjective probability of, 2transport of, 4–5

Hazardous waste, 3Henry's law constant, 19–20Hexachlorbutadiene, treatment wetland removal

of, 308Hexachlorobenzene, treatment wetland removal

of, 308High molecular weight natural organic matter, 58Humic soil organic matter, 42Hydrocarbons, natural attenuation of, 75Hydrogen

fermentation-induced production of, 149–151, 153–154

microorganism production ofcompetition, 152–153description of, 149–151

Hydrogenolysis, 55Hydrogen peroxide,

in situ

chemical oxidation uses of, 208–209, 211–213

Hydrolysisdefinition of, 22equation, 22plume effects, 23–24rates of, 23reaction products produced, 22–23

Hyperaccumulators, 246–247

I

Inorganic contaminantsbiodegradation of, 105treatment wetland removal of, 309

In situ

chemical oxidationadvantages of, 206–207applications of

description of, 218–2191,4-dioxane, 222–223

biodegradation effects, 223chemistry of, 208–218concerns regarding, 207description of, 205–206mechanism of action, 206oxidants used

description of, 208–209hydrogen peroxide, 208–209, 211–213

potassium permanganate, 209, 213–216predictions, 219–220

In situ

denitrification, 197–198

In situ

precipitation of metalsaquifer parameters, 195–196arsenic, 194–195chromium, 192–194contaminant removal, 196–197description of, 183, 188principles of, 187–195

In situ

reactive zonesadvantages of, 133–134definition of, 132description of, 8design of, 132effectiveness of, 132–133metals, 134–135nano-scale particle injections

irondescription of, 223–228history of use, 225organic contaminants treatable using,

231–232particle production, 228–230particle size, 226, 228reductive process, 224–225

permeable sediment applications, 231reductive dechlorination

biofilm, 168–169development considerations, 160–163fermentation, 167–168natural surfactant effect, 165–167performance data, 177–183reagent delivery, 164–165

schematic representation of, 132–133types of, 134–135

Instantaneous reaction model, 94–95Intrinsic remediation,

See

Monitored natural attenuation

Ionic strength, 49Ionizability, 50–51Iron, nano-scale injections in

in situ

reactive zonesdescription of, 223–228history of use, 225organic contaminants treatable using,


IRZ,

See

In situ

reactive zonesISCO,

See

In situ

chemical oxidationIsotherms

definition of, 43linear sorption, 43–44nonlinear, 44



L

Lakesnatural organic material in, 42recharging, 79

Landfillhistory of, 314–315leachates, 273, 317microbial activity in, 317moisture flow and content, 333waste stabilization phases in, 317–320

Landfill coverbarrier-type, 316capillary barrier, 321conventional

description of, 316–317phyto-cover vs., 326

design features of, 316evapotranspiration, 321function of, 315–316phyto-cover,

See

Phyto-coverrequirements for, 315

Light nonaqueous-phase liquid contaminants, 73Lindane biodegradation, 106Liquid-filled macropores, 14–15Low molecular weight natural organic matter, 58

M

Marine waters, contaminant levels in, 33Mass removal efficiency, 7Maximum growth rate, 37Mechanical dispersion, 84Mercury, biodegradation of, 107Metals,

See also

specific metalbiodegradation of, 105–106field filtration of, 121–124filtered vs. unfiltered samples, 120–124microbial transformation of, 190natural attenuation of, 75precipitates, 187

in situ

precipitationaquifer parameters, 195–196arsenic, 194–195chromium, 192–194contaminant removal, 196–197description of, 183, 188principles of, 187–195

in situ

reactive zones, 134–135soil concentrations of, 186, 241–242treatment wetlands for, 299–310

Methane monooxygenasesdescription of, 30types of, 203

Methanotrophs, 202–203Methylene chloride, reductive dechlorination of,

159Micelles, 289–290Microbial ecology, 6Microorganisms

electron donors, 92hydrogen production by, 149–151inorganic contaminants transformed by, 105reproductive mechanisms of, 90–91in rhizodegradation, 252

