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PHYTOREMEDIATION OF ARSENIC CONTAMINATED SOILS BY PTERIS VITTATA: ARSENIC REMOVAL AND BIOMASS DISPOSAL
By
EVANDRO BARBOSA DA SILVA
A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY
UNIVERSITY OF FLORIDA
2018
© 2018 Evandro Barbosa Da Silva
To Pâmela and Maria
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ACKNOWLEDGMENTS
I would like to thank God for my life and for guiding my path to this achievement.
I would like to express my sincere gratitude and appreciation to Dr. Lena Q. Ma,
my advisor, for letting me be a part of her group and for her support, trust, patience and
encouragement. I also would like to thank my co-chair, Dr. Ann C. Wilkie, for the
scientific inspiration and guidance she provided in my research and professional
development, encouraging me to keep pursuing a sustainable world and introducing me
to anaerobic digestion technology. I would also like to thank my committee members Dr.
Willie Harris and Dr. Bala Rathinasabapathi, for their valuable suggestions and the time
they spent on ensuring the smooth progress of my research. I would like to thank Dr.
Harris for assisting me with the XRD analysis.
Also, many thanks to my labmates: Jay Lessl, Rujira Tisarum, Ky Gress,
Andressa Freitas, Letuzia Oliveira and Peng Gao for their supports in my research and
personal life. Finally, yet importantly I would like to recognize my family specially my
wife and parents for providing me emotional support and giving me all the opportunities
and education necessary to pursuit my dreams.
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TABLE OF CONTENTS page
ACKNOWLEDGMENTS .................................................................................................. 4
LIST OF TABLES ............................................................................................................ 7
LIST OF FIGURES .......................................................................................................... 8
ABSTRACT ................................................................................................................... 12
CHAPTER
1 INTRODUCTION .................................................................................................... 15
Arsenic .................................................................................................................... 15 Arsenic in the Environment ..................................................................................... 16
Arsenic in Soil ................................................................................................... 17 Arsenic Toxicity ....................................................................................................... 18
Arsenic in Plants ..................................................................................................... 19 Remediation of As-Contaminated Soil .................................................................... 23 Root Exudates and Phytate .................................................................................... 25
Disposal of As-Laden Biomass ............................................................................... 26
2 ARSENIC REMOVAL BY PTERIS VITTATA FROM CONTAMINATED SOILS: A LONG TERM STUDY .......................................................................................... 31
Arsenic Phytoremediation ....................................................................................... 31
Materials and Methods............................................................................................ 33 Soil Sampling and Characterization ................................................................. 33
Experiment Setup ............................................................................................. 33 Sequential Extraction ....................................................................................... 34 Quality Assurance and Statistical Analyses ...................................................... 34
Results and Discussion........................................................................................... 35
Changes in Soil Arsenic Concentrations After 10 Harvests .............................. 35 Plant Growth and Arsenic Uptake in P. vittata .................................................. 36 Soil As Removal by P. vittata from Different Fractions ..................................... 37
3 ARSENIC REMOVAL FROM AS-HYPERACCUMULATOR PTERIS VITTATA BIOMASS: COUPLING EXTRACTION WITH PRECIPITATION ............................ 46
Biomass Disposal ................................................................................................... 46 Material and Methods ............................................................................................. 48
Chemical Reagents and PV Biomass ............................................................... 48 Water-Soluble As in PV Biomass ..................................................................... 48 Optimization of As Extraction from PV Biomass ............................................... 49
Precipitation of Water-Soluble As from PV Biomass ........................................ 50
6
Statistical Analyses .......................................................................................... 51
Results and Discussion........................................................................................... 51
Water-Soluble As in PV Biomass ..................................................................... 51 Arsenic Extraction from PV Biomass ................................................................ 52 Optimization of Ethanol Extraction from PV Biomass ....................................... 53 Precipitation of Water-Soluble As from PV Biomass ........................................ 54
4 ARSENIC REMOVAL FROM AS-RICH BIOMASS OF AS-HYPERACCUMULATOR PTERIS VITTATA: COUPLING ETHANOL EXTRACTION WITH ANAEROBIC DIGESTION .................................................... 62
Anaerobic Digestion ................................................................................................ 62 Material and Methods ............................................................................................. 64
Chemical Reagents and P. vittata Biomass...................................................... 64 Methane Index Potential from P. vittata Frond Biomass ................................... 65
Arsenic Analysis in plant biomass and mass balance ...................................... 66 Precipitation of Water-Soluble As from PV Biomass ........................................ 67 Statistical Analyses .......................................................................................... 67
Results and Discussion........................................................................................... 68 Methane Production from P. vittata Biomass .................................................... 68
Arsenic Partitioning After Anaerobic Digestion ................................................. 69 Precipitation of Water-Soluble As from PV Biomass ........................................ 70
5 CONCLUSIONS ..................................................................................................... 78
APPENDIX
METAL LEACHABILITY FROM COAL COMBUSTION RESIDUALS UNDER DIFFERENT PHS AND LIQUID/SOLID RATIOS .................................................... 80
Coal Combustion Residuals .................................................................................... 80
Materials and Methods............................................................................................ 82
Chemicals Reagents and CCR Samples .......................................................... 82 SPLP and LEAF Tests ..................................................................................... 83 Quality Assurance ............................................................................................ 84
Results and Discussion........................................................................................... 84 Total Metal Concentrations in CCRs ................................................................ 84
Metal Concentrations in Fly Ash Based on SPLP ............................................. 85
Metal Concentrations in Fly Ash Based on USEPA LEAF Method 1313 .......... 86
As and Se Concentrations ................................................................................ 87 Ba, Cr, Pb and Cd Concentrations ................................................................... 88 Metal Concentrations in Fly Ash Based on LEAF Method 1316 ....................... 90
Research Findings .................................................................................................. 91
LIST OF REFERENCES ............................................................................................. 106
BIOGRAPHICAL SKETCH .......................................................................................... 119
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LIST OF TABLES Table page 2-1 Changes in soil properties in two As-contaminated soils (CCA and DVA)
before and after 5 years of growth with P. vittata amended with phosphate rock (PR) and P fertilizer (P) (n=3). .................................................................... 40
2-2 Comparison of P. vittata phytoremediation studies in As-contaminated soils ..... 41
3-1 Elemental concentrations in six month old PV frond biomass obtained from a long-term phytoremediation experiment of As–contaminated soil (n=3). ............ 57
3-2 Water-soluble arsenic (%) and solution pH as a function of time and extraction method from PV frond biomass (shaking and no shaking) (n=3). ...... 57
3-3 Characterization of ethanol extraction effluent (n=3). ......................................... 58
4-1 Characterization of non-extracted and ethanol-extracted P. vittata biomass obtained from a long-term phytoremediation experiment of As–contaminated soil.. .................................................................................................................... 72
4-2 Cumulative methane (CH4) yield (LNCH4/kgVS) for ethanol-extracted P. vittata biomass with and without arsenic. Treatments are: treated control-biomass (TCB –No As) and treated As-rich biomass (TAsB) (n=3). ................... 73
A-1 Total concentrations of trace metals in 24 coal combustion residual samples from 7 representative power plants (mg/kg) ....................................................... 92
A-2 Concentration range of trace metals in coal combustion residuals based on literature (mg/kg). ............................................................................................... 93
A-3 Total concentrations of trace metals in 24 coal combustion residual samples from 7 representative power plants (mg kg-1) ..................................................... 94
A-4 pH of 24 coal combustion residual samples from 7 representative power plants before and after SPLP test ....................................................................... 95
A-5 SPLP concentrations of trace metals in 24 fly ash, bottom ash and FGD samples from 7 representative power plants (µg L-1). ......................................... 96
A-6 USEPA LEAF 1313 concentrations of trace metals in 8 fly ash samples from 7 representative power plants (n=3). .................................................................. 97
A-7 USEPA LEAF 1316 concentrations of trace metals in 8 fly ash samples from 7 representative power plants (n=3). ................................................................ 100
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LIST OF FIGURES
Figure page 1-1 Scheme of soil P uptake by plants and soil P distribution (Schachtman et al.,
1998) .................................................................................................................. 30
2-1 Frond biomass (AB) and As concentrations (CD) in P. vittata fronds during 5 years of growth in two As-contaminated soils (CCA and DVA). PR = phosphate rock and P = P fertilizer.. ................................................................... 42
2-2 Correlation of soil As concentration and frond As accumulation during 5 years of phytoremediation of two As-contaminated soils (CCA and DVA).. ........ 43
2-3 Roots biomass production during 3.5 years of phytoremediation of two As-contaminated soils (CCA and DVA) with P. vittata. PR= phosphate rock and P= P fertilizer.. .................................................................................................... 44
2-4 Arsenic distribution in the soluble (S), exchangeable (E) , amorphous (A), crystalline (C) and residual (R) fractions over 4.5 years in two contaminated soils (CCA-AC and DVA-BD) with P. vittata. PR= phosphate rock (AB) and P= P fertilizer (CD).............................................................................................. 45
3-1 Water-soluble arsenic from PV biomass as a function of time and extraction method ............................................................................................................... 58
3-2 Arsenic extraction from PV frond biomass using different extractants followed by HCl extraction. ............................................................................................... 59
3-3 Effect of time (a), particle size (b), solid-to-liquid ratio (c) and pH (d) on As extraction from P. vittata biomass using 35% ethanol. ....................................... 60
3-4 Effect of Mg salts (a), As:Mg ratio (b) and pH (c) in Mg–As precipitation. .......... 61
4-1 Cumulative methane yield (LNCH4/kgVS) for P. vittata control and Asbiomass with ethanol extraction.. ...................................................................................... 74
4-2 Initial and final volatile solids (%) (A) and remaining P. vittata biomass (B) after 35 d of anaerobic digestion ........................................................................ 75
4-3 Ethanol-extracted PV biomass arsenic partitioning among gas, solid and liquid phase (A) and As solid phase fractionation (PV biomass + precipitate) (B) after 35 d of anaerobic digestion (n=3). ........................................................ 76
4-4 Solution arsenic removal as As-Mg precipitate using MgCl2, As:Mg ratio of 1:400 and pH 9.5.. .............................................................................................. 77
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A-1 Total concentrations of trace metals in 24 fly ash, bottom ash and FGD from 7 power plants .................................................................................................. 102
A-2 SPLP concentrations of trace metals in 24 fly ash, bottom ash and FGD samples from 7 representative power plants. ................................................... 103
A-3 USEPA LEAF 1313 concentrations of trace metals in 8 fly ash samples from 7 representative power plants.. ......................................................................... 104
A-4 USEPA LEAF 1316 concentrations of trace metals in 8 fly ash samples from 7 representative power plants. .......................................................................... 105
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LIST OF ABBREVIATIONS AsIII Arsenite
AsV Arsenate
CCA Chromated Copper Arsenate
CCRs Coal Combustion Residues
CEC Cation Exchange Capacity
DI Deionized Water
DMA Dimethylarsinate
DOC Dissolved Organic Carbon
DVA Dipping Vat Soil
DVB Dipping Vat Soil B Horizon
FGCTL Florida Groundwater Cleanup Target Level
FGD Fuel Gas Desulfurization
FSCTL Florida Soil Cleanup Target Level
GF-AAS Graphite Furnace Atomic Absorption Spectrophotometry
H Hour
ICP-MS Inductively Coupled Plasma Mass Spectrometry
IP6 Myo-Inositol Hexakisphosphate
L/S Liquid to Solid Ratio
LEAF Leaching Environmental Assessment Framework
LMWOA Low Molecular Weight Organic Acid
MCL Maximum Contaminant Level
MIP Methane Index Potential
MMA Monomethylarsonate
Pi Inorganic Phosphorus
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PR Phosphate Rock
PV Pteris vittata
QA/QC Quality Assurance/Quality Control
S:L Solid to Liquid Ratio
USEPA United State Environmental Protection Agency
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Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy
PHYTOREMEDIATION OF ARSENIC CONTAMINATED SOILS BY PTERIS VITTATA:
ARSENIC REMOVAL AND BIOMASS DISPOSAL
By
Evandro Barbosa da Silva
May 2018
Chair: Lena Q. Ma Co-chair: Ann C. Wilkie Major: Soil and Water Science
Pteris vittata (PV) is the first-known As-hyperaccumulator, which not only
hyperaccumulates arsenic but also has the ability to extract insoluble As and P from
soils. In this study, its ability to continuously remove As from three contaminated soils
(26–126 mg kg−1) over 5 years was investigated and its biomass disposal methods were
optimized.
In the first experiment, the goal was to evaluate PV’s ability to continuously
remove As during 10 harvests and identify how soil As was affected by P availability.
The highest frond biomass production was 58.6, 51.9 and 42.4 g plant-1 year-1 in the
10th harvest, and frond As concentration decreased until replanting after which it
increased. Arsenic removal from soils averaged 10-15% of total soil As per harvest
during 1-6th harvests, which reduced to 3% during 7-10th harvests. All fractions were
affected by PV, except the residual fraction. The largest reduction occurred in the
amorphous (66%, from 61.2 to 21.8 mg kg-1) in the CCA soil and crystalline (61-85%,
from 1.99-4.35 to 0.61-0.75 mg kg-1) in DVA and DVB soils.
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In the second experiment, the goal was to optimize the parameters to extract As
from As-rich PV biomass. Water-soluble As varied from 6.8% to 61% in the biomass
depending on extraction time, with 99% of As being arsenate (AsV). Extraction with
2.1% HCl, 2.1% H3PO4, 1 M NaOH and 50% ethanol recovered 81, 78, 47 and 14% of
As in the soluble fraction. A follow-up extraction using HCl recovered additional 27-32%,
with ethanol recovering only 5%. Though ethanol showed the lowest extractable As, the
residual As in the biomass was also the lowest. Among the extractants, 35% ethanol
was the best to extract As from PV biomass. Approximately ~90% As was removed
from PV biomass using particle size < 1 mm at solid:liquid ratio 1:50 and pH 6 for 2 h.
Adding MgCl2 at As:Mg ratio of 1:400 with pH 9.5 was effective to precipitate soluble As,
resulting in 98% removal.
In the third experiment, the goal was to assess As partitioning among the three phases
(gas, liquid and solid) during anaerobic digestion of As-rich biomass of P. vittata. The
PV biomass was extracted with ethanol using particle size < 1 mm at solid:liquid ratio
1:50 and pH 6 for 2 h. Then, PV biomass with and without As and/or extraction were
digested at 35°C under anaerobic condition for 35 d. Liquid-displacement method was
used to measure methane gas production. Methane production was 145-160
LNCH4/kgVS (volatile solids basis) after 35 d for ethanol treated As-rich biomass (TAsB)
and control biomass (TCB), respectively. After digestion, volatile solids decreased from
93.9 and 94.1 to 14.9 and 17.5%TS (total solid) while PV biomass was reduced by 70.8
and 64.4% for TAsB and TCB, respectively. Ethanol extraction followed by anaerobic
digestion decreased As concentration in PV biomass by ~98%. At this level, PV
biomass would be considered a safe material by USEPA regulations. As a last step,
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51% of As in anaerobic digestate was recovered by As–Mg precipitation. Effective As
removal from PV biomass prior to disposal improves its phytoremediation process.
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CHAPTER 1 INTRODUCTION
Arsenic
Arsenic has atomic number of 33 and belongs to group 15 in the periodic table,
same group as phosphorus (P) and antimony (Sb). It can be found mainly as arsenates
and sulfides minerals (Wedepohl, 1969). Besides, arsenic has four oxidation states: -3,
0, +3 and +5. Arsenate (AsV) and arsenite (AsIII) are the predominating oxidation states
in the environment (IARC, 2004). In addition, AsV and AsIII are present as oxyanions
(H3AsO4 and H3AsO3) in solution with pKa’s of 2.2, 6.9, 11.5 and 9.2, 12.1, 13.4,
respectively (Goldberg and Johnston, 2001). Hence, at neutral pH, the dominating
species are H2AsO4-, HAsO4
2- and H3AsO3º. Arsenic availability in soils depends on its
oxidation states.
Arsenic is known for its high toxicity to plants, animals and humans. Due to its
carcinogenicity, it is listed as the #1 hazardous substance (ATSDR, 2007). Besides, it is
estimated that worldwide 36 million people live in As-contaminated areas. Arsenic
toxicity and speciation are affected by soil redox potential, pH and soil microbes activity
(USEPA, 2001).
Arsenic contamination is a serious problem in the US where soil in more than
600 areas are contaminated with As (USEPA, 2017). In fact, As ranks second most
frequent contaminant requiring to be cleaned up at the Superfund sites (USEPA, 1996)
Besides, As contaminated soils are the main exposures paths to food chain and
drinkable water (Frankenberger and Arshad, 2002). Other exposures paths include use
of pesticides, insecticides, defoliants, wood preservatives, and soil sterilants (Alves et
al., 2015; da Silva et al., 2018b; Gress et al., 2015)
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Thus, it is necessary to develop effective remediation techniques to reduce its
adverse effects (Lessl and Ma, 2013). Arsenic remediation techniques include capping,
solidification and stabilization, vitrification, soil washing, soil flushing and
phytoremediation. They can be applied in situ or excavated and transported out of site
for treatment (USEPA 2002). Phytoremediation includes different techniques such as
phytostabilization, phytoimmobilization, phytovolatilization and phytoextraction.
Phytoextraction using hyperaccumulator Pteris vittata (PV - Chinese brake Fern) is
becoming popular due to its cost-effectiveness. PV can accumulate up to 23 g kg-1 As in
the fronds (Ma et al., 2001; Lessl et al, 2013). In addition, PV has the unique ability to
acquire phosphorus (P) from soil with low P availability. However, disposal of As-rich
biomass is still an issue.
Arsenic in the Environment
Due to its toxicity and contamination in the environment, As is of major
environmental concern. Due to runoff, As may end up in rivers or percolate into
groundwater. It may be absorbed by plants entering the food chain, causing risks to
human health (Gress, 2014).
Arsenic is naturally present in low concentrations in soils and rocks, occurring
naturally in over 200 different mineral forms, of which ~60% are arsenates and 20%
sulfides (Mandal and Suzuki, 2002). Anthropogenic activities such as use of pesticides,
fertilizers, mining activity, coal combustion, and wood treatment all contribute to As
elevation in soil and water, increasing its risk to humans and environment (da Silva
2018a; Roychowdhury et al, 2002).
Arsenic has been widely used as pesticides, animals feed additives and wood
preservative. In the U.S., until recently more than 100 tons of feed additives were used
17
per year (Cortinas et al., 2006). However, studies showed high As levels in chicken
breast meat was related to additives use, leading to its suspension. Arsenic has also
been used as wood preservative, known as CCA (chromated copper arsenic). Before its
banishment in 2004, CCA wood made up more than 90% of outside wooden structures
(USEPA, 2008). Arsenic concentration in CCA wood can be as high as 1,200 mg kg-1
and last ~50 years (Townsend et al., 2003). The As in the CCA-wood acts as an As
sources due to its weathering and degrading in the environment.
