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LABORATORY STUDY OF INTRINSIC DEGRADATION OF ORGANIC POLLUTANTS IN COMPACED CLAYEY SOIL
by
Leila Hrapovic
Graduate Program in Engineering Science Department of Civil and Environmental Engineering
Submitted in partial fulflment of the requirements for the degree of
Doctor of Philosophy
Faculty of Graduate Studies The University of Western Ontario
London, Ontario December, 2000
0 Leila Hrapovic 200 1
LABORATORY STUDY OF INTRINSIC DEGRADATION OF ORGANIC
POLLUTANTS IN COMPACTED CLAYEY SOIL
This thesis examines the intrinsic biodegradation of selected organic pollutants
under diffisive transport in laboratory compacted clayey soil. A series of 16 ceUs
containing clayey soi1 were placed in contact with an essentiaily constant source solution
consisting of a synthetic blend of volatile fatty acids (VFAs), BTEX and chlorinated
aliphatic solvents. Results fiorn the tests indicate that as they difise through the soil,
most of the pollutants tested degrade slowly but steadily in the soil, even under adverse
conditions imposed by compaction and flow limitations. The process is modeled as
diffusion coupled with linear sorption and first order biological decay. Estimates of the
intrinsic degradation rates for selected chemicals under difisive transport through the
clay under taboratory conditions are given in this thesis.
Separate series of difision-degradation tests were performed with Keele Valley
Landfill leachate with an emphasis on the fate of dichioromethane alone. These tests
provide additional evidence of dichloromethane degradation in the compacted soil plugs.
Very fast anaerobic degradation of dichloromethane is also observed in the batch tests
with the Keele Valley Landlfill leachate.
Di ffision coefficients are obtained from independent short-term tests with the
same soil and solution medium. Batch equilibrium tests are performed in order to obtain
the sorption coefficients for the chemicals. The diffusion and linear sorption coefficients
deduced for the 11 pollutants examined are also presented. Results from the shon-term
tests are in excellent agreement with those obtained for modeling of the long-term
intrinsic degradation under diffusive transport. The impact of VFAs, dichloromethane,
benzene and xylenes on an aquifer under a hypothetical landfill was modeled using the
values of difisioq sorption and degradation coefficients determined in the experiments.
The results of the transport simulation for a simple banier with primary and secondary
leachate collection system and clay liner indicate that intrinsic degradation could
significantly reduce contamination of underlying groundwater.
Chapter 3
Significant portions of this chapter were published in the Journal of Geotechnical and Geoenvironrnentai Engineering, Vol. 123, No. 12, 1997, titled " Anaerobic Degradation of DCM Diffising through Clay", authored by R Kerry Rowe, Leila Hrapovic, Naim Kosanc and D. Roy Cullirnore, Copyright by American Society of Civil Engineers, 1801 Alexander Bell Drive, Reston, VA 2019 1-4400; Reprinted with permission fiom the publisher (ASCE) as granted on 09. Feb. 2001 (see Appendk 3, page 274)
Contributions:
R. K. Rowe initiated the project, contributed advice on the subject, assisted in test results interpretation and wrote the final version of the paper. L . Hrapovic: designed the apparatus, developed the testing procedure, performed the testing, modeled and interpreted test results and assisted in writing of the paper N. Kosanc provided advice and suggestions dunng the course of work and assisted in reviewing the paper. D.R. Cullimore provided advice and suggestion on microbiological testing and assisted in reviewing the paper.
Chapters 2,4,5,6 are not yet published
L. Hrapovic developed procedures for testing, performed the experiments and ail testing, modeled and interpreted the test results and wrote the first draft of each manuscript. R. K. Rowe: assisted in developing the procedures, provided constmctive advice and assisted in interpretation of results and writing of the manuscripts.
ACKNOWLEDGEMENTS
I would like to express sincere gratitude and appreciation to my supervisor, Dr. R.