Molecular dispersion, 83–84Monitored natural attenuation

capacity, 77–78definition of, 64description of, 8documenting of, 66evaluative approaches, 65–69monitoring and sampling of

case study, 124–125considerations, 109–110description of, 109–113dissolved oxygen, 113–116field portable meters, 111filtered vs. unfiltered metal samples,

120–124low-flow sampling, 124on-site, 111oxidation-reduction potential, 117–119pH, 119–120

in situ

, 112–113techniques for, 110–113

organizations that support, 64–65patterns of

contaminant sources, 72–77description of, 71questions for assessing, 71–72

processes that affectadvection, 81–82biodegradation, 90–109dilution, 79–81dispersion, 83–87stabilization, 88–89volatilization, 89

protocols, 70–71terminology associated with, 64–65

MTBE, aerobic oxidation of, 204–205

N

Nano-scale particle injections, in

in situ

reactive zones

irondescription of, 223–228


INDEX 415

history of use, 225organic contaminants treatable using,


permeable sediment applications, 231Naphthalene dioxygenase, 30Natural attenuation

capacity, 77–78definition of, 64description of, 8documenting of, 66evaluative approaches, 65–69monitoring and sampling of

case study, 124–125considerations, 109–110description of, 109–113dissolved oxygen, 113–116field portable meters, 111filtered vs. unfiltered metal samples,

120–124low-flow sampling, 124on-site, 111oxidation-reduction potential, 117–119pH, 119–120

in situ

, 112–113techniques for, 110–113

organizations that support, 64–65patterns of

contaminant sources, 72–77description of, 71questions for assessing, 71–72

processes that affectadvection, 81–82biodegradation, 90–109dilution, 79–81dispersion, 83–87stabilization, 88–89volatilization, 89

protocols, 70–71terminology associated with, 64–65

Natural attenuation capacity, 77–78Natural degradation,

See


Natural organic matterhigh molecular weight, 58low molecular weight, 58

Natural surfactant effect, 165–167Nitrate biodegradation, 107–108Nitrate esters, 104Nitrification, 198Nitroaromatics

biodegradation of, 103–104natural attenuation of, 75, 103–104

Nonaqueous-phase liquid contaminantsdefinition of, 73–74transfer of, 88

O

Octanol/water partition coefficients, 20Organic contaminants, 95Organic matter

colloidal, 49dissolved, 49soil

composition of, 42critical level of, 46definition of, 41–42humic, 42nonhumic, 42sediments vs., 42–43

sorption coefficient effects, 49–50Oxidants

in chemical oxidation systemsdescription of, 208–209hydrogen peroxide, 208–209, 211–213potassium permanganate, 209, 213–216

in natural systems, 56Oxidation-reduction,

See

REDOXOxidation-reduction potential

definition of, 117measurement of, 117–119reductive dechlorination effects, 142wetland soils, 290–292

Oxyanionsbiodegradation of, 107–108natural attenuation of, 76, 107–108

Oxygendissolved

description of, 113electrodes, 113–115in field, 116field calibrations, 115–116salinity correction factors, 392–398

solubility in water, 384–391Oxygenase, 30Oxygenated hydrocarbons

biodegradation of, 98–99natural attenuation of, 75, 98–99sources of, 98

Oxyhydroxides, 123Ozone

chemical structure of, 2171,4-dioxane oxidation, 222–223

in situ

chemical oxidation uses of, 208, 216–218



P

Particle diffusion, description of, 15Passive remediation,

See


pe, 51–52Pentachlorophenol biodegradation, 101–102Perchlorate

biodegradation of, 108chlorite, 199reduction of, 199–200

in situ

reactive zone for, 199–200Pesticide biodegradation, 104pH

definition of, 51measurement of, 119–120sorption coefficient effects, 49wetland soil, 292