Arsenic toxicity in the environment depends on its oxidation state, redox potential
and pH. Besides, a strong association of As with Fe minerals has been observed.
Actually, Fe/Al (hydr)oxides play an important role in controlling As availability in soils
(Kabata-Pendias, Pendias, 2011; Waltham, Eick, 2004). Arsenate (AsV) and arsenite
(AsIII) are the most predominant forms in soils. Arsenate dominates in aerobic
environments (Xu et al, 2007; Wang et al. 2011). The occurrence of AsIII is
predominantly under anaerobic conditions. However, it also presents in plants
rhizosphere due to arsenate reduction by microbial activities or by plant roots.
Schmöger et al. (2000) reported that AsIII is predominant in rhizosphere of
contaminated mine areas.
Arsenic in Soil
Arsenate (AsV) and arsenite (AsIII) are the most predominant As forms in soils.
Arsenate has extensive and strong interaction with soil minerals, while AsIII interaction
depends on specific chemical conditions (Fendorf et al., 2010). Hence, AsV’s availability
is low at pH < 8.5. In soil, AsV forms surface complexes with Al and Fe oxides (Bohn et
al., 2002). In the outer-sphere complexation, lack of ligand-bridged connection allows
18
faster desorption, while inner-sphere desorption requires covalent bound cleavage
(Catalano et al., 2008).
In addition, organic matter (OM) also affects As retention in soil (Grafe et al.,
2002). Arsenate exhibit a stronger preference to OM than AsIII with amine (NH2) groups
being responsible for most As retention onto OM (Fendorf et al., 2010; Manning and
Goldberg, 1997). Besides, the presence of cations such as Fe, Al and Mn enhance AsV
adsorption onto OM as they act as bridging complexes (Lin et al., 2004). Besides, AsV
also can precipitate with other metals and substitute P and S in minerals (Rochette et
al., 1998; Smedley and Kinniburgh, 2002).
Conversely, AsIII tends to be weakly retained in soils (Fendorf et al., 2010).
Arsenite adsorption to most soil minerals are more labile, except to Fe hydroxides with
which AsIII forms binuclear complex (Ona-Nguema et al., 2005). Under reducing
conditions, AsIII availability is controlled by sulfide precipitates forming mineral such as
arsenopyrite (FeAsS), arsenic-rich pyrite (Fe(S,As)2), orpiment (As2S3) and realgar
(AsS) (Bostick and Fendorf, 2003).
Arsenic Toxicity
Human exposure to arsenic occurs through inhalation, ingestion and skin
absorption (Shiomi and Nriagu, 1994). Both inhaled and ingested As may damage the
gastrointestinal and respiratory system, though vascular circulation will spread it to other
organ as well (Vahter, 2000). Besides, As’s half-life in the body depends on its exposure
path, with inhaled As presenting the shortest half-life due to biotransformation in the
liver. In addition, chronical exposure may cause skin, bladder and lung cancer.
Moreover, As exposure during pregnancy can cause fetal death and/or malformation in
19
animals, including humans. Generally, As toxicity decreases following the sequence of
AsIII > AsV > monomethylarsonic acid (MMA) > dimethylarsinic acid (DMA).
Unlike other toxic elements, organic As compounds are usually less toxic than
inorganic (Cohen et al., 2006; Fowler, 1977). Among inorganic As, AsIII is more toxic to
most organisms because it has high affinity for sulfhydryl groups, thereby disrupting the
structure and function of proteins involved in cellular metabolism (Ali et al., 2009).
In the U.S., it is estimated that 8% of drinking water exceeds the As maximum
contaminant Level (MCL = 10 µg L-1) and that 14% exceeds 5 µg L-1 (Focazio et al.,
2000). Besides water consumption, rice, flour and seafood are the main As food source
with a daily dietary intake of 8.4-14 µg d-1 depending on age group (Francesconi and
Edmonds, 1996; Schoof et al., 1999). Moreover, low-moderate arsenic level in drinking
water is being associated to type 2 diabetes (Wilbur, 2000).
Arsenic in Plants
Plant exposure to arsenic can reduce its biomass production; however, some
species do not show toxicity symptom. It is possible those plants have developed
strategies to tolerate As by accumulating more in the roots (Lessl et al., 2013) or due to
their ability to decrease As absorption by changing phosphate absorption mechanism,
reducing AsV intake (Meharg and Macnair, 1992; Meharg and Hartley-Whitaker, 2002).
However, some plants can accumulate high As concentration, suggesting a more
efficient detoxification mechanisms.
Many plants take up both AsV and AsIII using different mechanisms (Danh et al.,
2014; Wang et al., 2002, 2011). Arsenate is taken up by phosphate transporters
following Michaelis-Menten kinetics, presenting direct competition with P (Wang et al.,
2002, 2011) and PV is no exception (Poynton et al., 2004; Wang et al., 2002). However,
20
P transporters have higher affinity to phosphate than AsV (Meharg and Macnair, 1992;
Zhao et al., 2009). Yet, AsV’s uptake is inhibited by phosphate in PV. Wang et al.
(2002) reported that when AsV and phosphate are in similar concentration, phosphate
decreased AsV depletion from the solution up to 80%. On the contrary, Poynton et al.
(2004) showed a 30% decrease when the same concentration was tested.
Phosphate transporters involve cotransport of phosphate and/or arsenate and
protons (Zhao et al., 2009). The mechanism stoichiometry comprises at least 2H+ for
each H2PO4- or H2AsO4
- moved (Ullrich-Eberius et al., 1989). So far, many phosphate
transporter were discovered, with 100 transporters from Phosphate 1 (Pht1) family
being characterized (Bucher et al., 2007). Pht1 family is predominantly expressed in
plants’ roots and its transcript increased in low P environment (Shin et al., 2004).
Inorganic P is taken up by either low and high affinity transport system (Marschner et
al., 1988). In Arabidopsis thaliana, two phosphate transporters from Pht1 family (Pht1;1
and Pht1;4) were identified, playing a major role in phosphate uptake in both low and
high P availability (Shin et al., 2004). A mutant of A. thaliana with down-regulation of
Pht1;1 was more resistant to AsV when compared to wild type, indicating an important
role of Pht1;1 in AsV absorption (González et al., 2005).
Arsenate influx in PV follows Michaelis-Menten kinetics, i.e., it is mediated by
arsenate binding to a discrete active site (Poynton et al., 2004). The Michaelis constant
(Km) found by Wang et al. (2002) was 0.52 and 0.97 µM in the presence and absence
of P, respectively. Yet, in Poynton et al. (2004) study, the Km was 1.1 and 6.8 µM in the
presence and absence of P, respectively. The difference can be attributed to difference
in growth and P contents.
21
The mechanisms of AsIII uptake in plants is poorly understood until recently with
most AsIII influx mechanism studies in microorganism (Zhao et al., 2009). In bacteria,
yeast and humans, aquaglyceporins were AsIII transporter (Bhattacharjee and Rosen,
2007). Aquaglyceporin is a subfamily of the aquaporin superfamily that presents large
pores, allowing neutral molecules to pass passively such as glycerol (Danh et al., 2014).
Besides, aquaporin channels can be divided into three categories depending on their
pore structure and selectivity for different substrates (Mitani et al., 2008).
Arsenite in the environment is present mostly as neutral specie and plants take
up AsIII by the aquaporin channels. In addition, AsIII is analogous to silicic acid,
therefore, their competitive inhibition was observed (Wang et al., 2011). A competition
experiment suggested that in rice AsIII was taken up by the aquaporin channels
(Meharg and Jardine, 2003). Besides, the addition of glycerol decreased AsIII
accumulation in the roots and shoots by 53-71% (Mathews et al., 2011), suggesting
transport by aquaglyceroporins.
However, inhibitory effect in PV was not observed when silicic acid, antimonite
(AsIII analog) and Hg (aquaporin inhibitor) were added, indicating that it may not be the
main AsIII influx channel (Wang et al., 2010; Nagarajan and Ebbs 2007). In PV, it
seems AsIII influx mainly occur through active process, with passive diffusion being a
small component (Wang et al., 2011). Test using 2,4-dinitrophenol, an active uptake
inhibitor, showed 96% reduction in AsIII transport into the roots. Thus, it indicates that
AsIII was probably also transported by active uptake.
Alternatively, major intrinsic protein (MIP) family also allows the movement of
water and/or small neutral solutes. Besides, MIPs can be classified as aquaporins (He
22
et al., 2016). In PV, a new aquaporin protein, the ‘PvTIP4;1, was identified as a
tonoplast intrinsic protein homolog (He et al., 2016). The author expressed the
‘PvTIP4;1 in yeast and in A. thaliana and noticed that ‘PvTIP4;1 transformants were
much more sensitive to AsIII and they accumulated more As, indicating a possible
mediation of this protein in AsIII influx in PV.
After uptake, AsV is efficiently reduced to AsIII, becoming the main species in PV
fronds (Danh et al., 2014). However, there is no consensus where reduction takes
place. Some suggested that AsV reduction occurs in the fronds (Ellis et al., 2006;
Kertulis et al., 2005; Tu et al., 2004) but others affirmed that it occurs in the roots
followed by efficient AsIII translocation to the fronds (Duan et al., 2005; Mathews et al.,
2011; Su et al., 2008). Glutathione, carotenoids, ascorbate and catalase have been
reported to mediate AsV reduction in PV (Singh et al., 2006; Duan et al., 2005). Singh et
al. (2006) showed that As increased levels of ascorbate and glutathione in PV as a plant
defense system against oxidative stressed. Besides, it seems that glutathione is
recycled from glutathione disulfide by glutathione reductase and it is not affected by As
even though it has high affinity to thiol groups (Kertulis-Tartar et al., 2006).
Arsenite complexation with phytochelatins (PCs) is one of the main detoxification
mechanisms (Zhao et al., 2009). In sunflower, up to 14 PCs compounds were identified
after the plant being exposed to As, with GS-AsIII-PC2 and AsIII-PC3 being the main
As-PC compounds (Raab et al., 2005). AsIII-PC3 was also the main complex in the
arsenic-tolerant H. lanatus (Raab et al., 2004).
In most plants, As translocation is limited due to AsV reduction in the roots
followed by thiol group complexation, sequestrating in the root vacuoles (Zhao et al.,
23
2009). For example, when arsenate reductase was shut down in A. thaliana, As
accumulation in the leaves increased (Dhankher et al., 2006). On the other hand, As
might be transported in the phloem as PCs and/or thiol peptide complexes (Chen et al.,
2006). However, that is not the case for PV. In PV, As complexation with thiol groups in
the roots is minimal. Most of the As in the xylem sap is AsIII (Su et al., 2008; Zhao et al.,
2009). Besides, arsenate reductase, Arr2P, is required to convert AsV to AsIII with
further exportation from cells by Arr3p. Arsenate reductases are a key point in As
detoxification in PV and its activity in PV roots are highly correlated with As
concentration in the fronds (Zhang et al., 2002; Duan et al., 2005). In PV, the gene
PvACR2 was highly homologous with Saccharomyces cerevisiae gene ScACR2 (Ellis et
al., 2006). The overexpression of arsenate reductase in Arabidopsis resulted in As
hypersensitivity, probably due to AsIII toxicity in the leaves (Dhankher et al., 2006). Yet,
arsenic in PV is not fully understood and research in the molecular area might help to
better understand this plant.
Remediation of As-Contaminated Soil
Soil remediation techniques are costly and disrupt the environment and
sometimes unsuitable for large areas. Among all remediation techniques, none can be
applied in all situations. Often an exhaustive investigation is done by considering the
site characteristics, risk to human health, timeframe and available budget. Arsenic
remediation techniques can be in situ or ex situ, including capping, solidification and
stabilization, vitrification, soil washing, soil flushing and phytoremediation (USEPA
2002).
Phytoremediation is a low-cost green technology that can remedy contaminated
soils and waters. This technique involves the use of plants and associated organisms to
24
degrade, control or reduce the contaminants in the environment (Accioly et al., 2000).
Therefore, understanding the As dynamics in a soil-plant system is essential to
maximize remediation efficiency. Phytoextraction uses hyperaccumulator plants to
remove the contaminant from a soil (Lasat, 2002). Hyperaccumulator plants have the
ability to absorb and accumulate more than 1000 mg kg−1 of metal (Brooks et al., 1980).
The contaminants are accumulated in plant shoots, which can be collected, transported
and disposed off-site (Schnoor, 1997).
The advantages of phytoremediation include minimum impact in the area, low
environmental disturbance and favorable aesthetics (Nedelkoska and Doran, 2000).
However, phytoremediation efficiency depends on soil property, contaminant
bioavailability, and plants’ ability to take away the contaminant (Khan et al., 2000).
Therefore, it is important to understand soil-metal-plant specific interaction to improve
phytoremediation.
PV is the first As-known hyperaccumulator (Ma et al., 2001), though other ferns
have also been identified. However, not all ferns have the ability to hyperaccumulate As
(Visoottiviseth et al, 2002; Zhao et al, 2002). Most hyperaccumulator ferns belong to the
Genus Pteris including Pteris cretica, P. Longifolia and P. umbrosa (Zhao et al. 2002).
However, Pteris tremulam and Pteris stramina are not As hyperaccumulator (Meharg
and Jardine, 2003).
Phytoremediation success depends on two factors: (1) identify a plant able to
hyperaccumulate and (2) and know the optimal conditions for maximum removal (Tu et
al., 2003). For example, PV efficiently decreased As in groundwater in 3 days, reducing
As concentration from 46 to <10 µg L-1. However, P reduced PV ability to uptake As.
25
Besides, plant age also affected its efficiency with younger ferns being more effective
(Tu et al., 2004).
Root Exudates and Phytate
It is known that plants usually cannot uptake organic P. In soil, available
inorganic P content is low compared to plant requirements. For example, inorganic P
concentration in soil solution is ~10 µM (Figure 1-1) (Schachtman et al., 1998; Bieleski,
1973). On the other hand, organic P represents 20 to 80% of total soil P and phytate is
typically the major component (Richardson, 1994).
P. vittata is capable of accessing insoluble P in both acid and alkaline soils by
producing more adventitious roots and root hairs (Lessl and Ma, 2013). In addition, it
was noticed that P concentration in PV rhizosphere pore waters increased compared to
plantless soil (Lessl et al., 2013). Besides, large root systems enhances roots exudates,
resulting in rhizosphere acidification. In calcareous soil, roots exudates can affect P
associated with Ca, releasing P for plant uptake. Besides, it is possible that PV was
efficient to exudate low molecular-weight organic acids (LMWOA) while growing in As-
contaminated sites.
In the rhizosphere, plant roots release LMWOA (Tu et al, 2004). These LMWOA
can dissolve phosphates minerals via anion exchange and chelation with Fe and Al ions
associated with P (Johansson et al., 2008). For most plants, citric, oxalic and malic
acids are the main constituents of LMWOA. However, in PV rhizosphere, significant
amounts of phytate were identified besides oxalic and malic acid (Tu et al, 2004).
However, it is unclear why phytate concentration was high in the PV rhizosphere. It was
shown that a synergic effect between phytate and siderophores increased Fe uptake by
PV. Besides, phytate enhanced AsV and P transport by P. vittata (Liu et al., 2017a).
26
Both phytate and oxalic acid were effective in solubilizing As from AlAsO4,
FeAsO4 and CCA-contaminated soil (Liu et al., 2017b; Tu et al., 2004). Besides, since
As and P are analogues, it is possible that high phytate concentration in PV rhizosphere
affects As behavior in soil, increasing its availability (Tu et al, 2004). Therefore,
understanding the relation between LMWOA and PV is important to comprehend its
mechanisms of As accumulation.
Arsenic and P have similar chemical properties, competing for soil sorption and
plant uptake (Adriano, 2001). Low P levels helps plant to take up more As, and addition
of soluble P to soil causes more As release (Lessl et al., 2013; Zhang et al., 2002).
Besides, under P deficiency, plants exude LMWOA into soil, solubilizing P by changing
soil pH and P displacement. Furthermore, LMWOA may also complex with metal-P
compounds (Kirk et al., 1999; Neumann and Römheld, 1999). Likewise, LMWOA also
changes As availability in a similar manner.
PV’s root exudation plays an important role in As mobilization in soils. Besides, in
low-P environment, PV phytase activity was increased, resulting in hydrolysis of phytate
(Lessl et al., 2013). The increase in phytase activity was also reported in Polygonum
hydropiper rhizosphere. However, increase in phytate concentration was due to manure
application (Giles et al., 2011). For PV, it is possible it exudated phytase to hydrolyze phytate,
enhancing P availability. Therefore, if a plant possesses the ability to solubilize P from
phytate, it has the advantage to survive in As-contaminated soils.
Disposal of As-Laden Biomass
Proper husbandry practices can increase the potential of PV to remediate As-
contaminated sites. PV can produce 1.5 t ha-1 of frond biomass with As concentrations
27
up to 4,500 mg kg-1 (Lessl and Ma, 2013; Kertulis-Tartar et al., 2006). However, the
contaminated biomass disposal might represent a drawback of phytoremediation using
PV.
Conventional methods for biomass disposal use hazardous waste landfills.
However, improper disposal of arsenic-rich biomass may create further environmental
problems. The USEPA guidelines consider material with As concentration > 100 mg kg-1
unsafe. Thus, transferring large hazardous waste volumes and the costs incurred for
disposal dampens the advantages of using PV for phytoremediation. Other methods
have been used for treating biomass including pre-treatments (compaction and
pyrolysis) and post-treatments (incineration, ashing, and liquid extraction).
Conventional method to discard biomass is by incineration (Bondada and Ma,
2003). However, As combustion may release toxic As into the air (Cullen and Reimer,
1989). The other option is landfill deposition. Nonetheless, around 80% of total As in PV
biomass is water soluble (Tu et al., 2003). The annual biomass production of PV
biomass is 1.03 t (dw) ha−1 and As accumulation in PV frond can be as high as 23,000
mg kg-1 dw (Kertulis et al., 2005; Tu and Ma, 2002). Based on that, the potential As
leaching in landfill would be ~19 kg of As per hectare. Therefore, improper landfill
management would pose a risk to the environment, especially groundwater
contamination.
Composting is an effective strategy that could reduce the amount of biomass and
the cost of transport and disposal. However, As leaching and volatilization is of concern
due to arsenic reduction by microbes. This process occurs through methylation where
AsV is converted to AsIII or to organic forms (Cao et al., 2010; Turpeinen et al., 1999).
28
On the other hand, composting can stabilize metal in the biomass, reducing its available
fractions, especially water-soluble fraction (Singh and Kalamdhad, 2013). However,
some studies indicated composting increased metal availability due to the presence of
organic compounds (Greenway and Song, 2002). Therefore, it is necessary to assess
As transformation in PV biomass.