Kerry Rowe who provided support, guidance and encouragement through the entire
course of this work. His unique generosity, tolerance and many usefbl advices are
grat efully acknowledged.
I also wish to thank Drs. N. Kosaric and D. R. Collimore for many usefil
suggestions and valuable expertise provided to this topic.
Very special thanks are extended to members of the Geotechnical Research
Centre, in particular to Ms. E. Milliken, Ms. J. Lemon, Ms. C. Walter, M.. G. Lusk and
Mr. W. Logan for being available, helpfùl and kind to so many of us, graduate students.
I owe endless thanks to my collegues and fellow graduate students for their
fnendship, interest, enthusiasm and help they always offered. It is with affection and
respect that I remernber very dedicated involvement of H. P. Sangam, A. Cooke, M.
Armstrong, S. Micic, I. Rerning, C. Lake, S. Millward, M. Kr01 and G. Lima, and
assistance each of them generously provided when 1 was in need.
This study was supported by the Naturd Sciences and Engineering Research
Council of Canada under Collaborative Research Grant CPG 0163097.
Finally, I wholeheartedly acknowledge the moa forgiving and generous family of
fnends I have in London: the Dmovics, the Kohans, the Munozs, the Stucin-Pekedizs,
who gave me so strong suppon, encouragement and inspiration. It is to you that I owe the
most gratitude, recognition, and this thesis.
TABLE OF CONTENTS
CERTIFICATE OF EXAMINATION ........................................................................ ii ... ABSrlTRACT ................................................................................................................. 111
CO-AUTHORSEiIP ...................................................................................................... iv ACKNOWLEDGEMENTS .......................................................................................... v TABLE OF CONTENTS ............................................................................................ *vi LIST OF TABLES ...................................................................................................... i x LIST OF FIGURES ...................................................................................................... .x LIST OF APPENDICES ............................................................................................ xvi . . LIST OF ABBREVIATIONS, SYMBOLS, NOMENCLATURE .........................a. wii
CR4PTlER 1 INTRODUCTION .................................................................................. 1 1 . 1 General .................................................................................................................. 1 . . 1 -2 Research objectives ............................................................................................... -4 1.3 Thesis outline ....................................................................................................... 5 1.4 References ............................................................................................................. 6
CHAPTER 2 BATCH DICHLOROMETHANE DEGRADATION TESTS ............ 10 ......................................................................................................... 2.1 Introduction 10
2.2 Materials, Methods, Theoretical consideration and Data analysis ......................... 13 2.2.1 Materials: Keele Valley Landfill leachate, synthetic leachate and soil ........... 13 2.2.2 Methods ........................................................................................................ 15
2.2.2.1 Testing .................................................................................................. 15 2.2.2.2 Anal ytical measurements ....................................................................... 16
2.2.3 Theoretical considerations ............................................................................. 17 ................................................................................................. 2.2.4 Data analysis 20
2.3 Results and discussion ........................................................................................ 21 2.3.1 DCM degradation in the KVL leachate .......................................................... 21 2.3.2 DCM degradation in a KVLUSoil suspension ............................................... 26 2.3.3 DCM degradation in Soi Water controls ...................................................... -27 2.3.4 DCM degradation in the Synthetic IeachatdSoil suspensions ........................ 28
......................................................................................................... 2.4 Conclusions 29 ....................................................................... 2.5 References ............................ .,. 3 1