Photolytic reactionsdefinition of, 24direct, 24–25function of, 24indirect, 25

Photoreactions,

See

Photolytic reactionsPhreatophytes, 259–260Phytoaccumulation, 245–248, 259, 299Phyto-cover system

agronomic chemistry sampling of, 358–359benefits of

aesthetic, 327–328ecological, 327–328economics, 329gas permeability, 327maintenance, 329public safety, 329

in situ

biodegradation, 326–327case study example of, 344–347characteristics of, 321–323components of, 329–330conventional covers vs., 326description of, 321, 332–333designing of, 333–334examples of, 323–326illustration of, 323irrigation/irrigation system

considerations for, 329–330guidelines for, 352–356requirements, 362

maintenanceactive, 360–361passive, 361preventive, 359

models for designing, 333–335nonsoil amendments to, 331operation and maintenance schedule, 359–362

performance ofhydrologic water balance, 332–335potential evapotranspiration, 337–343precipitation, 335runoff, 335–337water balance model, 343

plants, 322, 331–332poplars, 321–322repairs, 359safety considerations, 359sample application of, 344–347schematic of, 334site inspections, 348–349soil moisture monitoring, 349–352trees

evaluations of, 356–358leaves of, 357–358root systems of, 349selection of, 321–322, 331–332soil moisture monitoring, 349–350stem evaluations, 356

understory planting, 331vegetative, 330–331water balance

model, 343summary overview of, 347–348

Phytodegradation, 248–250, 258Phytoextraction,

See

PhytoaccumulationPhytoremediation

advantages of, 240applications of, 258–259decision tree for, 262definition of, 240description of, 244disadvantages of, 240groundwater contaminants, 259–260, 263mechanism of action, 240phytoaccumulation, 245–248, 259, 299phytodegradation, 248–250, 258phytostabilization, 250–251, 258phytovolatilization, 251–252, 258rhizodegradation, 252–256, 258rhizofiltration, 256–259, 258, 299sediments, 260–261, 263soil-plant system chemicals

metals, 241–242organic compounds, 242–243

system designagronomic inputs, 266contaminant levels, 265groundwater capture zone, 267irrigation, 266maintenance, 266overview of, 261–265plant species, 265–266


INDEX 417

transpiration rate, 267treatability, 266

Phytostabilization, 250–251, 258Phytotransformation,

See

PhytodegradationPhytovolatilization, 251–252, 258Plants

chemical uptake inmetals, 241–242organic compounds, 242–243

enzyme systems of, 248–249phyto-cover system, 322, 331–332rhizodegradation uses of, 255rhizofiltration uses of, 257scientific names of, 405–407in treatment wetlands






water hyacinths, 282–283Pollutants, 5Pollution evaluations, 2Polychlorinated biphenyls

biodegradation of, 101plant uptake of, 243sources of, 101

Polycyclic aromatic hydrocarbonsbiodegradation of, 98, 308cometabolism of, 25half-life of, 295hydrolysis of, 25natural attenuation of, 98sources of, 98

Poplars, phyto-cover use of, 321–322Potassium permanganate,

in situ

chemical oxidation uses of, 209, 213–216

Precipitate, 187Primary biodegradation, 26Proton, 51

R

Reactive zones,

See

In situ

reactive zonesRecharge

definition of, 79

dilution estimations, 80natural attenuation effects, 79–81

REDOXcontaminant transformation, 58measurement of, 111, 119poise, 52–53reactions

description of, 53oxidants, 56–57oxidations, 53–54reductants, 57–58reductions, 54–56

wetland soils, 291zones, 51

Reductants, abiotic environmental, 57Reductive dechlorination

anaerobic oxidation, 158cometabolic, 142–144description of, 55electron acceptors, 159–160electron donor selection for, 147–158field studies of, 160–164groundwater solute transport of chlorinated

ethenes, 170–177halorespiring microorganisms, 144–147high constituent concentration areas, 169–170low constituent concentration areas, 169microorganisms for

electron donors for, 147–158halorespiring, 144–147hydrogen production, 149–151nutrients, 158

mixture of compounds effect, 155–158oxidation-reduction potential effects, 142

in situ

reactive zones forbiofilm, 168–169development considerations, 160–163fermentation, 167–168natural surfactant effect, 165–167performance data, 177–183reagent delivery, 164–165

temperature effects, 158tetrachloroethane, 141–144

Remediationevolution of, 7–11extractive techniques, 7–8goal of, 14phytoremediation,