Anaerobic digestion is a complex system where symbiotic microbes transform
organic materials into biogas under anoxic environment, leaving refractory organic
matter (Wilkie, 2008). However, metal availability is affected by contaminants and
components in biomass, and redox potential (Kuo et al., 2004). A study indicated that
As was immobilized in PV biomass during composting (Cao et al., 2010). They found
that total and water-soluble As concentration was reduced by 25-32%. Other study
showed the low occurrence of As volatilization in soil amended with organic material
(Moreno-Jiménez et al., 2013). Organic matter can be used as energy source, providing
favorable conditions for As reduction (Cao et al., 2003; Balasoiu et al., 2001). Cao et al.
(2003) attributed 16% As loss to microbial-mediated arsenic volatilization in the soil. On
the other hand, anaerobic digestion of As-contaminated sludge accounted <1% of
volatile species (Cortinas et al., 2006). Thus, anaerobic digestion may be used as to
pretreat As-rich biomass.
However, after anaerobic digestion, As in the effluent needs to be further treated.
Several techniques to remove As from aqueous media have been developed, for
example, by adsorption, electrocoagulation, membrane permeation and biological
methods (Kumar et al., 2004; Gecol et al., 2004; Sato et al., 2002).
29
These techniques are efficient in removing AsV from aqueous media and they
are used to treat polluted aqueous media. However, removal efficiency of AsIII is low
and the applicable concentration range is narrow. Besides, adsorbents cannot be
reused, becoming contaminated residues requiring proper disposal.
Another technique is As-metal precipitation. Precipitation can be accomplished
using different metal–As species over a wide pH range (Bothe and Brown, 1999;
Raposo et al., 2004). For example, Ca(OH)2 addition efficiently removes AsIII as high as
1 g/L via formation of Ca5(AsO3)3(OH).4H2O and Ca5(AsO3)3OH (Itakura et al., 2007).
Other elements such as Ba and Mg can also precipitate AsV (Raposo et al., 2004). The
removal of AsV by precipitation of Mg–As effectively removed As from liquor samples
(Park et al., 2010) with As concentration of 469 mg L-1.
The objectives of this research were to: (1) investigate PV’s ability to
continuously remove As from three contaminated soils over 5 years and 10 harvests (2)
to optimize As removal from PV biomass by testing different extractants, extraction
times, particle sizes and pH; and 3) to assess As removal and biomass degradation
under anaerobic condition.
30
Figure 1-1. Scheme of soil P uptake by plants and soil P distribution (Schachtman et al., 1998)
31
CHAPTER 2 ARSENIC REMOVAL BY PTERIS VITTATA FROM CONTAMINATED SOILS: A LONG
TERM STUDY
Arsenic Phytoremediation
Arsenic (As) is carcinogenic and toxic to humans, and is ranked as the top
pollutant in the environment (ATSDR, 2017). Humans are exposed to As through
consumption of contaminated food and water, and incidental ingestion of soils (Gress et
al., 2016, 2015). Arsenic is mainly present in soils as the inorganic forms, i.e., arsenite
(AsIII) and arsenate (AsV). Under aerobic conditions, AsV is the predominant form in
soils, which can be sorbed by Al/Fe oxides and Ca/Mg carbonates (Bissen and Frimmel,
2003; Bohn et al., 2002).
Arsenic is present in soil at low concentrations, ranging from 0.1 to 67 mg kg-1
(Mandal and Suzuki, 2002). However, anthropogenic activities have increased As levels
in soils (da Silva et al., 2018b), with ~36 million people worldwide living in As-
contaminated areas (ATSDR, 2007). Among sources, As-treated wood represents a
threat to public health (Gress et al., 2015; Lessl and Ma, 2013). Before being banned in
2004, >90% of outdoor wood products were made of As-treated wood (USEPA, 2008),
which may increase soil As at 0.5–1.2 mg kg-1 annually (Lebow et al., 2004).
Contaminated soils need to be remediated, but it is costly and time-consuming (Belluck
et al., 2003).
Phytoremediation is a low-cost technology that utilizes hyperaccumulator plants
to remove metals from soils. Adapting to contaminated soils, hyperaccumulator plants
have developed tolerance to As by accumulation (Chen et al., 2017; Ma et al., 2001).
Pteris vittata (Chinese brake fern) is the first known As-hyperaccumulator. It can
32
accumulate up to 23 g kg-1 As in the fronds with rapid As translocation (da Silva et al.,
2018a; Danh et al., 2014; Tu and Ma, 2002).
To grow in nutrient-poor soils, P. vittata developed the ability to solubilize
insoluble P including phosphate rock (Lessl et al., 2014). As analogs, P competes with
As uptake in plants. As such, P. vittata is more efficient in taking up As in low-P soils,
which enhances root growth and exudation (Lessl and Ma, 2013; Santos et al., 2008).
Root exudation from P. vittata helps to solubilize As and P from soils, making it efficient
to extract non-labile As and P from soils (Liu et al., 2017; Gonzaga et al., 2006).
Arsenic availability in soils has been assessed using sequential extraction, which
separates it into different fractions including soluble, exchangeable, amorphous,
crystalline and residual, in decreasing order of availability (Wenzel et al., 2001). While
soluble plus exchangeable fractions are labile and available for plant uptake,
amorphous and crystalline forms are non-labile and the residual fraction is the most
recalcitrant. During plant uptake, soluble and exchangeable fractions can be
replenished slowly from non-labile fraction (Fitz et al., 2003; Hinsinger et al., 2005).
Thus, assessing changes in As fractionation in soil helps to evaluate its efficiency in
phytoremediation.
In 2009, a long-term experiment was initiated to investigate P. vittata’s ability to
continuously remove As from contaminated soils (Lessl et al., 2014). The study showed
that maintaining low P levels in soils enhanced As removal over 3.5 years and 7
harvests (Lessl et al., 2014). Built upon that study, this study was designed to evaluate
P. vittata’s ability in continuing to remove As from soils over 5 years. Specifically, our
objectives were to (1) investigate P. vittata’s ability to continuously remove As from two
33
contaminated soils over 10 harvests, (2) identify the impact of P status on As removal
from soils, and 3) examine the changes in As-fractions in soils impacted by As uptake
by P. vittata over 5 years.
Materials and Methods
Soil Sampling and Characterization
Two As-contaminated soils (A horizon) were collected from contaminated areas.
One soil was from an abandoned As-treated wood facility (CCA, Grossarenic Paleudult)
and one was from an abandoned cattle-dipping vat (DVA, Arenic Albaqualfs) (Lessl et
al., 2014). The soils were air-dried, and sieved through a 2 mm sieve. Initial As
concentration was obtained through HNO3/H2O2 digestion (USEPA Method 3050B) on a
hot block (Environmental Express, Ventura, CA). The supernatants were filtered (0.45
μm) and analyzed for total As concentration using inductively coupled plasma mass
spectrometry (ICP-MS; Perkin-Elmer Corp., Norwalk, CT).
Experiment Setup
An experiment was initiated in December 2009 to investigate P. vitatta’s ability to
remediate As-contaminated soils (Lessl et al., 2014). For the experiment, raised beds
(0.36 m² by 35 cm) were constructed and filled with soils. While phosphoric rock (PR; 15
g kg−1) was added to half of the beds, P fertilizer was added to the other half (P) (four
beds soil−1 treatment−1). Three-month old P. vittata were planted (15 cm between
plants) at 9 plants per bed. For the PR treatment, P-free fertilizer (N:P:K 10:0:10) was
added, while P fertilizer was added (N:P:K 6:4:6) to the P treatment bimonthly. Hydrated
lime was applied in the DVA soil (40 g bed-1) to raise pH to ~6. Every six months,
mature fronds were harvested together with soil samples. Experiment data from 1-7th
harvests during January 2010 to January 2013 were discussed by Lessl et al. (2014).
34
Three additional harvests were made (8-10th) for the long-term experiment until
December 2014. After growing for 5 years from 2010 to 2014, P. vittata plants were
root-bound.
Sequential Extraction
Soil As was separated into five fractions including soluble, exchangeable,
amorphous, crystalline and residual fractions based on sequential extraction (Wenzel et
al., 2001). Briefly, 2.0 g of soil was weighed. In between extractants, soil samples were
centrifuged for 20 min at 5000 × g, with the supernatant being collected for analysis and
suspending the soil for the following extraction. Extraction steps included: soluble –
shaken for 4 h in 25 mL 0.05 M (NH4)2SO4; exchangeable – shaken for 16 h in 25 mL
0.05 M NH4H4PO4; amorphous– shaken for 4 h in the dark in 25 mL 0.2 M NH4-oxalate
buffer at pH 3.25; crystalline– shaken for 30 min in 25 mL 0.2 M NH4-oxalate buffer +
0.1 M ascorbic acid (pH 3.25) at 96oC; and residual – digestion using HNO3/H2O2
(USEPA Method 3050B). Arsenic in each fraction was analyzed by ICP-MS.
Quality Assurance and Statistical Analyses
QA/QC in samples analysis included blanks, spikes and triplicates for every 20
samples. Recovery was determined using spikes, and relative standard deviations of
triplicate analyses were obtained. The performance of ICP-MS was checked by running
an intermediate calibration standard for every 20 samples. All calibration standard
checks were within the acceptable range (80–120%).
All data are presented as the mean of three replicates with standard deviation.
Significant differences were determined by using one-way analysis of variance
(ANOVA) and treatment means were compared by Tukey's multiple range tests at p <
0.05 using software (R3.2.2) (Team, 2005). The relationship between frond As uptake
35
and soil As concentration for each fraction was compared using simple regression
analysis.
Results and Discussion
Changes in Soil Arsenic Concentrations After 10 Harvests
The two As-contaminated soils texture were loamy sand and sand (for CCA and
DVA, respectively), typical of Florida, which received two P treatments: P-limiting
(phosphate rock-PR) and P-sufficient (P fertilizer-P) (Table 1; Lessl and Ma, 2013).
Liming increased DVA soil pH from 5.1 to 6.1 as P. vittata prefers to grow in alkaline
soils (Table 1). Initial As concentrations in soils (26.7 and 129 mg kg-1 for DVA and
CCA) exceeded the Florida Soil Cleanup Target Level for residential soils (FSCTL = 2.1
mg kg-1) by 12-61 times. Available As in soils was much lower at 1.75 and 3.57 mg kg-1
for DVA and CCA. After growing for 5 years, rhizosphere pH was lower than bulk soil
pH in all treatments, reducing from 6.1 to 5.8-5.9 (DVA) and 7.4 to 6.9-7.1 (CCA).
Amendment with PR did not increase pH nor increase available P in soils (Table 1).
Unlike pH, available Ca was much greater in PR treatments, probably due to PR
dissolution by P. vittata. However, available P was greater in P treatments, reflecting
soluble P from fertilizer addition (Table 1). This was consistent with the P level in
porewater showing undetectable P levels in PR treatment and 0.25 mg L-1 in P
treatment (Lessl and Ma, 2013). After 10 harvests, available As decreased by 59-63%
in CCA and DVA soils (3.57 to 1.52-1.44 and 1.75 to 0.64-0.66 mg kg-1) (Table 1),
which are all below the FSCTL for residential soils at 2.1 mg kg-1. Similarly, after 10
harvests, soil As concentrations decreased by 46-49% in CCA and DVA soils (129 to
68.9-70.1 and 26.7 to 15.6-16.8 mg kg-1) (Table 1).
36
Plant Growth and Arsenic Uptake in P. vittata
Plant growth can be divided into two periods: 1-5th and 6-10th harvests (Figure
1AB). During the first 5 harvests, biomass production increased with each harvest,
partially because plants were not fully grown as it took 12-24 months to reach maturity
(Yong et al., 2010). Plant biomass was stabilized during the 6-8th harvests, and
increased during the 9-10th harvests. In addition, during the first few harvests (8 for
CCA and 3 for DVA), plant biomass was greater with PR-treatment than P-treatment.
For example, during the 1st harvest for CCA and DVA soils, frond biomass was 20.1 and
22.9 g plant-1 year-1 in PR-treatment compared to 13.9 and 13.8 g plant-1 year-1 in P-
treatment (Figure 1AB). However, after 10 harvests, generally there was no significant
difference in biomass production. The fact that P. vittata was able to grow well in CCA-
soil near pH 7 for 5 years with only PR as its P supply indicated its ability to acquire
non-available P to sustain biomass production, which is rather unique among plants.
This was accomplished with lower P concentration in P. vittata biomass in PR-treatment
than P-treatment (1,957 vs. 2,132 mg kg-1) (data not shown).
For CCA and DVB soils, the highest biomass production occurred during the 9th
(62.1-63.9 and 35.6-63.5 g plant-1) and 10th harvests (58.6-60.7 and 51.9-57.1 g plant-1)
(Figure 1AB). Plant As uptake and soil As concentrations was negatively correlated in
CCA soil (R² =0.89-0.95) and DVA soil (R² = 0.36-0.59) (Figure 2). These data showed
that plant As uptake effectively decreased soil As concentration, which in turn led to
lower As uptake by plants (Figure 1CD). Compared to other studies, As removal was
1.1 – 5 times more efficient (Table 2). For example, the bioconcentration factors (ratios
of As concentrations in fronds to soils) were 20-33 in this study compared to 7.5-17 in
37
other studies (Shelmerdine et al., 2009; Gonzaga et al., 2008; Kertulis-Tartar et al.,
2006).
This impressive growth in a P-limited environment is unique for P. vittata (Lessl
et al., 2014), showing P. vittata’s ability to access insoluble P and As in alkaline soils.
This is possible due to production of more adventitious roots and root hairs, with larger
root surface area to release more organic acids including phytate (Lessl et al., 2013; Liu
et al., 2016). Larger root systems helped rhizosphere acidification compared to bulk
soils after 10 harvests (Figure 3; Table 1), probably due to enhanced roots. The fact that
the P concentrations increased in rhizosphere porewater (0.25 mg L-1) compared to bulk
soils (<0.01 mg L-1) in PR-treatment is consistent with our data (Lessl and Ma, 2013).
However, with decreasing As concentration in soils as more As was removed by
P. vittata, it became more difficult to acquire insoluble As from soils (Figures 1 and 2).
For example, As accumulation in the fronds decreased from 3,470 to 1,261 mg kg-1 as
As soil content decreased from 129 to 68.9 mg kg-1 in CCA soil in PR-treatment. Arsenic
accumulation in plants and soil As removal generally decreased after they reached a
plateau during the 6th harvest (Figures 1 and 2). Arsenic removal averaged 10% and 7-
9% per harvest for PR-treatment and P-treatment during the 16th harvests. With lower
soil As, the rate reduced to 3 and 6% for CCA-soil, with limited reduction for DVA-soil
during the 710th harvests. Our data indicated the limitation of using plants to remove
As from soils, i.e., with time, its effectiveness decreased due to a lower As pool in soils.
Soil As Removal by P. vittata from Different Fractions
To better understand plant As removal from soils, soil As was separated into
different fractions. In general, As distribution in the two soils followed the order: soluble
38
≤ exchangeable < residual ≤ crystalline < amorphous (Figure 4). The amorphous
fraction represented the largest pool in the soils and, together with the crystalline
fraction, they made up the non-labile fraction. In the CCA-soil, initial As concentration in
the crystalline+amorphous fraction was 63.3-80.3 mg kg-1 (49-62%) and the residual
fraction was 7.40-10.9 mg kg-1 (5-8%) (Figure 4). In comparison, the
crystalline+amorphous fraction in DVA soil was much lower (30-47%). P. vittata was
effective to extract As from all fractions except the residual fraction, especially from the
amorphous and crystalline Al-Fe oxide fraction (Figure 4). In the earlier harvests, P.
vittata was more efficient in depleting non-labile As in PR- than P-treatments. For
example, up to the 7th harvest for CCA-soil, As reduction in the amorphous fraction
under PR-treatment was 26% (61.2 to 45.3 mg kg-1) compared to 2% (61.5 to 60.3 mg
kg-1) under P-treatment. However, after 4.5 years, As depletion was similar in PR- and
P-treatments. For both treatments, the largest reduction occurred in the amorphous
fraction (61.2-61.5 to 20.8-21.8 mg kg-1 or 64-66%) for CCA-soil, while it was in the
crystalline fraction (2.18-4.35 to 0.61-1.10 mg kg-1 or 50-86%) for DVA-soil. For DVA-
soil under PR-treatment, P. vittata was more effective in removing As from the non-
labile fraction, but it was similar in CCA-soil. During the 4.5 years, soluble and
exchangeable fractions remained similar, probably due to As resupply from the non-
labile fractions.
Initially, P. vittata was more efficient in solubilizing soil As in PR-treatment than
P-treatment (Figure 1). This was because root biomass in PR-treatment was higher
than P-treatment up to 7th harvest. Effective As removal was probably associated with
more root exudate release from greater root biomass in response to P limitation (Figure
39
3). In addition, P. vittata biomass production (20.1-43.8 g plant-1) in PR-treatment was
greater than in P-treatment (13.9-31.7 g plant-1) up to the 8th harvest (Figure 1AB). In
both soils, larger root biomass enhanced root exudate release, which was supported by
decreased rhizosphere pH compared to bulk soil (Table 1), thereby increasing P and Ca
release from PR-treatment for both soils (from 1541 and 132 to 2042 and 664 mg kg-1
or 132 and 500 %) (Table 1). Other studies also showed the ability of P. vittata root
exudates to solubilize PR, increasing its P and Ca uptake (Fu et al., 2017). Liu et al.
(2016) also noted that P. vittata accumulated 2.6–26 times more As with root exudates.
The data suggest that P. vittata was effective in releasing root exudates to enhance As
release from soils. This was supported by an increase in dissolved organic carbon in P.
vittata rhizosphere porewater, which was 24.6 and 16.1 mg L-1 for PR- and P-treatment
(Lessl and Ma, 2013). However, as plants matured, similar As removal was observed
between the two treatments. In addition, As removal became more difficult within time
as more As was taken up by P. vittata (Figure 2).
40
Table 2-1. Changes in soil properties in two As-contaminated soils (CCA and DVA) before and after 5 years of growth with P. vittata amended with phosphate rock (PR) and P fertilizer (P) (n=3).
CCA CCA-PR CCA-P DVA DVA-
PR DVA-P
Initial¹ ------- After ------- Initial¹ ------- After -------
Bulk soil pH 7.2 7.4 7.4 5.1 6.1 6.1
Rhizosphere soil pH -- 6.9 7.1 -- 5.8 5.9
Total As (mg kg-1) 129 68.9 70.1 26.7 15.6 16.8
Available As (mg kg-1)2 3.57 1.52 1.44 1.75 0.64 0.66
Mehlich III P (mg kg-1) 75.5 7.51 23.5 24.3 14.6 24.6
Mehlich III Ca (mg kg-1) 1541 2042 829 132 664 362
Mehlich III Mg (mg kg-1) 115 275 228 18.2 66.0 135
¹ From Lessl and Ma, 2013; 2 0.05 mM ammonium sulfate
41
Table 2-2. Comparison of P. vittata phytoremediation studies in As-contaminated soils
Harvest Initial soil As
(mg kg-1)
Final soil As
(mg kg-1)¹
Frond biomass
(g plant-1 y-1)¹
Frond As
(mg kg-1) Bioconcentration factor²
CCA-this study1 129 68.9 40.5 2118 20
DVA-this study1 26.7 15.6 27.5 651 33
Shelmerdine et al. (2009) 132 131 11.2 60 7.5
Gonzaga et al. (2008) 110 99.5 40.0 492 8.4
Kertulis-Tartar et al. (2006) 234 149 7.36 3,880 17
¹ Average of 10 harvests (1-7th harvest from Lessl et al., 2014).² Ratios of As concentrations of fronds to soils
42
Figure 2-1. Frond biomass (AB) and As concentrations (CD) in P. vittata fronds during 5
years of growth in two As-contaminated soils (CCA and DVA). PR = phosphate rock and P = P fertilizer. Bars represent standard deviation (n=3).