CHAPTER 3 LABORATORY TESTING OF ANAEROBIC DEGRADATION OF DICHLOROMETBANE UNDER DlFFUSIVE TRANSPORT THROUGH CLAY ...................................................................................................................................... 47
3.1 Introduction ......................................................................................................... 47 3.2 Test program ....................................................................................................... -48 3.3 Analytical methods .................................... .... ...................................................... 49
........................................................... ......................... 3 -4 Diffision and sorption .. 50 ..................................... 3.5 DCM diffision-degradation tests with synthetic leachate 5 1
.................... ....................................*............................... 3 S.1 Methodology ......... 1 .............................................. 3.5.2 Test results and discussion ................. ..,......... 52
3.6 DCM dimision-degradation test with KVL leachate ............................................. 5 5 3 . 6.1 Methodology ................................................................................................. 55 3.6.2 Test results and discussion ............................................................................ 56
........................................................................ 3.6.2.1 Clay Only: Cells 3 and 4 56 .................................................... 3.6.2.2 Sand over a clay plug: Cells 5, 6 and 7 58 ................................................... 3.6.2.3 Sand, granules and clay: Cells 8 and 9 5 9
.................................................................................... 3.7 Summary and conclusions 61 .......................................................................................................... 3 -8 References -63
CHAPTER 4 INTPUNSIC DEGRADATION OF VOLATILE FATTY AClDS IN LABORATORY COMPACTED CLAYEY SOIL .... ................ ................................. 79
4.1 Introduction ......................................................................................................... 79 4.2 Sorption of VFAs on clayey soils ........................................................................ -80
.............................................................................. 4.2.1 Materials and test method 81 4.2.2 Theoretical considerations and data analysis ................................................. -83 4.2.3 Analytical measurements ............................................................................. -85
.................................................. 4.2.4 Results and discussion: Batch sorption tests 85 ............... 4.3 Di f i s ion of volatile fatty acids (Wh) through compacted clayey soi1 86
4.3.1 Materials and rnethod .................................................................................... 87 .............................................................................. 4.3.2 Analytical measurements 89 ............................................................................. 4.3.2.1 Gas chromatography -89
...................... 4.3.2.2 Bacterial population size and ATP content measurements 91 .......................................................... 4.3.3 Results and discussion: Diffision tests 93
4.4 Intrinsic degradation of volatile fatty acids (MAS) under diffusive transport in compacted clayey soi1 ................................................................................................ 98
.................................................................................. 4.4.1 Materials and method 107 ............................................................................ 4.4.2 Anal ytical measurements 109
..................... 4.4.3 Results and discussion: Laboratory intrinsic degradation tests 109 .................................................................................. 4.5 Summary and conclusions 128
...................................................................... ..................... ... 4.6 References .. .. 130
CHAPTER 5 INTRINSIC DEGRADATION OF VOLATILE ORGANTC ................................ COMPOUNDS m O U G H COMPACTED CLAYEY SOIL 160
5.1 Introduction ................................................................................................... . 160 5.2 Sorption of volatile organic compounds (VOC) on clayey soi1 ........................... 161
.................................................................................. 5.2.1 Materials and method 163 ............................................................................ 5.2.2 Anal yt ical measurements 164
................................................ 5.2.3 Results and discussion: Batch sorption tests 164 5.3 Diffusion of volatile organic compounds (VOC s) through compacted clayey soi1 ...