See

PhytoremediationResource Conservation and Recovery Act, 314Retardation, equations, 85–86Rhizodegradation, 252–256, 258Rhizofiltration, 256–259, 258, 299Rhizosphere, 254Risk

assessment of, 2–3



description of, 2

S

Sedimentsphytoremediation of, 260–261, 263soil organic matter vs., 42–43treatment wetland, 293

Selenium biodegradation, 108Sequestration, 33Smectites, 40Sodium permanganate, 214Soil

characteristics of, 40–41distribution coefficient of, 45metals concentration in, 186, 241–242sorption coefficients, 43–48in treatment wetlands

biological influences, 292cation exchange capacity, 289–290characteristics of, 287–289clays, 287hydric, 287microbial processes, 292–293mineral, 287organic, 287–288oxidation reactions, 290–291peats, 287–288pH, 292reduction reactions, 290–291

Soil organic mattercomposition of, 42critical level of, 46definition of, 41–42humic, 42nonhumic, 42sediments vs., 42–43

Soil watercomposition of, 42functions of, 42

Solid waste, 318Solubility

definition of, 20–21single compound in an immiscible liquid, 21single compound is an immiscible liquid, 22

SOM,

See

Soil organic matterSorption coefficients

competitive, 49description of, 38–39factors that affect, 39, 48–51kinetic considerations, 50pH effects, 49soil, 43–48

temperature effects, 48wetland soils, 289

Source areas, 8Source zones

definition of, 72types of, 72, 76

Streams, recharging, 79

T

Temperaturereductive dechlorination effects, 158sorption coefficient effects, 48

Tetrachloroethanehydrogenolysis of, 141phytodegradation of, 250reductive dechlorination of, 135, 141–144,

152–153treatment wetland removal of, 306

1,1,2,2-Tetrachloroethane, treatment wetland removal of, 306

Toluene dioxygenase, 30–31Toluene monooxygenase, 30Toxicity characteristic leaching procedure, 241Transport processes, description of, 14–15Treatment wetlands,

See

Wetlands, constructed treatment

Trees, in phyto-cover systemevaluations of, 356–358leaves of, 357–358replacement of, 362root systems of, 349selection of, 321–322, 331–332soil moisture monitoring, 349–350stem evaluations, 356

1,1,1-Trichloroethanehydrolysis effects, 23–24reductive dechlorination of, 159

Trichloroethylene, enhanced reductive dechlorination of, 135

Trinitrotoluene biodegradation, 104

U

Ultimate biodegradation, 26

V

Vapor pressure, 18Vicinal dehalogenation, 55


INDEX 419

Volatile organic compounds, Henry's law constant, 19

Volatilizationdefinition of, 89natural attenuation effects, 89treatment wetland use of, 294–295

W

Water hyacinth-based wetland systems, 282–283Wetlands, constructed treatment

algae in, 279–280bacteria in, 278–279benefits of, 272components of, 270–271contaminant removal mechanisms

hydraulic retention time, 297–299partitioning and storage, 295–296volatilization, 294–295

description of, 270dissolved oxygen concentration, 294in Europe, 272–273fish in, 310free water surface, 276fungi in, 278–279groundwater remediation use of, 299–310horizontal flow systems


285hydrology of, 309–310hydroperiods, 292inorganics removal using, 309issues regarding, 273–274metals removal, 300–305morphology of, 309–310popularity of, 271regulation of, 272, 275–276research of, 274soils

biological influences, 292cation exchange capacity, 289–290characteristics of, 287–289clays, 287hydric, 287microbial processes, 292–293mineral, 287organic, 287–288oxidation reactions, 290–291peats, 287–288pH, 292reduction reactions, 290–291

subsurface flow, 276–277toxic organics removal using, 306–309in United States, 272–273vegetation in






water hyacinths, 282–283vertical flow systems


285wildlife considerations, 274–275, 310

Z

ZoBell's solution, 399–400Zwitterions, 217



Documents

Natural_and_Enhanced.pdf