43
Figure 2-2. Correlation of soil As concentration and frond As accumulation during 5 years of phytoremediation of two As-contaminated soils (CCA and DVA). Bars represent standard error (n=3). ** Significant at p < 0.01.
y = 791 - 14.7x + 0.27x²** R² = 0.95
6080100120140
600
1200
1800
2400
3000
3600
As soil concentration (mg/kg)
As a
ccu
mu
lati
on
in
th
e f
ron
ds (
mg
/kg
)
6080100120140
As a
ccu
mu
lati
on
in
th
e f
ron
ds (
mg
/kg
)
0
500
1000
1500
2000
2500
3000y = - 511 + 18.1x R² = 0.36
As soil concentration (mg/kg)
As soil concentration (mg/kg)
51015202530
As a
ccu
mu
lati
on
in
th
e f
ron
ds (
mg
/kg
)
0
200
400
600
800
1000y = - 559 + 50.1x** R² = 0.59
1215182124
0
400
800
1200
1600
2000y = - 1086 + 98.8x** R² = 0.89
As soil concentration (mg/kg)
As
ac
cu
mu
lati
on
in
th
e f
ron
ds
(m
g/k
g)
D: P-soil (DVA)
A: PR-soil (CCA) B: P-soil (CCA)
C: PR-soil (DVA)
44
Figure 2-3. Roots biomass production during 3.5 years of phytoremediation of two As-contaminated soils (CCA and DVA) with P. vittata. PR= phosphate rock and P= P fertilizer. Bars represent standard deviation (n=3). (Lessl et al., 2014).
CCA PR CCA P DVA PR DVA P
Ro
ot
we
igh
t (g
kg
-1 s
oil
)
0
5
10
15
20
25
30
35
H 3 H 4 H 7
45
Figure 2-4. Arsenic distribution in the soluble (S), exchangeable (E) , amorphous (A), crystalline (C) and residual (R) fractions over 4.5 years in two contaminated soils (CCA-AC and DVA-BD) with P. vittata. PR= phosphate rock (AB) and P= P fertilizer (CD). Bars represent standard deviation (n=3) (Partially published by Lessl et al., 2014).
46
CHAPTER 3 ARSENIC REMOVAL FROM AS-HYPERACCUMULATOR PTERIS VITTATA
BIOMASS: COUPLING EXTRACTION WITH PRECIPITATION
Biomass Disposal
Arsenic (As) is naturally present in low concentrations in soils, ranging from 0.1
to 67 mg kg-1 (Mandal and Suzuki, 2002). Anthropogenic activities such as the use of
pesticides and fertilizers, mining activity, coal combustion, and wood treatment have all
contributed to As elevation in soils, increasing its risk to humans (Gress et al., 2016;
Roychowdhury et al., 2002). It is estimated that ~36 million people worldwide live in As-
contaminated areas, making it the #1 hazardous substance on the USEPA priority list
(ATSDR, 2007). Arsenic toxicity in soils depends on its oxidation state, redox potential
and pH, with Fe/Al (hydr)oxides controlling As availability in soils (Walthan and Eick,
2004).
Human exposure to arsenic occurs via many pathways, with consumption of
contaminated food and water being the most important (Gress et al., 2016). Arsenic
exposure may cause various cancers. Generally, As toxicity decreases following the
order of arsenite (AsIII) > arsenate (AsV) > monomethylarsonic acid (MMA) >
dimethylarsinic acid (DMA) (ATSDR, 2007). In addition, plant exposure to arsenic can
reduce its biomass production. Plants have developed strategies to tolerate As by
accumulating it in the roots (Chen et al., 2017; Lessl et al., 2014), or changing
phosphate transporters to reduce its uptake (El-Zohri et al., 2015; Wang et al., 2015).
Reprinted with permission from da Silva, E.B., de Oliveira, L.M., Wilkie, A.C., Liu, Y. and Ma, L.Q., 2018. Arsenic removal from As-hyperaccumulator Pteris vittata biomass: Coupling extraction with precipitation. Chemosphere, 193, 288-294. Doi:10.1016/j.chemosphere.2017.10.116
47
However, some plants can accumulate high As concentrations, making them
hyperaccumulators (Ma et al., 2001).
Phytoremediation is a low-cost technology that utilizes hyperaccumulator plants
to extract metals from soil. Pteris vittata (PV; Chinese brake fern) is the first known As-
hyperaccumulator and it can accumulate up to 23,000 mg kg-1 As in the fronds (Ma et
al., 2001; Tu and Ma, 2002). The advantages of phytoremediation include minimum
disturbance of the area, low environmental impact and favorable aesthetics
(Nedelkoska and Doran, 2000). The efficiency of phytoremediation is affected by soil
properties, metal bioavailability, and the plants’ ability to accumulate metals (Krämer,
2005).
Proper husbandry practices enhance PV’s effectiveness to remediate As-
contaminated sites (Kertulis-Tartar et al., 2006; Lessl and Ma, 2013). P. vittata takes up
As and rapidly translocates it into the fronds, the main site of accumulation (Danh et al.,
2014; Yang et al., 2007). Though most of the As in the fronds is present as AsIII, it is
oxidized to AsV in dry biomass (Tu et al., 2003). The disposal of PV biomass can be a
drawback for its application in phytoremediation. Usually, the biomass is either disposed
at regulated landfills or incinerated (Chaney et al., 2007). Other methods include
compaction, pyrolysis, ashing, and liquid extraction; however, they require specialized
equipment and transport of large amounts of biomass. Therefore, an effective As
recovery method is needed.
Various chemicals including acid, base and chelate have been used to recover
As, with HCl and NaOH being efficient in solubilizing As from plants at 90-92% (Alam et
al., 2001; Jang et al., 2007; Sullivan et al., 2003). In addition, methanol and ethanol
48
have been used to extract As from plant biomass for speciation (Amaral et al., 2013;
Zhang et al., 2002). Zhao et al. (2015) showed that 1:3 ethanol:water was effective to
recover ~80% of As from PV fronds. Besides extraction, As can be precipitated as
Mg3(AsO4)2 (Park et al., 2010), which can be used to separate As from the solution.
The goal of this study was to optimize As removal from PV biomass. The specific
objectives were: 1) to optimize As removal from PV biomass by testing different
extractants, extraction times, particle sizes and pH; and 2) to recover soluble As by
precipitation with different Mg salts. Removing most of the As from PV biomass can
help to improve its application in phytoremediation.
Material and Methods
Chemical Reagents and PV Biomass
All chemicals were of analytical grade or better. Nitric acid (trace metal grade),
H2O2, NaOH, H3PO4, and ethanol were obtained from Fisher Scientific (Waltham, MA).
The Sep-Pak AccellPlus QMA Plus Short cartridges were obtained from Waters
Corporation (WAT020545, Milford, MA). Before use, all labware was washed and
soaked in 1 M nitric acid for 24 h and rinsed several times with DI water.
P. vittata biomass was obtained from a long-term phytoremediation experiment
with 126 mg kg-1 As in As–contaminated soil (Lessl et al., 2014). PV fronds were
harvested every six months, oven dried at 65 °C and shredded to < 2 mm size.
Concentrations of As and other elements in PV fronds are shown in Table 3-1. Residual
humidity was ~7%, and biomass weight was corrected.
Water-Soluble As in PV Biomass
To determine arsenic concentration, PV biomass was digested using
HNO3/H2O2 via USEPA Method 3050B on a hot block (Environmental Express, Ventura,
49
CA). Briefly, 0.5 g of dried plant biomass was suspended in 15 mL 1:1 nitric acid and
heated at 105 °C for 6 h. After cooling, 1 mL 30% H2O2 was added and digested for an
additional 30 min before bringing samples to a 50 mL volume with DI water. Arsenic
concentration was analyzed using inductively coupled plasma mass spectrometry (ICP-
MS, Perkin-Elmer Corp., Norwalk, CT).
Water-soluble As in dried PV biomass was extracted using double-distilled water
(pH~6.5) at a solid:liquid ratio of 1:25 in 50 mL plastic bottles for 2, 4, 8, 12, 16 or 20 h.
Half of the samples were shaken at 60 rpm in a rotary shaker and the other half were
kept static. Speciation of water-soluble As was determined after centrifugation at 4200
rpm for 15 min and filtration with Whatman N. 42 filter paper using an As speciation
cartridge (WAT020545, Waters Corporation, Milford, MA), which retains arsenate
(Mathews et al., 2011). Total As and AsIII was determined using ICP-MS, with AsV
being the difference between total As and AsIII.
Optimization of As Extraction from PV Biomass
Preliminary studies showed that 80 °C was most efficient to extract As from PV
biomass (data not shown), so all experiments were conducted at 80 °C. Based on the
literature (Alam et al., 2001; Jang et al., 2007; Sullivan et al., 2003), the following
extractants were chosen: 2.1% HCl, 2.1% H3PO4, 1 M NaOH and 50% ethanol, which
was followed by a second extraction using 2.1% HCl. To extract As, 1 g of PV dry
biomass was placed in a 50 mL vial. After adding 25 mL of extractant, the samples were
placed in a water bath at 80 °C for 15 h and supernatant was collected after
centrifugation at 4200 rpm for 15 min and filtration with Whatman N. 42 filters. For the
second extraction, after adding 25 mL of HCl, the samples were placed in a water bath
50
at 80 °C for 2 h. Arsenic concentration was determined using ICP-MS after
centrifugation at 4200 rpm for 15 min and filtration with Whatman N. 42 filters.
After finding that ethanol was the best extractant, efforts were made to optimize
its efficiency. Different conditions were tested, including extraction time (0.5, 2, 8, 15 or
24 h), particle size (<1, <2 or >2 mm), solid:liquid ratio (S:L, 1:25, 1:50, 1:100, 1:200 or
1:300) and pH (6, 7 or 8). Extraction time was optimized using particle size < 1 mm PV
dry biomass and 35% ethanol at S:L 1:25 and pH 6. Similarly, particle size was
determined based on optimal extraction time and 35% ethanol at S:L 1:25 and pH 6.
Optimal S:L ratio was assessed using the optimized time and particle size, and 35%
ethanol at pH 6. Finally, optimal pH was determined using the optimal time, particle size
and S:L ratio, and 35% ethanol.
Precipitation of Water-Soluble As from PV Biomass
During As extraction using ethanol solution, colloidal precipitate was formed,
lowering As concentration in solution (data not shown). Ethanol has been used in
chlorophyll extraction, which is denatured above 60 °C (Ritchie, 2006). Precipitate was
inferred by stoichiometry (data not shown) to be As-Mg precipitate, mostly probably to
be magnesium arsenate (Mg3(AsO4)2). To optimize As-Mg precipitation, we used
different As:Mg ratios (1:3, 1:10, 1:50, 1:200 or 1:400), pH (7, 8, 8.5, 9 or 9.5) and
different salts (Mg(OH)2, MgCl2, MgO, MgCO3 and MgSO4). The solution was
centrifuged at 4200 rpm for 15 min. and filtered with Whatman N. 42 filters. The As in
solution was determined using ICP-MS. In addition, standard reference materials from
the National Institute of Science and Technology (NIST 1547 – peach leaves,
Gaithersburg, MD) and appropriate reagent blanks, internal standards and spikes were
used as quality checks to ensure method accuracy and precision.
51
Statistical Analyses
All data are presented as the mean of three replicates with standard deviation.
Significant differences were determined by using one-way analysis of variance
(ANOVA) and treatment means were compared by Tukey's multiple range tests at p <
0.05 using software (R 3.2.2) (Team, 2005).
Results and Discussion
Water-Soluble As in PV Biomass
Arsenic and Mg concentrations in the biomass were 2,665 and 2,630 mg kg-1,
respectively (Table 3-1). Water-soluble As varied with extraction time (6.8 – 61%) with
no difference between shaking versus no shaking treatments (Table 3-2 and Figure 3-
1), showing a potential for secondary contamination if the biomass was not properly
handled. Arsenic in dried biomass consisted of ~99% AsV although AsIII is the primary
form in fresh PV biomass (Duan et al., 2005; Zhang et al., 2002). As plants senesce,
AsIII is oxidized to AsV (Chrobok et al., 2016; Tu et al., 2003). During extraction,
solution pH increased with time (Table 3-2). As pH increases, organic compounds
become more negatively charged, facilitating anion desorption (Carbonell-Barrachina et
al., 1999). Besides, solution pH after 20-h extraction (6.8) was close to the arsenic acid
pKa2 of 6.96, favoring formation of H2AsO4- and HAsO4
-2, thereby increasing As
extraction.
These results indicate a potential for secondary contamination if disposed in a
landfill. Microbial activities in mildly alkaline and anaerobic conditions can reduce AsV to
AsIII, further increasing its mobility (Kjeldsen et al., 2002). Addition of organic
amendment rich in sulfur compounds and amorphous Fe oxide can be used to reduce
As mobility (Carbonell-Barrachina et al., 1999; Dixit and Hering, 2003).
52
Arsenic Extraction from PV Biomass
Common extractants for As in plants include 2.1% HCl, 2.1% H3PO4, 1 M NaOH
and 50% ethanol (Alam et al., 2001; Zhao et al., 2015). To determine the most efficient
method, their extraction efficiency was tested (Figure 3-2). During first extraction, HCl
and H3PO4 were most effective, with ethanol being least effective. The recoveries by
HCl, H3PO4, NaOH and ethanol were 81, 78, 47 and 14%, respectively, with 5-32%
being recovered in the second extraction by HCl. After two extractions, ethanol had the
lowest recovery at 20%; however, it also showed the least As remaining in PV biomass.
Low As recovery was probably attributed to As precipitation with Mg or absorption onto
Mg(OH)2 (Park et al., 2010).
Acid extraction is efficient and widely used for As speciation in plants at low As
concentration (Narukawa and Chiba, 2010; Williams et al., 2005). However, difference
in tissue matrices and As concentration impact its efficiency (Heitkemper et al., 2001).
Extraction with HNO3 at 90°C recovered >90% of As from PV biomass (Zhao et al.,
2015). However, some acids including HCl may interfere in As determination using ICP-
MS (Cai et al., 2000).
Dilute H3PO4 is efficient in extracting As from organic compounds, sediments and
soils (Bohari et al., 2002; Giacomino et al., 2010; Tokunaga and Hakuta, 2002).
However, As extraction from plant material is variable with recovery ranging from 0 –
94% (Bohari et al., 2002; Foster et al., 2007; Kuehnelt et al., 2001). At 2.1%, H3PO4 was
efficient in extracting 78% As from PV biomass (Figure 3-2). However, it presented a
drawback when recovering As from the solution by competing for the sorbent with much
greater concentration than As. Dilute NaOH was also efficient in plant As extraction (He
et al., 2002). NaOH removes As by breaking S-As bonds and by hydroxyl ion ligand
53
replacement, with the high pH avoiding its readsorption (Bohari et al., 2002; Jang et al.,
2005). However, NaOH was ineffective on PV biomass, extracting only 47%. This was
probably due to As precipitation with Mg or absorption onto Mg(OH)2 at pH > 12 (Figure
3-2) (Park et al., 2010). Also, NaOH partially solubilized the PV biomass, making it
difficult to separate biomass from solution and, thus, it is impractical.
Ethanol is a powerful and non-toxic extractant for As speciation (Amaral et al.,
2013). Zhao et al. (2015) obtained satisfactory As recovery in PV biomass using ethanol
coupled with sonication when compared to other methods including methanol. In this
study, ethanol was least effective, with 20% recovery (Figure 3-2). Interestingly, it also
had the least residual As in the biomass, leaving 70% unknown. Upon reviewing the
procedure, a colloidal precipitate was noticed in the vial bottom. Thus, low recovery was
probably related to chlorophyllic Mg release into solution, resulting in formation of
Mg3(AsO4)2 at pH 7-10. Based on As removal at > 90% via spontaneous precipitation,
ethanol was the best extractant.
Optimization of Ethanol Extraction from PV Biomass
Based on a preliminary test, 35% ethanol was optimal after one extraction (data
not shown), which was different from Zhao et al. (2015) who reported 25% ethanol. The
use of sonication in their study compared to higher temperature in our study may
explain the difference. Other factors include time, particle size and pH. Zhao et al.
(2015) reported >90% As recovery after 0.5 h compared to 72% recovery in this study
(Figure 3-3a). The difference might be attributed to differences in As speciation in PV
biomass. In Zhao et al. (2015), As was 93% AsIII in fresh PV biomass compared to 99%
as AsV in dried PV biomass in this study. Compared to AsIII, AsV is more accumulated
in cell walls, thus requiring more time to extract (Yuan et al., 2005; Zhao et al., 2015). In
54
fact, after increasing extraction time to 2 h, no difference in As recovery compared to
Zhao et al. (2015) was noticed (Figure 3-3a). Therefore, 2 h was chosen as extract time.
In terms of particle size, there was a difference with smaller particles resulting in greater
recovery (Figure 3-3b). This result was expected based on the larger surface area of
smaller particles, which enhanced As extraction.
Similar to particle size, solid-to-liquid ratio affected As extraction with 1:50 ratio
being most effective at 90-100% recovery (Figure 3-3c). Zhao et al. (2015) used 1:300
S:L ratio, which recovered <90% As in this study. Given the large amount of biomass,
lower S:L is better. Though it was expected that higher pH would result in higher As
recovery, this was not the case (Figure 3-3d). Hydroxyl ions can replace AsV at high
pH, thereby increasing its extraction (Jang et al., 2005). It was possible that formation of
Mg3(AsO4)2 or absorption onto Mg(OH)2 might explain the difference. Based on our
data, the optimal As removal procedure (2 h, particle size < 1mm, S:L of 1:50 and pH 6)
produced satisfactory results, removing ~90% As from PV biomass.
Precipitation of Water-Soluble As from PV Biomass
After extraction, recovered soluble As needs to be treated. Arsenic concentration
in the ethanol extraction solution was 28.7 mg L-1 (Table 3-3). Technologies used to
remove As from effluents include adsorption, precipitation with Fe oxides, and
electrocoagulation (Bissen and Frimmel, 2003b; Sullivan et al., 2003). In addition, AsV
can be precipitated with Mg, which presents low solubility (Magalhaes, 2002).