............................................................................................................... 167 5.3.1 Materials and rnethod .................................................................................. 168
............................................................................ 5.3.2 Analytical measurernents 169 ........................................................ 5.3.3 Results and discussion: Diffision tests 170
5.4. Intnnsic degradation of VOCs under diffisive transport through compacted clayey .................................................................................................................... soi1 174
............................. .......*...... 5.4.1 Materials and method .... 178 ............................................................................ 5 .4.2 Analytical measurernents 179
5 .4.3 Results and discussion: Intrinsic degradation tests ....................................... 179 .................................................................................. 5 . 5 S u m m q and conclusions 190
......................................................................................................... 5.6 References 192
CEIAPTER 6 PREDICTXON OF CONTAMINATION IMPACT FOR SELECTED ORGANIC CHEMICALS IN AQUIFER FROM A BYPOTHETICAL LANDFILL .................................................................................................................................... 229
6.1 Introduction ....................................................................................................... 229 6.2 Hypothetical hydrogeological setting and choice of contaminants ...................... 229 6.3 Biodegradation parameters ................................................................................. 233 6.4 Results and discussion ....................................................................................... 236
6.4.1 DOC-VFAS ................................................................................................. 236 6.4.2 DCM ........................................................................................................... 239 6.4.3 Benzene ......................................... ,.. .......................................................... 240 6.4.4 Xylenes ....................................................................................................... 241
6.5 General remarks on transport and degradation simulations ................................. 242 6.6 References ............................. ,.. ......................................................................... 244
CHAPTER 7 CONCLUSIONS AND RECOMMENDATIONS ............................. 255 7.1 Summary and Conclusions ................................................................................. 255 7.2 Recomrnendations for firture work ..................................................................... 258
............................................................................................................... GLOSS ARY 260 APPENDIX 1 - SUPPLEMENT TO CHAPTER 1 ...........................*...................... 267
.................................................. APPENDIX 2 - SUPPLEMENT TO CELAPTER 2 269
................................................. APPENDK 3 - SUPPLEhlENT TO CHAPTER 3 2 7 3 APPENDIX 4 - SUPPLEMENT TO CEFAPTER 4 ................................................ 2 7 7 APPENDN 5 - SUPPLEMENT TO CHAPTER 5 .................................................. 287
LIST OF TABLES
CEIAPTER 1 Table 1 . 1 Sumrnary of physical - chernical properties for selected organic
compounds .................................................................................................... 8
Table 1. 2 Drinking water objectives for tested organic chemicals ................................. 9
CHAPTER 2
Table 2. 1 Composition of media used for the batch degradation tests .......................... 36
Table 2. 2 Characteristics of the soil (Halton Tili) used in the batch degradation tests.. . . . . . . . .. . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 7
Table 2. 3 Cornparison of Monod without growth, Zero and First -order models ............. ................. ................................. . .... . ...................... . . . 38
Table 2. 4 Cornparison of Michaelis-Menten and growth linked models ...................... 39
CEIAPTER 3
Table 3. 1 Clay characteristics ............ .. . . ... . .. . . .. .. . . . .... .. . .. . . . . , . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . , . 65 Table 3. 2 Summary of bacteriological background data .............................................. 65 Table 3. 3 Dimensions of diffision - degradation cells ................................................ 66
CHAPTER 4
Table 4. 1 Summary of diffision and linear sorption coefficients for the VFAs and Halton till used for modeling .... ...... .... ... ... ..... ..... . ...... . . . . . . . 137
Table 4. 2 Composition of synthetic Keele Valley Landfill leachate (KVLL) ............................................................................................ 138