Hoernesite [Mg3(AsO4)2], which has Ksp of 10-30.32 and forms in pH range of 7-10,
has been widely studied (Raposo et al., 2004; Zhu et al., 2005). However, its application
in As removal is limited. Under alkaline conditions at stoichiometric As:Mg ratio > 0.5,
precipitation of Mg3(AsO4)2 may occur (Park et al., 2010). In fact, spontaneous
55
precipitation was observed at effluent pH > 7 and As:Mg ratio of 1:1.9 in our study,
decreasing As concentration. Thus, this process was optimized using different Mg salts.
There was no difference among Mg salts even though they have different
solubility constants (5.61x10-12, 2.37x10-8 and 6.82x10-6 for Mg(OH)2, MgO and MgCO3,
respectively; (Haynes, 2014) (Figure 3-4a). One advantage of using Mg(OH)2 or MgCO3
was the potential to increase pH by 0.7 and 1.9 units, respectively. However, their low
solubility is a disadvantage for this process. Therefore, MgCl2 was chosen due to its
high solubility (ksp 738) and low cost.
Arsenic concentration was reduced from 28.7 mg L-1 to < 2 mg L-1 at pH 9 with
As:Mg ratio of 1:400 (Figure 3-4b). Speciation modeling predicted As concentration of
0.8 mg L-1 at pH 9.5 and As:Mg ratio of 1:3 (Park et al., 2010). In this study, As:Mg ratio
at 1:3 decreased As concentration by ~35%. This difference might be explained by As
competition with phosphate ions (HPO4-2) and the presence of organic ligands.
Phosphate and arsenate are analogues so phosphate presence decreases the
effectiveness of AsV removal. Besides, precipitation of Mg3(PO4)2 may occur at pH 6-10
and P:Mg ratio of 3:2 (Tamimi et al., 2011).
Among all factors, pH was the master variable in As removal, with the optimal pH
being 9.5 (Figure 3-4c), which is similar to Park et al. (2010). As pH increased, soluble
As concentration decreased from 12 mg L-1 to 0.4 mg L-1. Besides, at pH > 9.5,
Mg(OH)2 can also precipitate (Tabelin et al., 2013), which has positive charge to sorb
oxyanions such as AsV. However, at pH > 11, As was again soluble due to higher
stability of Mg(OH)2 (Park et al., 2010).
56
Addition of MgCl2 was efficient to remove As from ethanol extraction solution
from PV biomass. The As-Mg precipitated can be reused or sent for waste disposal, but
in much smaller quantity. Further, after removing As precipitate, pH can be increased to
> 11 to precipitate Mg as Mg(OH)2, allowing it to be reused in As removal after acid
dissolution.
57
Table 3-1. Elemental concentrations in six month old PV frond biomass obtained from a long-term phytoremediation experiment of As–contaminated soil (n=3).
Element Concentration
pH (0.01 M CaCl2) 5.14
------ mg kg-1 ------
As 2665 ± 31
Cu 4.5 ± 0.2
Mn 67 ± 2.0
Zn 53 ± 2.0
Fe 69 ± 1.2
P 1193 ± 57
Ca 2235 ± 113
Mg 2630 ± 100
K 12016 ± 827
Table 3-2. Water-soluble arsenic (%) and solution pH as a function of time and extraction method from PV frond biomass (static or shaking at 60 rpm in a rotary shaker) (n=3).
Time (h) No shaking pH Shaking pH
2 6.8 ± 1.6 B 4.9 ± 0.5 7.6 ± 0.5 b 5.7 ± 0.2
4 6.1 ± 0.7 B 5.9 ± 0.1 6.1 ± 0.2 b 6.3 ± 0.0
8 5.3 ± 0.5 B 6.4 ± 0.0 7.6 ± 0.9 b 6.7 ± 0.2
12 7.4 ± 1.1 B 6.3 ± 0.3 7.2 ± 0.4 b 6.0 ± 0.1
16 7.4 ± 3.8 B 6.1 ± 0.6 6.6 ± 0.3 b 6.3 ± 0.3
20 60.8 ± 1.2 A 6.8 ± 0.6 57.4 ± 1.8 a 6.8 ± 0.1
Treatments followed by the same letters are not significantly different at p < 0.05
58
Table 3-3. Characterization of optimized ethanol extraction effluent (n=3).
pH As Mg P Ca Fe As:Mg
molar ratio
------------------------------------------ mg L-1 ----------------------------------- 5.51 ± 0.02 28.7 ± 3.7 18.0 ± 0.27 9.46 ± 1.4 8.24 ± 0.14 0.04 ± 0.01 1:1.9
Figure 3-1. Water-soluble arsenic from PV biomass as a function of time and extraction method (static or shaking at 60 rpm in a rotary shaker) at 1:25 solid:liquid ratio (n=3) (Data presented in Table 3-2). Treatments followed by the same letters are not significantly different at p < 0.05.
Time
2 4 8 12 16 20
Ars
enic
in
solu
tion
(%
)
0
5
10
15
30
40
50
60 no Shaking Shaking
B
BB
B
B
b
bb
bb
Aa
59
Figure 3-2. Arsenic extraction from PV frond biomass using different extractants
followed by HCl extraction (recovery of 123, 116, 82 and 20% for 2.1% HCl, 2.1 % H3PO4, 1 M NaOH and 50% ethanol, respectively) (n=3).
HCl
81%
32%
10%
78%
27%
11%
H3PO
4
14%
5%
1%
70%
Ethanol
1st extraction
2nd
extraction - 2.1% HCl
Remaining in biomass
Not recovered
47%
27%
8%
29%
NaOH
47%
27%
8%
29%
NaOH
60
Figure 3-3. Effect of time (a), particle size (b), solid-to-liquid ratio (c) and pH (d) on As extraction from P. vittata biomass using 35% ethanol. Time was tested using particle size < 1 mm, S:L ratio 1:25 and pH 6. Particle size was tested using 2 h extraction, S:L ratio 1:25 and pH 6. Solid-to-liquid ratio was tested using 2 h extraction, particle size < 1 mm and pH 6. pH was tested using 2 h extraction, particle size < 1 mm and S:L ratio 1:50. (n=3)
61
Figure 3-4. Effect of Mg salts (a), As:Mg ratio (b) and pH (c) in Mg–As precipitation (n=3). Effect of Mg salts was tested using As:Mg ratio of 1:50 and pH 9. As:Mg ratio was tested using MgCl2 and pH 9. Optimal pH was tested using MgCl2 and As:Mg ratio of 1:400.
62
CHAPTER 4 ARSENIC REMOVAL FROM AS-RICH BIOMASS OF AS-HYPERACCUMULATOR
PTERIS VITTATA: COUPLING ETHANOL EXTRACTION WITH ANAEROBIC DIGESTION
Anaerobic Digestion
Arsenic is a metalloid and often presents in low concentrations in soils, ranging
from 0.1-67 mg kg-1 (Kabata-Pendias, 2011; Mandal and Suzuki, 2002). However,
human activities have increased its concentrations in the environment (da Silva et al.,
2018b). Arsenate (AsV) and Arsenite (AsIII) are the main As forms in soils, with AsV
predominating in oxic conditions while AsIII predominates in anoxic environments (Bohn
et al., 2002).
Arsenic is toxic and chronic exposure to As might cause cancers (ATSDR, 2007;
Gress et al., 2015, 2014). Arsenic availability in soils depends on redox potential, pH,
and content of Fe/Al (hydr)oxides (Alves et al., 2016; Walthan and Eick, 2004). Arsenic
exposure is mainly related to consumption of contaminated food and water, with
inorganic As being more toxic than organic As (ATSDR, 2017; Gress et al., 2016).
Moreover, plants exposure to arsenic can reduce their growth. However, some plants
have developed mechanisms to cope with As, either reducing its uptake or
accumulating it in the roots (El-Zohri et al., 2015; Lessl et al., 2014; Wang et al., 2015).
Unlike typical plants, hyperaccumulator plants can accumulate high concentrations of
As. For example, As-hyperaccumulator Pteris vittata (Chinese brake fern) can
accumulate up to 23,000 mg kg-1 As in the fronds (Ma et al., 2001; Tu and Ma, 2002).
In the US, there are more than 600 sites contaminated with As that require
remediation (USEPA, 2017). However, conventional remediation technologies are
expensive and time-consuming (Belluck et al., 2003). Phytoremediation is a plant-based
63
technology that uses hyperaccumulator plants to remove metals from a soil. The
contaminants are accumulated in the shoots, which can be collected, transported and
disposed off-site (Rahman and Hasegawa, 2011). Once accumulated in plant biomass,
proper disposal of contaminant-rich biomass is important (Krämer, 2005).
The USEPA considers material with As concentration > 100 mg kg-1 As to be
unsafe (USEPA, 2002). Thus, biomass disposal can be a limitation for
phytoremediation. Conventional methods for disposing of contaminated-biomass often
use hazardous-waste landfills or incineration (Chaney et al., 2007). In Florida alone, it
was estimated that, by 2000, 12.8 kiloton of As had been disposed in unlined landfills as
As-treated wood (Saxe et al., 2007). Moreover, a batch study showed that, under
anaerobic conditions, As-treated wood leached 11% of the As, with solution
concentration being 4,000 times greater than the maximum contaminant level for
groundwater (MCL = 10 µg L-1) (Jambeck et al., 2006). Other biomass disposal
technologies include compaction, pyrolysis, ashing, and liquid extraction. Recently, an
alternative method to treat As-rich biomass using ethanol extraction followed by Mg
precipitation was developed (da Silva et al., 2018a). The method removed ~90% As
from P. vittata biomass, and As concentration in the solution was decreased from 28.7
mg L-1 to ~0.4 mg L-1 by As-Mg precipitation. However, the 10% As remaining in the
biomass might still be an issue.
Microbe-mediated composting has potential to reduce biomass and release As,
thereby reducing transportation and disposal cost. Composting can stabilize metals in
the biomass, reducing their availability (Singh and Kalamdhad, 2013). However, arsenic
leaching and volatilization are concerns during composting as As can be methylated
64
during the process (Cao et al., 2003; Balasoiu et al., 2001). Arsenic loss was reported
during aerobic composting, which was attributed to losses of volatile As (Cao et al.,
2010). However, anaerobic digestion of As sludge showed that < 1% of total As was
volatile (Cortinas et al., 2006). Anaerobic digestion is a complex system in which
symbiotic microbes transform organic materials into biogas in an anoxic environment,
leaving refractory organic matter (Wilkie, 2008). Anaerobic digestion has potential to
reduce biomass and to extract As from PV biomass. Thus, coupling ethanol extraction
(da Silva et al., 2018a) with anaerobic digestion may be a powerful method to treat As-
rich biomass. However, it is necessary to assess As transformation during the process.
The goal of this study was to remove residual As from As-rich P. vittata biomass
by coupling ethanol extraction with anaerobic digestion. The specific objectives were to:
1) examine substrate utilization kinetics of P. vittata biomass during anaerobic digestion;
2) Assess As partitioning during anaerobic digestion. Removing most of the As from P.
vittata biomass to avoid As loss helps to improve its application in phytoremediation.
Material and Methods
Chemical Reagents and P. vittata Biomass
All chemicals were of analytical grade. Nitric acid (trace metal grade) and H2O2
were obtained from Fisher Scientific (Waltham, MA). Sep-Pak AccellPlus QMA Plus
Short cartridges for As speciation were obtained from Waters Corporation (Milford, MA).
Before use, all labware was washed and soaked in 1 M nitric acid for 24 h and rinsed
several times with DI water.
P. vittata (PV) biomass with no As was obtained from a non-contaminated site in
Gainesville, FL. Arsenic-rich PV biomass was obtained from a long-term
phytoremediation experiment with 126 mg kg-1 As in the soil (Lessl et al., 2014). PV
65
fronds were harvested, oven dried at 65°C and shredded to < 1 mm size.
Concentrations of As and other elements in PV fronds are shown in Table 1. Since only
frond biomass was harvested for As-removal, all biomass in this study was from P.
vittata fronds.
Methane Index Potential from P. vittata Frond Biomass
Anaerobic digestion has potential to reduce biomass and to extract As from PV
biomass, and coupling it with ethanol extraction might enhance As extraction. Thus, PV
biomass was first extracted with ethanol solution to remove As from biomass (da Silva
et al., 2018a). Treatments were: ethanol-treated control-biomass (TCB – without As)
and ethanol-treated As-rich biomass (TAsB). For ethanol extraction, in a falcon vial, 1 g
of PV biomass was added followed by 35% ethanol solution and kept in a water bath at
80 °C for 2 h, using particle size < 1 mm, and with S:L ratio 1:50 at pH 6. After that,
samples were centrifuged at 4200 rpm for 15 min and filtrated with Whatman N. 42
filters. Arsenic concentrations in the treated biomass and in solution were analyzed as
described below.
Methane index potential (MIP) assays were used to assess methane yields from
PV biomass. PV biomass was added to 250-mL anaerobic serum bottles at biomass
loading of 2.0 g volatile solids per liter (gVS L-1). The inoculum was obtained from a
mesophilic anaerobic digester fed with waste food and characterized similarly as PV
biomass (Fallahi et al., 2016). Total solids (TS) was determined by drying the samples
for 24 h in an oven at 105°C. Volatile solids (VS) was determined by igniting the sample
in a muffle furnace at 550°C (APHA, 1995). Total solids and VS were determined before
and after anaerobic digestion. The inoculum had a pH of 7.62 ± 0.01, 0.66% TS, 28.9%
VS and As concentration < 0.01 µg L-1. A total of 200 mL of inoculum and 0.43 and 0.44
66
g of PV biomass (for TAsB and TCB, respectively) was added. Inoculum blanks were
included as a negative control and positive controls were glucose, cellulose and starch.
P. vittata biomass, positive and negative controls were assayed in triplicate.
Bottles were sealed with a rubber septum, crimped with an aluminum cap,
inverted to prevent potential gas leakage, and incubated at 35°C for 35 d. A liquid-
displacement method with 3M KOH as barrier solution was used to measure methane
gas production (Wilkie et al., 2004). During peak methane production, gas
measurements were recorded daily and later at longer intervals based on gas
production. The methane volumes were corrected by subtracting the mean methane
volume of the inoculum controls and normalized to standard temperature and pressure
(0°C and 760 mm Hg). The methane yields are reported as normalized liter (i.e., LN) of
methane produced per kg of VS added. After anaerobic digestion, biomass and VS
mass balances were performed. Biomass mass balance was obtained by subtracting
final biomass and inoculum TS from initial biomass TS. Volatile solids mass balance
was obtained by subtracting final biomass and inoculum VS from initial biomass VS.
Arsenic Analysis in plant biomass and mass balance
Arsenic concentration in PV biomass was determined by HNO3/H2O2 digestion
on a hot block (Environmental Express, Ventura, CA) using USEPA Method 3050B.
Briefly, 0.5 g of dried plant biomass was suspended in 15 mL 1:1 nitric acid and heated
at 105 °C for 8 h. After cooling, 1 mL 30% H2O2 was added and the sample was
digested for an additional 30 min before bringing samples to a 50 mL volume with DI
water. Arsenic concentration was analyzed using inductively coupled plasma mass
spectrometry (ICP-MS, Perkin-Elmer Corp., Norwalk, CT).
67
After anaerobic digestion, As mass balance was calculated. Samples were
centrifuged at 4,200 rpm for 15 min and filtered with Whatman N. 42 filters. Filtered
supernatant was the liquid phase while solid material was the solid phase. Undisturbed
centrifuged solid phase was oven dried for 24 h at 60 °C and fractionated into two
components: PV biomass and other solids. Then, a homogeneous sub-sample of both
liquid and solid phases was digested as described above and As concentration was
determined by ICP-MS. Arsenic in the gas phase was determined by trapping the gas in
20 mL of 2 M nitric acid and analyzing for total As using ICP-MS.
Precipitation of Water-Soluble As from PV Biomass
After anaerobic digestion, water-soluble As was precipitated from the anaerobic
digestate supernatant (da Silva et al., 2018a). Briefly, removal of soluble As was
achieved by adding MgCl2 at As:Mg ratio of 1:400 and pH 9.5. Then, the solution was
centrifuged at 4200 rpm for 15 min. and filtered with Whatman N. 42 filters. Arsenic in
solution was determined using ICP-MS. In addition, standard reference materials from
the National Institute of Science and Technology (NIST 1547 – peach leaves,
Gaithersburg, MD) and appropriate reagent blanks, internal standards and spikes were
used as quality checks to ensure method accuracy and precision.
Statistical Analyses
All data are presented as the mean of three replicates with standard deviation.
Significant differences were determined by using one-way analysis of variance
(ANOVA) and treatment means were compared by Tukey's multiple range test at p <
0.05 using software (R 3.2.2) (Team, 2005).
68
Results and Discussion
As expected, ethanol extracted soluble metals (Table 4-1), with As and K
presenting the highest removed amounts from As-rich PV biomass, decreasing 93%
(2,665 to 197 mg kg-1) and 74% (25,916 to 6,860 mg kg-1), respectively. Ethanol also
extracted soluble carbon from PV biomass. Total solids was reduced by 2% while
volatile solids increased by 3% (Table 4-1).
Methane Production from P. vittata Biomass
Typical of plants, P. vittata biomass had high organic C, with VS of 91.1 and
94.1%TS for TCB and TAsB, respectively (Table 1). The data from positive controls
assays showed that the inoculum presented satisfactory methanogenic activity to
convert glucose, cellulose and starch.
Ethanol-extracted PV biomass methane yield was 145 and 160 LNCH4/kgVS after
35 d for TAsB and TCB, respectively (Table 4-2 and Figure 4-1). Compared to methane
feedstocks such as grass, perennial grass and maize, PV biomass had lower CH4 yield
(Monlau et al., 2012; Thomsen et al., 2014; Triolo et al., 2011). Methane yield of
perennial grass was 410 LNCH4/kgVS while wild grass yielded 306 LNCH4/kgVS (Triolo
et al., 2011). However, compared to other feedstocks, PV biomass methane yield was
similar. For example, sorghum and rice straw CH4 yields were 210 and 195 LNCH4/kgVS
(Dinuccio et al., 2010; Monlau et al., 2012). In this study, extracted control PV biomass
had a higher CH4 yield compared to extracted As-rich PV biomass (Table 4-2 and
Figure 4-1). Methane production kinetics in PV biomass were similar compared to other
feedstocks. For example, maize and wild grass yielded ~72 and 62% (290 and 190
LNCH4/kgVS) after 14 d, respectively (Triolo et al., 2011), while methane production in
TCB reached 68% (110 ± 0.86 LNCH4/kgVS) and in TAsB reached 73% (107 ± 0.86
69
LNCH4/kgVS) in the same timeframe. Differences in methane yield can be related to
microbial accessibility limitation to lignocellulosic substrates during the fermentative
process, mostly due to biomass compositional and structural characteristics (Monlau et
al., 2012; Triolo et al., 2011). Though methane production was not the focus of this
study, it does add value to anaerobic digestion as a method to recover As and avoid
landfill disposal of As-rich biomass.