Table 4. 3 Change of soil porosity induced by intrinsic degradation of VFAs ............................................................................................... 139
CHAPTER 5 Table 5. 1 Summary of sorption parameters for the VOCs and Halton Till ................. 201 Table S . 2 Summary of diffision and linear sorption coefficients for the
VOCs and Halton Till used for rnodeling ............................................... 202 Table 5 . 3 Surnmary of half-lives used for modeling of lab-scale intrinsic
degradation of VOCs in compacted Halton Till .............................. .......... 203
CEIAPTER 6 Table 6. 1 Layer data for hypothetical landfill ............................................................ 246 Table 6. 2 Change of Darcy velocities in the layers during operation of the
hypothetical landfill.. . .. . ... .. . .- ..- .. -. -. ., . . . . . -. . .. . . . . . - . . . . . . . . . . . . . . . . . . . . . . . . . . . - - . . . . 246
LIST OF FIGURES
CHAPTER 2
Degradation of DCM in the KVLL at 24OC - Batches 1, 2, 3, 4, 5 & 6 Data and the fit h e s to the Michaelis-Menten kinetics ................................. 40
Effect of subsequent addition of DCM on its degradation rate: Batches 4 & 5 at 24' and 10°C with fit lines to the Michaelis-
.................................................................................... Menten (M-M) kinetics 41
Effect of subsequent addition of DCM on its degradation rate: Batches 4 & 5 at 24' and 10°C with the iines fit to Zero-order and Growth-linked models ..................................................................................... 42
Effect of subsequent addition of DCM on its degradation rate; Batches 4 & 5 at 10°C - Cornparison between the Zero-order and the First-order kinetics .................................................................................... 43
Degradation pattern and growth iinked kinetics lines upon single addition of DCM in KVLL: Batches 7 & 8 at 24" and 10°C .............................. 44
Degradation pattern and kinetics lines upon single addition of DCM in soi1 - KVLL suspensions - Batches 1, 7, 8 & 9 at 24°C and batches 7 & 8 at 10°C ..................................................................................... 45
............................................ Monitoring DCM concentration in M3M controls 46
DCM degradation in SoiVSynthetic leachate suspensions ................................ 46
Fig. 3. 1
Fig. 3. 2
Fig. 3. 3
Fig. 3. 4
Fig. 3. 5
Fig. 3 . 6
DCM batch sorption test results for Halton Till (modified from ......................................................... ............. Rowe & Barone, 199 1). ... 67
Schematic of the leachate diffusion-degradation test ce11 ................... ..... 68
Difision-degradation tests with synthetic leachate: Celi 1 - ..................................................................................... Receptor solution 69
Diffusion-degradation tests with synthetic leachate: Ce112 - ........................................................................................ Receptor solution 70
ATP concentration profile for Ce1 2: Diffision-degradation test with synthetic ieachate - t = 230 days ............... ... .............................. 71
Difision-degradation tests with KVL Leachate - Cells 3 & 4 - ......................................................................................... receptor solution 7 2
Fig. 3. 7
Fig. 3. 8
Fig. 3. 9
Fig. 3. 10
Fig. 3. 11
Fig. 3. 12
ATP concnentration profile for Ce11 4: Diffusion-degradation ........................................................ test with KVL Leachate - t = 153 days 73
Diffision-degradation test with KVL leachate - ceils 5 & 6 - Receptor solution ...................................... .. .............................................. 74
Difision-degradation test with KVL leachate - Ce11 7 - Receptor solution ........................................................................................ 75
Diffision-degradation test with KVL leachate - cells 8 & 9 - ........................................................................................ Effect of porosity 76
Diffision-degradation test with KVL leachate - Ce11 8 - Receptor solution ........................................................................................ 77
ATP concentration profile for Ce11 8: Diffusion-degradation test with KVL leachate - t =147 days ................................................................. 78
CHAPTER 4
Fig. 4. 1
Fig. 4. 2
Fig. 4. 3
Fig. 4. 4
Fig. 4. 4.1
Fig. 4. 5
Fig. 4. 6
Fig. 4. 7.1
Fig. 4. 8.1
Fig. 4. 9.1
Sorttion of VFAs ont0 the Halton till: linear isotherms with 95 % confidence interval of Kd: (a) acetate (b) propionate (c)
..................................................................................................... butyrate 140
Diffusion of acetate through Halton till: (a) source and receptor ........................................................................ solutions @) depth profiles 14 1
Diffusion of propionate through Halton till: (a) source and receptor solutions @) depth profiles ........................................................... 141