Arsenic Partitioning After Anaerobic Digestion
Arsenic in PV biomass contained mostly AsV, at 99%, though AsIII is the major
species in fresh biomass (Duan et al., 2005; Zhang et al., 2002). Many extractants have
been used to extract As from plant biomass including 2.1% HCl, 2.1% H3PO4, 1 M
NaOH and 35-50% ethanol (Alam et al., 2001; da Silva et al., 2018a; Zhao et al., 2015).
While AsIII is often stored in plant vacuoles, more AsV is accumulated in cell walls, and
thus it is more recalcitrant (Yuan et al., 2005). For PV biomass, 35% ethanol extracted
~90% of total As (da Silva et al., 2018a). However, the 10% As remaining in the
biomass might be an issue. Therefore, it is important to understand arsenic partitioning
and biomass degradation during anaerobic digestion.
After 35 d of anaerobic digestion, both VS and PV biomass decreased
significantly (Figure 4-2) and most As was in the liquid phase (Figure 4-3). Volatile
solids decreased from 91.1 and 94.1 to 14.9 and 17.5 %TS, while PV biomass
decreased by 71.4 and 65.1% for TCB and TAsB, respectively (Figure 4-2). Moreover,
As concentration did not affect PV biomass degradation, as the remaining biomass was
similar (28.6 and 34.9%, for TCB and TAsB, respectively). However, anaerobic
digestion decreased As concentration in the ethanol-extracted PV biomass (Figure 4-3).
Arsenic in the digested biomass was 11.7% with As concentrations of 60 mg kg-1
70
(Figure 4-3B). Therefore, anaerobic digestion of PV biomass after ethanol extraction
has an advantage as As concentration in the biomass decreased from 2,665 to 60 mg
kg-1, or by ~98%. At this level, PV biomass would be considered a safe material since
As concentration is lower than 100 mg kg-1 (USEPA, 2002). Consequently, PV biomass
would not require hazardous-waste landfill disposal, reducing cost and eliminating a
secondary contamination risk (loss by volatilization and/or in leachate).
Anaerobic digestion was more efficient than aerobic composting in biomass
reduction, As removal and As recovery, i.e., no As loss by volatilization. Arsenic loss by
volatilization was reported at 18% in aerobic composting and ~100% via incineration
(Cao et al., 2010; Yan et al., 2008). Aerobic composting of As-rich PV biomass reduced
biomass and As by 38% and 25%, respectively (Cao et al., 2010). In this study, biomass
reduction was 64.4% and As removal from biomass was up to 98% when anaerobic
digestion was coupled with ethanol extraction. Therefore, based on our data, anaerobic
digestion of PV biomass after ethanol extraction produced satisfactory results, allowing
the PV material to be classified as safe (USEPA, 2002).
Precipitation of Water-Soluble As from PV Biomass
After anaerobic digestion, water-soluble As in the digestate also needs to be
treated. Arsenic concentration in the solution was 0.53 mg L-1 for TAsB (Figure 4-4).
Technologies for arsenic removal from solutions include precipitation with Fe oxides,
adsorption and electrocoagulation (Bissen and Frimmel, 2003b; Sullivan et al., 2003). In
addition, AsV can be precipitated with Mg (da Silva et al., 2018a).
Among the factors that influence As-Mg precipitation, pH and As:Mg molar ratio
are the most important. Spontaneous precipitation may occur in pH 7-10, though in pH <
9.5 Mg(OH)2 starts to precipitate, affecting efficiency (Tabelin et al., 2013). Addition of
71
MgCl2 at 1:400 As:Mg molar ratio and increasing pH to 9.5 reduced As concentration in
solution to 0.26 mg L-1 or by 51% in the TAsB treatment (Figure 4-4). The precipitate
can be reused or sent for waste disposal in much smaller amounts compared to the
initial biomass. Besides, after removing As precipitate, Mg can be precipitated from
solution as Mg(OH)2 by increasing pH > 11, allowing it to be reused in As removal after
acid dissolution (da Silva et al., 2018a).
72
Table 4-1. Characterization of non-extracted and ethanol-extracted P. vittata biomass obtained from a long-term phytoremediation experiment of As–contaminated soil. ± standard deviation (n=3).
As biomass before ethanol extraction
As biomass after ethanol extraction
Control biomass before ethanol
extraction
pH (0.01 M CaCl2) 5.14 5.51 5.61
Total solids (%) 93.1 ± 0.3 91.2 ± 0.6 92.6 ± 0.1
Volatile Solids (%TS) 90.6 ± 0.1 94.1 ± 0.3 91.1 ± 0.1
Total COD (g/kg) 877 ± 121 1076 ± 198 1020 ± 151
------------------------------- mg kg-1 ---------------------------
As 2,665 ± 31 197 ± 2.79 3.10 ± 0.03
Cu 9.9 ± 0.2 6.6 ± 0.5 8.1 ± 1.2
Zn 63.1 ± 2.1 51.7 ± 6.1 51 ± 3.5
Ca 7,710 ± 180 4,690 ± 160 3,500 ± 250
Mg 6,890 ± 420 3,270 ± 160 4,410 ± 198
K 25,916 ± 1,510 6,860 ± 430 12,815 ± 751
73
Table 4-2. Cumulative methane (CH4) yield (LNCH4/kgVS) for ethanol-extracted P. vittata biomass with and without arsenic. Treatments are: treated control-biomass (TCB –No As) and treated As-rich biomass (TAsB) (n=3).
Days TCB TAsB Difference in CH4 yield
(%)
0 0.00 ± 0.00 0.00 ± 0.00 0.00
1 1.9 ± 0.00 1.9 ± 2.25 0.00
2 9.6 ± 0.33 8.6 ± 0.33 10.87
3 17.6 ± 1.42 18.8 ± 2.03 -6.00
5 47.5 ± 5.37 46.0 ± 1.72 3.27
7 73.0 ± 3.96 68.7 ± 1.81 6.28
9 87.7 ± 1.49 82.2 ± 1.13 6.62
11 97.4 ± 1.72 94.0 ± 0.56 3.59
14 110 ± 0.86 107 ± 0.86 2.82
18 123 ± 1.17 119 ± 0.86 2.99
22 133 ± 0.98 131 ± 0.33 1.14
26 140 ± 0.98 141 ± 0.65 -0.80
30 149 ± 0.86 141 ± 0.33 5.46
35 160 ± 0.00 145 ± 0.33 9.82
74
Figure 4-1. Cumulative methane yield (LNCH4/kgVS) for P. vittata control and As
biomass with ethanol extraction. Treatments are: ethanol-treated control biomass (TCB – No As) and ethanol-treated As-rich biomass (TAsB). Bars represent standard deviation (n=3) (Data presented in table 4-2).
Days
0 5 10 15 20 25 30 35
Cu
mu
lati
ve
Me
tha
ne
Yie
ld (
LNC
H4/k
gV
S)
0
40
80
120
160
200
TCB TAsB
75
Figure 4-2. Initial and final volatile solids (%TS) (A) and remaining P. vittata biomass (B)
after 35 d of anaerobic digestion. Treatments are: ethanol-treated control biomass (TCB – No As) and ethanol-treated As-rich biomass (TAsB). Treatments followed by the same letters are not significantly different at p < 0.05. Bars represent standard deviation (n=3).
76
Figure 4-3. Ethanol-extracted PV biomass arsenic partitioning among gas, solid and
liquid phase (A) and As solid phase fractionation (PV biomass + effluent) (B) after 35 d of anaerobic digestion. Phases followed by the same letters are not significantly different at p < 0.05. Bars represent standard deviation (n=3).
Other solids
77
Figure 4-4. Solution arsenic removal as As-Mg precipitate using MgCl2, As:Mg ratio of
1:400 and pH 9.5. Bars represent standard deviation (n=3).
Initial Final
Ars
en
ic i
n s
olu
tio
n (
mg
L-1
)
0.0
0.2
0.3
0.5
0.6
0.8
78
CHAPTER 5 CONCLUSIONS
In our study, P. vittata was efficient in removing As from two contaminated soils
with the greatest reduction occurring in the amorphous and crystalline fractions.
However, as more As was taken up by P. vittata, less As was removed from soils due to
lower soil As concentration. Though, initially, PR-treatment was more effective in
removing As from soils than P-treatment, with time, As depletion was similar in two
treatments. Still the fact that P. vittata grew well in CCA-soil near pH 7 for 5 years with
only PR as its P supply indicated its ability to acquire non-available P to sustain
biomass production, which is rather unique among plants.
An alternative method to remove As from As-laden biomass was assessed using
35% ethanol extraction followed by As precipitation with Mg. Optimal ethanol extraction
was achieved using 2 h, particle size < 1 mm, S:L ratio 1:50 at pH 6. After extraction, As
removal from the solution was achieved using MgCl2 at As:Mg ratio of 1:400 and pH
9.5. Approximately 60% As from dried frond biomass of P. vittata was water soluble,
with 99% As being AsV. Ethanol was effective in As removal (> 90%) followed by
spontaneous precipitation. MgCl2 addition in the effluent decreased As concentration
from 28.7 mg L-1 to ~0.4 mg L-1.
However, for large-scale application the remaining 10% As in biomass still can
pose risk. Thus, an alternative method to treat As-rich biomass was developed by
coupling ethanol extraction with anaerobic digestion, allowing PV biomass to be
classified as a safe material. Methane yield, a by-product of anaerobic digestion, was
145 and 160 LNCH4/kgVS after 35 d for TAsB and TCB, respectively. On the other hand,
VS in PV biomass decreased from 91.1 and 94.1 to 14.9 and 17.5%TS, while PV
79
biomass decreased by 71.4 and 65.1% for TCB and TAsB, respectively. Anaerobic
digestion of PV biomass after ethanol extraction decreased As concentration from 2,665
to 60 mg kg-1, or by ~98%, in the biomass. At this level, PV biomass would be
considered a safe material by USEPA regulations (< 100 mg kg-1). Finally, 51% of As in
anaerobic digestate was recovered by As-Mg precipitation, decreasing the As
concentration from 0.53 to 0.26 mg L-1.
Ethanol extraction followed by anaerobic digestion and As–Mg precipitation of
digestate was efficient to recover As from As-rich PV biomass. The process allows safe
disposal of PV biomass with the added benefit of biogas production potential. In
addition, As–Mg precipitate can be reused in different products such as wood
preservatives.
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APPENDIX METAL LEACHABILITY FROM COAL COMBUSTION RESIDUALS UNDER
DIFFERENT PHS AND LIQUID/SOLID RATIOS
Coal Combustion Residuals
Coal combustion residuals (CCRs) are one of the largest waste streams
generated in the USA (Riazi and Gupta, 2015). They are byproducts of coal combustion
during electricity generation and consist of fly ash, bottom ash and flue gas
desulfurization (FGD) residue (USEPA, 2009a). In 2013, over 115 million tons of CCRs
were produced in the USA, with ~51.4 tons being beneficially used for mine waste
treatment, cement and concrete mixture, fill materials and metal recovery (ACCA, 2016;
Clarke, 1993). Due to updated Clean Air Act (USEPA, 2004), CCRs production reached
140 million tons in 2015 (Riazi and Gupta, 2015; USEPA, 2004).
Fly ash is fine in size (0.5-100 µm), which is generated during coal combustion
and carried up with flue gas. To minimize its release into the atmosphere, emission
control devices are used to separate fly ash in the flue gas from the air stream (Schnelle
et al., 2002). Bottom ash is the remaining residue in the boiler after fly ash removal and
is formed in dry-bottom boilers and stokers (USEPA, 1996). FGD residues are produced
by air-emission control devices, which contain lime-based materials to trap SO2 as CaS
or CaSO4 to reduce its atmospheric concentrations (Kosson et al., 2009). In some
cases, a selective catalytic reduction process is used to reduce NOx gas emissions by
spraying ammonia into the flue gas (Saracco et al., 1996).
Reprinted with permission from da Silva, E.B., Li, S., de Oliveira, L.M., Gress, J., Dong, X., Wilkie, A.C., Townsend, T., Ma, L.Q., 2018. Metal leachability from coal combustion residuals under different pHs and liquid/solid ratios. J. Hazard. Mater. 341, 66–74. doi:10.1016/j.jhazmat.2017.07.010
81
CCRs contain variable amounts of soluble salts and trace metals that may leach
out and cause environmental problems (Ayoko et al., 2005; Riazi and Gupta, 2015).
Depending on the combustion process, CCRs can be enriched with trace metals such
as As, Cu, Pb, Se and Zn (Kosson et al., 2009; Ruhl et al., 2012, 2010). Coal source
impacts trace metal concentrations in CCRs (Nawaz, 2013). For example, Appalachian
and Illinois Basin coals have higher Pb, Cr, Ni, and As concentrations than coals from
the Rocky Mountains and Northern Plains, while Gulf Cost coals show the highest Hg
concentration (Riazi and Gupta, 2015; Yudovich and Ketris, 2005). Because different
coals contain different metal concentrations and pH values, metal leachability from
CCRs can vary widely.
Metal concentrations in CCRs are generally low, rarely reaching the hazardous
levels based on USEPA's Toxicity Characteristic Leaching Procedure (TCLP) (USEPA,
2014). However, if not managed properly, CCRs have the potential to cause
environmental contamination due to leaching of trace metals (Ruhl et al., 2012). For
example, trace metals including As, Se, B, Sr and Ba were detected in water from the
impacted area 18 months after the collapse of the Kingston facility (Ruhl et al., 2010).
The collapse of another facility in 2014 released ~39,000 tons of ash into the Dan River,
causing As, Se and Ba contamination in Duke river sediment (Lemly, 2015; USEPA,
2014). Therefore, it is important to assess the potential risk associated with CCRs to
minimize their impact on the environment.
The Synthetic Precipitation Leaching Procedure (SPLP) has been used to
determine metal leachability under controlled laboratory conditions (Jang et al., 2002).
However, it does not consider site-specific conditions or leaching behavior over different
82
time (Riazi and Gupta, 2015). Leaching Environmental Assessment Framework (LEAF)
methods were developed to examine metal leaching over a range of pHs (2 to 13;
USEPA Method 1313) and liquid/solid ratios (0.5 to 10 mL/g; USEPA Method 1316),
which helps to assess metal leaching behavior in landfills or surface impoundments
(USEPA, 2014). Combined with geochemical speciation modeling, LEAF methods can
estimate metal leachability under different environmental conditions and land disposal
characteristics (Kosson et al., 2009; Roessler et al., 2015). Even though tests with wild
range of pHs and L/S has been done in the past, they used different protocols and
some report are incomplete, thus limited information is available regarding metal
leachability of CCRs (Kosson et al., 2002; USEPA, 2009b).
The purpose of this study was to characterize and assess metal leachability in
representative CCRs samples from coal-fired power plants, including As, Ba, Cd, Cr,
Pb, and Se. The specific objectives were to: (1) measure fly ash metal leachability
under different pH conditions using LEAF Method 1313 (pH 2-13); (2) evaluate their
leachability under different liquid-solid ratios using LEAF Method 1316 (0.5-10 mL/g);
and (3) compare the LEAF results with those from SPLP.
Materials and Methods
Chemicals Reagents and CCR Samples
All chemicals were of analytical grade or better. Nitric acid (trace metal grade),
H2O2, HCl and KOH were obtained from Fisher Scientific (Waltham, MA). Before use, all
labware was washed and soaked in 1 M nitric acid for 24 h and rinsed several times
with DI water.
Twenty four (24) CCRs samples including 12 fly ash, 8 bottom ash, and 4 FGD
residues were obtained from 7 representative coal-power plants. Composite samples
83
were obtained following the in-stream and storage area methods. The in-stream method
allowed collection of CCRs samples from transitory storage areas or conveyance
systems. The storage area method allowed collection of CCRs samples from storage
areas such as storage piles. All CCRs samples were subjected to total and SPLP
analysis. Among those samples, eight fly ash samples were analyzed using both LEAF
methods (USEPA 1313 and 1316). All samples were derived from Bituminous type coal
(Eastern, Central Appalachian and Illinois basin coal).
SPLP and LEAF Tests
The CCRs water pH was measured after shaking at a solid/liquid ratio of 1:10
(w/v) for 1 h and passing through 0.45 µm membrane filters. CCR samples were
digested by USEPA Method 3050B for total concentrations (USEPA, 1996). Briefly, 1 g
of sample was suspended in 15 mL 1:1 nitric acid and heated at 105 °C for 6 h. After
cooling, 1 mL 30% H2O2 was added and digested for an additional 30 min before
bringing samples to a 50 mL volume with double DI water. Concentrations were
analyzed by inductively coupled plasma mass spectrometry (ICP-MS NexIon 300,
PerkinElmer Corp., Norwalk, CT).The SPLP was used to measure metal leaching
behavior in CCRs samples. The SPLP was shaken at 1:20 solid/extractant ratios for 20
h in a rotary shaker. The resulting solutions were filtered (0.45 µm), acidified with
concentrated HNO3, and pH was measured before and after acid addition. The total
content of As, Cd, Co, Cr, Cu, Ni, Pb and Se was analyzed by USEPA Method 6020
using ICP-MS (USEPA, 2007).
The LEAF Method 1313 uses dilute HNO3 or KOH to adjust pH in nine parallel
batch tests at a liquid/solid ratio of 10 mL/g, with the final pH values at 2, 4, 5.5, 7, 8, 9,
10.5, 12, or 13. The pH measurement was done within 30 minutes to avoid
84
neutralization of the solution. An extraction with no acid/base addition was done to
assess the pH of CCRs. The LEAF Method 1316 is a five parallel batch extraction as a
function of liquid/solid ratios. Water was added to 20 g of ash samples to reach 0.5, 1.0,
2.0, 5.0, or 10.0 mL/g. All samples were shaken for 24 h, followed by filtration and
analyzed using ICP-MS. For the LEAF method were chosen metals of most
environment concern (As, Ba, Cr, Cd, Pb and Se). All analyses were performed in
triplicate.
Quality Assurance
The QA/QC in SPLP and LEAF analyses included blanks, spikes and triplicates
for every 20 samples. Recovery was determined using spikes (80-120%) and relative
standard deviations of triplicate analysis were obtained. The performance of the ICP-
MS was checked by running an intermediate calibration standard for every 20 samples.
All the calibration standard checks were within the acceptable range (80-120%).
Results and Discussion
Total Metal Concentrations in CCRs
Trace metal concentrations in CCRs varied widely (Figure A-1 and Table A-1).
Nickel, Zn and Pb were the highest, with some CCRs exceeding the Florida residential
Soil Cleanup Target Levels (SCTLs) for Ni and Pb (340 and 400 mg kg-1 for Ni and Pb)
and one sample was close to Florida industrial SCTL for Pb at 1,400 mg kg-1. All As
concentrations and some Cr and Cu also exceeded the Florida residential SCTLs (2.1,
210 and 150 mg kg-1 for As, Cr and Cu, respectively). Among metals, As and Cu are
more toxic due to As’ carcinogenicity and Cu’s adverse impacts on aquatic biota
(ATSDR, 2007; Real et al., 2003). Compared to CCRs in the US and other countries
(Table A-2), some concentrations of Co, Ni, Se and Zn in the fly ash, and As, Ni and Se
85
in the bottom ash were higher. In fact, the highest Ni concentration was up to 27 times
greater than those in previous studies (Gallardo et al., 2015; McNally et al., 2012; Rowe
et al., 2002).
All metals were enriched in fly ash samples compared to bottom ash and FGD.
Among them, As, Cd and Pb have the highest vapor pressures and the lowest boiling
points, resulting in their higher accumulation in fly ash (Sukandar et al., 2006; Xiao et
al., 2015). On the other hand, Cu and Cr have relatively low vapor pressures and higher
boiling points so they are retained in slag or bottom ash. However, some samples
presented higher concentrations of Cu and Cr in fly ash than bottom ash (Davison et al.,
1974; Juda-Rezler and Kowalczyk, 2013), which might be related to the use of a
filtration device, such as an electrostatic precipitator, in pulverized coal boilers (Juda-
Rezler and Kowalczyk, 2013). Total concentrations for As, Ba, Cd, Cr, Hg, Pb and Se
for all CCRs can be found in Tables A3
Metal Concentrations in Fly Ash Based on SPLP
SPLP metal concentrations in FGD were the lowest while fly ash had the highest,
especially for As, Cr, Ni and Pb (Figure A-2). The final SPLP pH values were close to
the sample pH values (Table A-4), indicating a high buffering capacity of fly ash.
Buffering capacity of CCRs is correlated to their CaO concentration (Bin-Shafique et al.,
2006).
SPLP concentration of trace metals is related to its total concentration in CCRs
samples, with pH playing an important role. For example, Ni concentration is the lowest
at pH 8-10, while Cd concentration decreases at pH > 6 (Izquierdo and Querol, 2012;
Kabata-Pendias et al., 1984), which is consistent with our results where SPLP data
were the greatest in samples exhibiting pH < 6 (Figure A-2 and Table A-4). In this study,
86
Pb concentration was the highest in fly ash, being ~750 times the Florida Groundwater
Cleanup Target Levels (FGCTL) at 15 µg L-1 Pb. Besides, As concentration was also
high, being up to 5 times the FGCTL at 10 µg L-1. These results are different from
McNally et al. (McNally et al., 2012) who tested fly ash in Pennsylvania using SPLP,
and reported a much lower Pb concentration at 0.09 mg L-1 but similar As concentration
at 0.71 mg L-1. In case of FGD, even though total concentration was low compared to
bottom and fly ash, some SPLP concentrations of As, Cd, Ni and Pb exceeded FGCTL
(10, 5, 100 and 15 µg L-1) (Figure A-2). The highest Ni concentration was up to 4 times
of FGCTL of 100 µg L-1 Ni. Metal solubilization in FGD might be related to metal
containing carbonates dissolution by H2SO4 and HNO3 used in the SPLP (N H
Koralegedara et al., 2017).
As noticed in this study, pH controls metal solubility, therefore, it is important to
measure CCRs buffering capacity and final solution pH. For example, lower Cr
concentration is obtained at pH < 4 compared to pH 5-7 (Nathan et al., 1999). Besides,
Ni solubility is controlled by pH (Roy and Stegemann, 2017). The disadvantage of SPLP
is that it simulates the effect of acid rain on metal leaching in a single test condition
(Kosson et al., 2014), which is insufficient to describe metal leaching behavior under
different environmental scenarios and over time (Kosson et al., 2002; Thorneloe et al.,
2010). Concentrations of As, Ba, Cd, Cr, Hg, Pb and Se for all CCRs based on SPLP
can be found in Table A-5.
Metal Concentrations in Fly Ash Based on USEPA LEAF Method 1313
The parallel batch test under different pHs (USEPA Method 1313) can be used to
determine metal leachability over a broad range of environmental conditions. Figure A-3
shows the results for trace metals that showed significant leaching (As, Ba, Cd, Cr, Pb
87
and Se) in fly ash. Concentrations of As, Ba, Cd, Cr, Hg, Pb and Se based on USEPA
Method 1313 for all CCRs can be found in Table A-6. All metals excluding Cd showed
amphoteric behavior, presenting higher solubility at low and high pH values (Figure A-
3). Cadmium concentration decreased at pH > 7, with the highest being at pH < 4. pH is
a master variable that can affect element speciation, therefore solubility (Bohn et al.,
2002). In addition, low pH affects Ca-containing compounds, causing disturbance in the
ash matrix and releasing associated metals (Fedje et al., 2010).
Besides pH, metal leaching from CCRs is also influenced by solid to liquid ratio
and other chemicals in the CCRs matrix (Tang and Steenari, 2016). For example, Cu
leaching is affected by the ash mineralogy (presence of crystalline or amorphous
minerals) and presence of organic compounds (Fedje et al., 2010). Moreover, Cd, Cu
and Pb sorption is usually low at pH < 4 (Gerçel and Gerçel, 2007). Tang and Steenari
(Tang and Steenari, 2016) also noticed that metals present as chlorides are released in
high concentrations from CCRs.
As and Se Concentrations
Arsenic and Se showed typical oxyanionic behavior with increasing
concentrations as pH decreased (Figure A-3). Arsenic concentrations (0.70 – 6497 µg L-
1, 25th percentile = 11.6) in most fly ash samples were higher than the FGCTL at 10 µg
L-1. Nevertheless, at typical environmental pHs (4 – 8), As concentration was the lowest
(0.70 – 204 µg L-1, 25th percentile = 7.8), with more than 30% being below the FGCTL.
For Se, lowest leaching occurred at pH 5.5 – 7 (Figure A-3). When compared to SPLP
results, Se LEAF concentrations (0.00 – 2740 µg L-1, 25th percentile = 49) were similar,
with most samples exceeding the FGCTL at 35 µg L-1 while As SPLP results were lower
than the LEAF results. This is likely due to difference in the extraction solution, i.e.,
88
different acids used to adjust the pHs. Kosson et al. (Kosson et al., 2014) also noticed
similar behavior, though their concentrations were much lower than those in this study
(Figures A-2 and A-3). However, for As, those results are only relevant under aerobic
environments where As is predominantly arsenate (AsV). Under anaerobic conditions,
As tends to exist as arsenite (AsIII), which is more mobile due to its neutral charge.
During combustion, arsenic associated with pyrite in coal is decomposed,
releasing soluble AsV species in the ash (Goodarzi et al., 2008). They are
characterized by a pH-dependent leaching. In acidic fly ash, leaching increases with pH;
whereas, in alkaline fly ash, this trend is reversed, displaying a plateau of maximum
solubility at pH 7–11 (Izquierdo and Querol, 2012). Other substances in fly ash may
also play a role in arsenic leaching, including Ca, Fe oxyhydroxides and Al and Mn
oxides (Izquierdo and Querol, 2012). For Se, its solubility is controlled by redox potential
and pH. In addition, the presence of Ca and Mg also affects Se speciation (Kabata-
Pendias et al., 1984).
Ba, Cr, Pb and Cd Concentrations
All barium concentrations at all pHs were lower than the FGCTL of 2,000 µg L-1.
Conversely, Kosson et al. (Kosson et al., 2014) noticed concentrations higher than the
FGCTL in pH > 5.5. Different from Ba, Cr concentration was the greatest at extremes
pH (> 8), with the lowest being at pH 4 – 5.5 (0.00 – 563 µg L-1), with 75% of the
samples being lower than the FGCTL of 100 µg L-1 (Figure A-3). In a previous study,
hexavalent Cr (CrVI) was the main form of Cr at pH > 7 (Kosson et al., 2009). USEPA’s
study also found that post-combustion NOx controls enhanced Cr concentration in fly
ash (Kosson et al., 2009). As an oxyanion, CrVI is mobile in the environment and it is a
89
strong oxidizer, being toxic to biological systems and a human carcinogen (Wilbur,
2000).
Similar to Cr, Pb leaching was the highest at pH < 5.5 (0.46 – 4,909 µg L-1,
median = 57.1 µg L-1) and pH > 10.5 (0.00 – 81 µg L-1, median = 16.0 µg L-1) (Figure A-
3). Most of the SPLP results were higher than those using LEAF 1313 method. At pH
4-10, Pb concentrations were low (0.00 – 732 µg L-1), with only two samples exceeding
the FGCTL of 15 µg L-1. Amphoteric behavior of Pb was noted by Kosson et al.
(Kosson et al., 2009). In high pH (> 7), Pb can precipitate as carbonate or phosphate,
which are more stable (Kabata-Pendias et al., 1984).
Cadmium had a typical leaching behavior of decreasing concentration with
increasing pH (Figure A-3). However, all samples exceeded the FGCTL for Cd at 5 µg
L-1. In USEPA’s CCRs characterization, Cd also exceeded the limit (Kosson et al.,
2009). As pH increases (pH > 7.5), Cd starts to precipitate so CdCO3 and/or Cd3(PO4)2
play an important role in controlling Cd solubility (Kabata-Pendias et al., 1984). Most
studies reported low Cd concentration for fly ash based on SPLP or TCLP (Kosson et
al., 2009; Nathan et al., 1999). The difference might be in the final pH because neither
SPLP nor TCLP controls final solution pH. Due to the buffering capacity of fly ashes, pH
tends to be > 7, leading to low Cd leaching. Therefore, Cd may be a limiting factor in
reusing fly ash due its high toxicity and concentration if pH is not properly managed.
Metal leachability under different pHs can be used to predict their fate and
transport in risk assessment. However, metal leaching depends on site conditions, ash
type, and solution pH (Roessler et al., 2015). Likewise, other factors such as sorption
and/or co-precipitation with metal oxides also determine their fate in the environment
90
(Evans, 2008; Kogbara et al., 2014). Kosson et al. (Kosson et al., 2014) studied metal
leaching from fly ash-based concrete (up to 45% fly ash) and reported As leaching up to
10 mg L-1 using LEAF Method 1313. On the other hand, another study reported that hot-
mix asphalt could facilitate metal leaching and mobility in the environment (Roessler et
al., 2015). Thus, it is recommended that determination of maximum solubility capacity
and pH are needed prior to beneficial use of these CCRs.
Metal Concentrations in Fly Ash Based on LEAF Method 1316
Metal leaching in CCRs is impacted by pH as well as liquid to solid ratio (LS).
Therefore, it is important to understand how LS influences metal release. When the LS
ratio was increased from 0.5 to 10, metal concentrations decreased in most samples
except for Ba and Pb (Figure A-4). Concentrations of As, Ba, Cd, Cr, Hg, Pb and Se
based on EPA Method 1316 for all CCRs can be found in Table A-7. This might indicate
that all elements reached their maximum leaching at low LS of 0.5 (Nadeesha H
Koralegedara et al., 2017). For Ba and Pb, its precipitation with sulfates, phosphate and
carbonates might be responsible for its lower concentration (Kabata-Pendias et al.,
1984). The highest Pb concentration of 1,889 µg L-1 was observed at pH < 2 and the
highest Ba concentration of 1,501 µg L-1 was observed at pH > 11. Those data are in
agreement with Ba and Pb LEAF 1313 behavior and with Kosson et al. (Kosson et al.,
2014).
Concentrations of As, Cd, Cr, and Se in most samples exceeded the FGCTL
even at LS ratio of 0.5. Compared to the SPLP data at LS ratio of 20, no trend was
observed. For some samples, the SPLP followed the same trend, but for others, the
SPLP data were higher. Lead concentration was low in most samples, possibly due to
PbSO4 precipitation from Pb reaction with CaSO4 (Tang and Steenari, 2016).
91
Research Findings
This work characterized CCRs samples and applied the new USEPA LEAF
Method 1313 and 1316 to test metal leachability. All metals showed wide concentration
ranges, with most lower than Florida industrial SCTLs. Cobalt and Ni exceeded Florida
residential SCTLs in some cases while As, Cr, Cu, Cd and Pb were enriched in fly ash.
SPLP data were higher in fly ash and bottom ash compared to FGD. All samples
showed high buffering capacity with final SPLP pH close to the initial pH.
Most of the metals in fly ash showed amphoteric behavior with higher
concentrations at very low and high pH values based on LEAF Method 1313 except Cd.
Barium was the only element that did not exceed the FGCTL at all pHs. Concentrations
for all metals in fly ash was low at typical soil pH levels. When the LS ratio was
increased, metal concentrations in fly ash decreased in most cases. However, Ba and
Pb showed an opposite trend, probably due to possible precipitation as sulfates,
phosphates and carbonates. Comparison between the SPLP and LEAF data for As, Cd,
Cr, Pb and Se in fly ash showed different methods produced different results.
As pH and LS ratio were the major variables that controlled metal release from
CCRs, therefore, assessing wide range of those variables is possible to predict metal
behavior in the environment. Thus, it is important to use appropriate method to predict
metal solubility characteristics prior to beneficial use. It is also important to note that
metal solubility in fly ash also depends on other factors in the environment. Beneficial
use should be engineered to isolate CCRs from either rainfall infiltration or contact with
groundwater to reduce metal leachability, thereby minimizing potential adverse impacts
on the environment.
92
Table A-1. Total concentrations of trace metals in 24 coal combustion residual samples from 7 representative power plants (mg/kg)
As Cd Co Cr Cu Ni Pb Se Zn
Fly ash (n=11)
Max 73.0 4.80 398 398 692 9,768 1,018 58.0 3212
Min 36.8 0.20 2.62 6.94 23.8 13.0 21.0 3.83 10.7
Mean 52.6 2.13 44.5 63.7 103 916 156 21.6 376
St dev 15.3 1.46 117 114 196 2936 310 14.5 942
Median 46.0 2.55 8.44 16.0 37.3 32.0 32.0 21.0 113
Bottom ash (n=8)
Max 53.0 1.24 249 72.8 299 5992 369 39.0 524
Min 1.55 0.12 2.03 1.67 6.98 5.97 0.54 0.80 6.20
Mean 11.3 0.57 32.3 22.9 52.4 767 50.5 7.68 101
St dev 17.1 0.50 81.2 26.6 100 2111 129 13.3 178
Median 5.76 0.38 5.55 10.7 14.5 12.5 3.87 1.71 28.0
FGD (n=4)
Max 1.37 0.39 0.29 13.7 3.87 11.1 11.1 17.0 66.4
Min 0.40 0.01 0.19 0.19 1.74 0.37 0.37 0.00 4.69
Mean 0.81 0.13 0.22 6.47 2.64 3.34 3.34 7.09 22.1
St dev 0.42 0.18 0.05 7.29 1.00 5.18 5.18 7.13 29.6
Median 0.74 0.06 0.20 6.00 2.47 0.94 0.94 5.68 8.62
93
Table A-2. Concentration range of trace metals in coal combustion residuals based on literature (mg/kg).
As Cd Co Cu Cr Pb Ni Se Zn
Fly ash
Max 1,385 17 180 1,452 651 2,120 353 47 2,880
Min 1.1 <0.11 7.3 45 11 <1.40 23 <1.40 25
Bottom ash
Max 56.0 <5.50 NA 146 4,710 843 1,267 8.20 717
Min 0.36 NA NA 20.0 20.0 1.40 <12 <1.25 3.80
FGD
Max 11.0 0.37 <1.00 3.20 24.0 2.00 2.40 32.0 23.0
Min <1.90 <0.02 <1.00 <0.40 0.60 <1.00 <0.20 <2.50 <1.25
Source: Gallardo et al., 2015; McNally et al., 2012; Rowe et al., 2002; EPRI, 2010; Nathan et al., 1999; NA: data not available
94
Table A-3. Total concentrations of trace metals in 24 coal combustion residual samples from 7 representative power plants (mg kg-1)
As Cd Co Cu Cr Hg Ni Pb Se Zn
Fly ash
1 36.9 2.7 5.87 23.8 63.5 0.7 25.6 25.2 21.9 113
2 36.8 3.0 5.59 25.9 74.2 0.6 29.6 24.2 24.3 104
3 39.9 3.2 6.11 29.0 65.8 1.3 54.8 36.6 32.4 146
4 72.6 0.3 2.62 37.3 6.9 8.7 13.0 1018 7.1 10.7
5 73.0 0.2 16 61.3 16.1 0.8 24.0 21.4 21.1 26.0
6 43.9 2.5 8.31 35.1 65.6 1.1 31.5 31.0 16.1 147
7 40.7 2.5 8.44 33.8 74.2 1.5 33.1 29.6 16.0 147
8 72.8 4.8 398 692 547 0.4 9768 427 57.7 3212
9 45.8 0.6 15 64.6 24.4 0.2 30.2 37.7 3.8 67.4
10 46.6 2.7 8.7 17.3 53.0 2.5 1655 114.3 26.5 508
11 66.1 0.8 13.7 70.5 34.3 -0.2 33.3 38.2 23.1 47.0
12 49.6 2.8 10 55.1 116 0.1 38.2 31.6 13.6 120 Bottom ash
1 7.69 1.16 13.3 39.5 63.7 4.78 72.3 11.0 3.855 160
2 3.49 0.00 2.03 10.9 1.82 1.16 5.97 0.61 0.90 7.83
3 1.55 0.00 2.40 14.9 1.67 0.17 7.93 0.54 0.80 6.20
4 3.90 0.15 2.45 6.98 8.39 0.16 7.70 4.06 1.05 26.7
5 2.59 0.12 2.15 7.75 8.49 0.38 7.54 3.35 0.83 28.5
6 53.0 1.24 249 299 72.8 1.10 5992 369 38.9 524
7 10.6 0.29 4.71 25.8 10.7 0.00 23.9 11.9 12.6 33.5
8 7.61 0.47 7.74 14.4 22.4 1.19 16.6 3.68 2.37 23.0 FGD
1 0.40 0.39 0.20 1.92 11.8 4.20 1.71 11.15 0.00 10.5
2 0.60 0.01 0.19 3.87 8.43 0.66 1.84 0.37 5.80 4.69
3 0.87 0.10 0.29 1.74 13.7 0.29 2.81 1.03 17.0 66.4
4 1.37 0.01 0.20 3.02 10.5 1.12 1.70 0.85 5.56 6.74
95
Table A-4. pH of 24 coal combustion residual samples from 7 representative power plants before and after SPLP test
Fly ash (n=11) Bottom ash (n=8) FGD (n=4)
Before After Before After Before After
MEAN 8.61 8.89 7.23 8.28 7.15 7.50
MAX 11.9 12.2 10.8 10.9 10.8 10.9
MIN 1.74 1.83 1.74 3.58 1.74 2.44
ST DEVIATION 3.53 3.58 2.00 1.95 0.64 0.11
MEDIAN 9.57 10.2 8.44 9.24 7.96 8.14
25th PERCENTILE 6.37 6.89 7.53 8.21 7.44 8.04
96
Table A-5. SPLP concentrations of trace metals in 24 fly ash, bottom ash and FGD samples from 7 representative power plants (µg L-1).
As Cd Co Cr Cu Hg Ni Pb Se Zn Fly ash
1 77.3 5.33 1.67 135 31.7 0.33 2.67 0.00 7.05 115
2 111 8.33 4.67 357 50.3 0.00 24.7 3.7 32.0 0.00
3 80.7 5.67 4.00 102 27.0 0.00 52.0 21.7 58.3 0.00
4 587 4.33 70.3 123 748 146 347.7 11403 78.7 0.00
5 81.3 0.10 7.25 10.9 14.1 0.00 14.8 1.60 101 0.00
6 21.9 4.25 0.65 76.0 5.10 0.40 0.95 19.4 165 0.00
7 14.4 3.80 0.35 65.0 3.30 0.00 1.40 20.7 165 34.1
8 40.4 4.10 282 97.1 195 0.00 7714 154 242 2838
9 87.3 24.0 83.7 54.0 1395 1.67 310 35.3 252 479
10 46.0 4.50 1.50 125 21.4 1.80 5.05 10.6 285 24.9
11 51.5 0.55 1.00 189 12.3 0.00 1.60 3.00 506 4.50
12 87.2 23.2 67.5 11.3 7.67 0.00 16.0 43.0 82.2 3118 Bottom ash
1 25.0 0.67 61.7 28.3 22.7 1.40 43.7 0.00 0.00 0.00
2 27.3 0.00 0.00 0.95 1.60 0.00 0.00 0.30 1.75 0.00
3 15.3 0.00 0.15 0.95 18.5 0.00 0.00 2.20 0.00 6.20
4 13.7 0.00 0.25 6.50 1.50 0.00 1.30 1.70 1.30 0.00
5 7.60 0.00 0.10 2.15 0.00 0.00 0.00 0.65 0.70 0.00
6 4.33 6.33 40.7 0.00 3.00 15.3 318 10.0 0.00 393
7 13.2 0.05 14.5 13.3 19.2 0.00 419 19.6 2.05 33.2
8 3.95 0.00 0.15 16.0 4.20 0.00 0.05 1.75 18.9 0.00 FGD
1 0.00 1.00 0.00 0.33 7.00 0.00 0.00 0.00 30.7 0.00
2 0.65 0.00 0.05 1.70 1.55 0.00 0.00 0.15 10.6 0.00
3 4.00 0.33 0.00 5.00 12.7 4.67 11.7 11.0 0.00 13.0
4 16.7 0.00 0.30 8.25 1.55 0.00 0.60 1.45 6.00 4.75
97
Table A-6. USEPA LEAF 1313 concentrations of trace metals in 8 fly ash samples from 7 representative power plants (n=3).
Sample pH As Ba Cd Cr Hg Pb Se
-------------------------------------- µg L-1 ------------------------------------
1
2
134 397 221 2079 0.00 95.5 162
2 2503 380 243 3950 0.00 189 626
3 1066 887 304 4201 0.00 271 753
4 936 7.25 8.01 310 0.00 1721 25.3
5 6497 1281 18.7 763 0.00 101 2197
6 847 437 255 3911 0.00 232 350
7 455 813 267 3803 0.00 268 287
8 1271 719 42.8 1356 0.00 205 226
9 184 98.9 241 583 0.00 2.20 628
10 389 248 551 2242 0.00 4909 488
11 854 749 63.5 1768 0.00 102 522
12 329 943 253 5558 0.00 394 299
1
4
4.86 144 94.4 13.3 0.00 2.83 49.1
2 12.1 417 191 18.1 2.30 0.46 79.4
3 10.6 181 270 10.8 0.00 11.8 121
4 4.70 46.9 9.58 1.37 0.00 732 30.1
5 101 272 16.4 169 0.00 3.56 234
6 14.9 428 234 114 0.00 7.36 153
7 13.7 632 250 104 0.00 7.98 144
8 95.9 202 35.3 563 0.00 2.14 129
9 20.2 56.0 60.9 8.04 0.00 1.54 18.3
10 27.6 39.9 307 0.00 0.00 5.37 296
11 15.6 82.3 46.1 8.66 0.00 13.2 59.2
12 16.4 292 253 200 0.00 18.7 153
1
5.5
6.31 130 59.3 8.52 0.00 0.00 60.0
2 4.86 365 141 70.2 5.54 0.00 48.6
3 4.32 155 239 0.64 0.65 0.64 77.6
4 0.71 78.3 1.71 0.00 0.00 10.3 17.0
5 18.3 100 5.31 0.00 0.00 1.95 120
6 8.35 420 181 5.44 0.00 0.00 269
7 5.02 471 197 18.5 0.00 0.00 211
8 12.0 139 33.6 384 0.00 0.00 51.7
9 10.6 20.9 6.42 0.00 0.00 0.72 6.27
10 8.24 46.5 194 1.69 0.00 0.04 360
11 13.0 131 47.6 0.70 1.23 0.75 38.8
12 3.30 244 229 1.06 0.00 0.00 40.7
98
Table A-6. Continued
Sample pH As Ba Cd Cr Hg Pb Se
1
7
8.54 138 10.6 21.2 0.00 0.00 94.8
2 8.82 331 40.6 169 3.35 4.56 88.8
3 5.03 133 16.6 17.7 0.00 0.00 124
4 0.70 90.4 1.05 0.00 0.00 0.00 17.1
5 37.8 194 11.4 22.3 0.00 1.10 248
6 68.4 377 44.8 87.2 0.00 0.00 313
7 53.1 413 27.2 123 0.00 0.00 303
8 96.2 117 5.82 460 0.00 0.61 10.9
9 25.0 13.1 14.7 0.66 0.00 0.00 9.8
10 9.84 42.8 38.1 12.4 0.00 0.00 379
11 11.9 130 4.19 16.1 2.22 0.00 97.6
12 3.09 193 81.7 26.7 0.00 0.00 49.1
1
8
10.5 154 7.36 39.7 0.00 0.00 140
2 19.2 348 7.98 308 0.00 0.00 193
3 11.0 132 10.52 44.3 0.39 0.00 237
4 0.98 94.2 0.89 0.00 0.00 0.00 22.0
5 169 200 1.51 71.7 0.00 0.00 1129
6 67.4 500 12.2 104 0.00 0.00 367
7 51.4 611 12.4 134 0.00 0.00 345
8 204 111 2.77 490 0.00 0.00 18.2
9 49.1 7.6 16.0 4.35 0.00 0.00 16.0
10 152 4.5 5.96 55.1 0.00 0.00 895
11 13.8 131 3.32 26.4 2.37 0.00 115
12 12.7 138 8.09 115 0.00 0.00 131
1
9
11.7 240 6.75 88.9 0.96 0.15 305
2 17.9 353 7.27 328 0.00 0.00 236
3 12.2 49.5 6.81 98.6 0.00 0.00 443
4 1.18 82.6 0.96 0.00 0.00 0.00 26.8
5 230 360 1.52 69.9 0.00 0.00 1685
6 56.4 781 11.1 115 0.00 0.00 401
7 33.4 890 11.1 152 0.00 0.00 373
8 159 149 2.58 483 0.00 2.13 19.1
9 46.5 6.65 14.3 10.5 0.00 0.00 5.12
10 412 1.78 6.36 67.0 0.00 0.00 1544
11 137 296 3.06 77.9 7.36 0.00 902
12 9.48 150 8.27 159 0.00 0.00 199
99
Table A-6. Continued
Sample pH As Ba Cd Cr Hg Pb Se
1
10.5
17.7 65.9 6.07 104 0.00 0.00 308
2 13.1 945 8.28 376 0.00 0.00 354
3 18.5 241 9.40 144 2.05 0.00 856
4 3.34 57.2 1.78 0.00 5.76 0.00 37.5
5 276 696 1.48 75.5 0.00 0.00 2215
6 19.1 1880 10.3 169 0.00 0.00 290
7 18.0 1675 10.1 184 0.00 0.00 306
8 36.0 791 2.70 482 0.00 0.07 23.7
9 130 2.37 15.8 61.4 0.00 0.00 0.00
10 1029 1.13 7.53 74.5 0.00 0.00 2663
11 202 509 3.03 98.8 8.14 0.00 1331
12 13.8 439 9.43 244 0.00 0.00 150
1
12
35.1 438 5.77 129 0.00 1.99 687
2 3.17 166 7.71 404 0.00 0.00 27.4
3 30.4 327 9.23 169 4.00 0.00 901
4 36.3 14.3 1.71 0.00 14.8 0.00 187
5 1410 612 1.34 73.5 0.00 3.47 2285
6 5.89 682 15.7 511 0.00 16.0 39.4
7 5.17 871 14.4 414 0.00 23.5 29.2
8 113 1123 2.66 380 0.00 19.0 35.8
9 1384 2.94 19.7 244 0.00 0.00 100
10 1715 0.11 7.68 115 0.00 2.03 2740
11 1065 304 2.73 119 8.29 4.71 1429
12 20.9 336 8.8 181 0.00 2.80 17.7
1
13
121 590 5.47 112 0.00 18.1 823
2 45.5 330 5.23 294 0.00 26.4 680
3 111 378 8.42 160 46.4 5.23 1055
4 698 2.16 1.87 0.00 60.6 0.48 756
5 4208 842 1.13 83.3 0.00 19.8 2072
6 56.6 1617 6.10 440 0.00 64.3 586
7 63.6 1508 7.65 488 0.00 78.8 742
8 457 1220 2.49 347 0.00 65.0 30.5
9 2355 3.07 22.0 434 0.00 0.00 445
10 1414 0.99 6.15 114 0.00 30.3 1966
11 3872 252 2.54 141 6.20 35.7 1258
12 86.4 306 5.69 127 0.00 81.3 442
100
Table A-7. USEPA LEAF 1316 concentrations of trace metals in 8 fly ash samples from 7 representative power plants (n=3).
Sample LS ratio As Ba Cd Cr Hg Pb Se ------------------------------------ µg L-1 ---------------------------------------
1
LS=0.5
17 122 137 616 8.6 0.00 2485
2 6.8 289 76 1699 0.0 2.58 346
3 64 74 216 1074 10 0.00 4324
4 3118 2.1 23 967 6.8 726 31
5 16 37 15 0.7 0.0 0.0 208
6 3.0 607 39 682 0.0 15 189
7 3.7 646 40 696 0.0 15 213
8 1.42 89 18 2051 0.0 0.00 16
9 0.0 0.0 0.0 0.00 0.0 0.00 0.0
10 10 18 8.6 0.9 0.0 0.0 22
11 71 108 666 155 0.0 21 498
12 3.0 51 82 1184 0.0 0.00 87
1
LS=1
12 152 57 538 5.4 0.00 1497
2 9.3 377 51 983 0.0 1.70 281
3 37 91 83 825 4.9 0.00 2384
4 2758 2.3 20 806 0.5 997 47
5 20 53 8.5 13 0.0 0.0 258
6 5.6 1111 35 590 0.0 22 200
7 5.8 1037 35 620 0.0 22 194
8 53 226 19 2131 0.0 0.00 13
9 21 14 6.67 0.00 0.0 0.00 7.8
10 38 44 3287 2.5 0.0 0.9 2906
11 54 135 380 100 0.0 20 337
12 17 130 48 1295 0.0 7.64 79
1
LS=2
8.0 149 32 396 6.1 0.00 908
2 10 513 36 640 1.2 3.70 170
3 19 92 41 527 5.5 0.00 1365
4 2649 3.7 19 693 0.7 1640 53
5 22 64 14 5.3 0.0 0.0 181
6 6.7 1313 27 376 0.0 26 166
7 6.7 1261 26 398 0.0 28 149
8 57 411 13 1313 2.2 0.00 13
9 14 15 7.2 0.0 0.5 0.00 9.93
10 34 40 1210 0.0 1.4 0.0 1035
11 40 141 217 77 0.0 18 222
12 19 190 32 773 0.0 0.00 48
101
Table A-7. Continued
Sample LS ratio As Ba Cd Cr Hg Pb Se
1
LS=5
13 181 17 191 1.1 0.00 583
2 13 622 19 351 6.1 6.52 95
3 18 134 20 248 11.5 814
4 1739 4.9 12 506 7.5 1889 48
5 18 87 6.6 2.3 2.4 0.0 147
6 7.8 1374 13 161 2.3 27 130
7 8.2 1501 13 172 1.6 32 124
8 46 597 5.7 696 15.4 4.10 11
9 11 17 6.8 0.0 8.2 0.00 9.54
10 22 47 516 0.0 15 6.1 467
11 25 97 90 32 9.7 11 94
12 19 285 18 399 2.7 0.00 33
1
LS=10
18 66 6.1 104 0.0 0.00 308
2 3.2 166 7.7 404 0.0 0.00 27
3 12 50 6.8 99 0.0 0.00 443
4 936 7.2 8.0 310 0.0 1721 25
5 18 100 5.3 0.0 0.0 1.9 120
6 5.9 682 16 511 0.0 16 39
7 5.2 871 14 414 0.0 24 29
8 55 601 3.2 440 3.6 8.08 5.63
9 11 21 6.4 0.0 0.0 2.72 6.3
10 28 40 307 0.0 0.0 5.4 296
11 16 82 46 8.7 0.0 13 59
12 19 318 11 423 0.0 0.00 9.8
102
Figure A-1. Total concentrations of trace metals in 24 fly ash, bottom ash and FGD from 7 power plants. The Box-and-Whisker plots show the following: the minimum value (the lower whisker), the 25th quartile, the median, the 75th quartile, the maximum value (the upper whisker) and outlier (●). (n= 11, 8 and 4 for fly ash, bottom ash, and FGD)
103
Figure A-2. SPLP concentrations of trace metals in 24 fly ash, bottom ash and FGD samples from 7 representative power plants. The Box-and-Whisker plots show the following: the minimum value (the lower whisker), the 25th quartile, the median, the 75th quartile, the maximum value (the upper whisker) and outlier (●). (n= 11, 8 and 4 for fly ash, bottom ash, and FGD)
104
Figure A-3. USEPA LEAF 1313 concentrations of trace metals in 8 fly ash samples from
7 representative power plants. The Box-and-Whisker plots show the following: the minimum value (the lower whisker), the 25th quartile, the median, the 75th quartile, the maximum value (the upper whisker) and outlier (●).
pH
2 4 6 7 8 9 11 12 13
Cr
co
nc
en
tra
tio
n (
µg
/kg
)
0
100
200
300
400
500
1500
3000
4500
6000
pH
2 4 6 7 8 9 11 12 13
Cd
co
nc
en
tra
tio
n (
µg
/kg
)
0
5
10
15
20
25
120
240
360
480
600pH
2 4 6 7 8 9 11 12 13
As c
on
cen
trati
on
(µ
g/k
g)
0
35
70
105
140
175
1500
3000
4500
6000
pH
2 4 6 7 8 9 11 12 13
Ba
co
nc
en
tra
tio
n (
µg
/kg
)
0
50
100
150
200
450
900
1350
1800
pH
2 4 6 7 8 9 11 12 13
Se c
on
cen
trati
on
(µ
g/k
g)
0
75
150
225
300
750
1500
2250
3000
pH2 4 6 7 8 9 11 12 13
Pb
co
nc
en
tra
tio
n (
µg
/kg
)
0
100
200
300
400
1000
2000
3000
4000
5000
As
Co
nc
en
tra
tio
n (µ
g/L
)
Ba
Co
nc
en
tra
tio
n (µ
g/L
)
Cr
Co
nc
en
tra
tio
n (
µg
/L)
Cd
Co
nc
en
tra
tio
n (µ
g/L
)
Pb
Co
nc
en
tra
tio
n (µ
g/L
)
Se
Co
nc
en
tra
tio
n (
µg
/L)
2 4 5.5 7 8 9 10.5 12 13 2 4 5.5 7 8 9 10.5 12 13
2 4 5.5 7 8 9 10.5 12 13 2 4 5.5 7 8 9 10.5 12 13
2 4 5.5 7 8 9 10.5 12 13
pH
2 4 5.5 7 8 9 10.5 12 13
pH
105
Figure A-4. USEPA LEAF 1316 concentrations of trace metals in 8 fly ash samples from
7 representative power plants. The Box-and-Whisker plots show the following: the minimum value (the lower whisker), the 25th quartile, the median, the 75th quartile, the maximum value (the upper whisker) and outlier (●).
2D Graph 1
LS Ratio
0.5 1.0 2.0 5.0 10.0
Cd
co
ncen
trati
on
(µ
g/k
g)
0
25
50
75
100
800
1600
2400
3200
0.5 1 2 5 10
2D Graph 1
LS Ratio
0.5 1.0 2.0 5.0 10.0
Cr
co
ncen
trati
on
(µ
g/k
g)
0
500
1000
1500
2000
2500
0.5 1 2 5 10
2D Graph 1
LS Ratio
0.51.0 2.0 5.0 10.0
As
co
nc
en
tra
tio
n (
µg
/kg
)
0
15
30
45
60
75
750
1500
2250
3000
2D Graph 1
LS Ratio
0.5 1.0 2.0 5.0 10.0
Ba
co
nc
en
tra
tio
n (
µg
/kg
)
0
350
700
1050
1400
1750
0.5 1 2 5 10
2D Graph 1
LS Ratio
0.5 1.0 2.0 5.0 10.0
Se c
on
cen
trati
on
(µ
g/k
g)
0
50
100
150
200
800
1600
2400
3200
4000
0.5 1 2 5 10
2D Graph 1
LS Ratio
0.5 1.0 2.0 5.0 10.0
Pb
co
nc
en
tra
tio
n (
µg
/kg
)
0
5
10
15
20
500
1000
1500
2000
0.5 1 2 5 10
As
Co
nc
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g/L
)
Ba
Co
nc
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n (µ
g/L
)
Cr
Co
nc
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n (
µg
/L)
Cd
Co
nc
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n (µ
g/L
)
Pb
Co
nc
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n (µ
g/L
)
Se
Co
nc
en
tra
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n (
µg
/L)
106
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BIOGRAPHICAL SKETCH
Evandro Barbosa da Silva, the second of the four siblings, was born in Sorocaba,
Brazil. He went to University of Sao Paulo to pursue his bachelor’s degree in
environmental management. Following this, he switched his major to soil science and
obtained his master degree in the same University. His master’s thesis was titled
“Watershed soils contents of potential toxics elements in soil from a watershed and
human health risk assessment”. In 2013, he came to the University of Florida to pursue
a Ph.D. in soil and water science supported by the Brazilian National Council for
Scientific and Technological Development under the supervision of Dr. Lena Q. Ma. He
worked on phytoremediation of arsenic contaminated soils by Pteris vittata and As-
laden biomass disposal methods and he is expected to receive his Ph.D. from the
University of Florida in the spring of 2018.