Diffusion of butyrate through Haiton till: (a) source and receptor ........................................................................ solutions @) depth profiles 14 1
Magnified Fig. 4.4 with details on diffision of butyrate through Halton till: Influence of D (diffision coefficient), r , . ~ ~ (half-lives)
....................................................... and Kd (linear sorption) on the best fit 142
Ce11 assemblies used for testing of intnnsic degradation: (zipper) glass cells with 3 and 5 cm plugs c o ~ e c t e d to the feed network;
................. (lower) close-up showing 8-ce11 assembly with 3 cm soi1 plugs 143
Intnnsic degradation of organic chernicals fiom synthetic KVL leachate in cornpacted Halton till: (Iefl) Cell II-lafter 162 days;
................................................................ (right) Cell IV-2 afier 163 days; 144
Variation of porosity in 3 cm thick soil plugs (Halton till, compacted); (a) time profiles (b) depth profiles ......................................... 145
Distribution of HAB and SRI3 in Halton till: 3 cm compacted soi1 plugs; (a)@) time profiles (c)(d) depth profiles .................................... 146
Distribution of ATP and f, in Halton till: 3 cm cornpacted soil plugs; (a)@) time profiles (c)(d) depth profiles .......................................... 147
Fig. 4. 10.1 Intrinsic degradation of acetate through Halton tili: 3 cm compacted soil plugs; (a) source & receptor solutions (b) depth profiles ...................................................................................................... 148
Fig. 4. 1 1. l Intrinsic degradation of propionate through halton tili: 3 cm compacted soil plugs; (a) source & receptor solutions (b) depth profiles ...................................................................................................... 149
Fig. 4. 12.1 Intrinsic degradation of butyrate through Halton till: 3 cm compacted soil plugs; (a) source & receptor solutions @) depths profiles ...................................................................................................... 150
Fig. 4. 13.1 Intrinsic degradation of VFAs (as DOC) through Halton till: 3 cm compacted soil plugs; (a) source & receptor solutions (b)
........................................................................................... depths profiles 15 1
Fig. 4. 7.2 Variation of porosity in 5 cm thick soil plugs (Halton till, compacted); (a) time profiles @) depth profiles ........................................ 152
Fig. 4. 8.2 Distribution of HAB and SRI3 in Halton till: 5 cm compacted .................................... soi1 plugs; (a)@) time profiles (c)(d) depth profiles 153
Fig. 4. 9.2 Distribution of ATP and f, in Halton till: 5 cm compaaed soi1 plugs; (a)@) time profiles (c)(d) depth profiles .......................................... 154
Fig. 4. 10.21ntrinsic degradation of acetate in through Halton till: 5 cm compacted soil plugs; (a) source & receptor profiles (b) depth profiles ...................................................................................................... 155
Fig. 4. 1 1.2 Intrinsic degradation of propionate through Halton till: 5 cm compacted soil plugs; (a) source & receptor solution (b) depth profiles ...................................................................................................... 156
Fig. 4. 12.2 Intrinsic degradation of butyrate through Halton till: 5 cm compacted soil plugs; (a) source & receptor solutions (b) depth
...................................................................................................... profiles 157
Fig. 4. 13 -2 Intrinsic degradation of VFAs (as DOC) through Halton till: 5 cm compacted soi1 plugs; (a) source & receptor solutions (b)
...................................... .................................................. depth profiles .... 158
Fig. 4. 14 Intrinsic degradation of organic chernicals from synthetic KVL leachate in compacted Halton till: Cell4-2 (with 5 cm thick plug) afier 20 1 days ............................................................................... 159
xii
Fig. 5.1.1
Fig. 5.1.2
Fig. 5.2.1
Fig. 5.2.2
Fig. 5.3
Fig. 5.4
Fig. 5.5
Fig. 5.6
Fig. 5.7
Fig. 5.8
Fig. 5.9
Fig. 5.10
Fig. 5.1 1
Fig. 5.12
Fig. 5.13
Fig. 5.14.1
Sorption of chlorinated aliphatics ont0 the Halton till: linear isotherms and 95 % confidence interval for Kd (a) DCM (b) 1,2-
............................................................................................ DCA (c) TCE 204
Sorption of chlorinated aliphatics ont0 the Halton till: data with Freundlich and Langmuir isotherms (a) DCM (b) 1,2-DCA (c) TCE .......................................................................................................... 205
Sorption of BTEX ont0 the Halton till: Data with linear isotherms (a) benzene (b) toluene (c) ethyl-benzene d) xylenes ................. 206
Sorption of BTEX ont0 the Halton till: Data with Freundlich and Langmuir isotherms (a) benzene @) toluene (c) ethyl-benzene (d) xylenes ........................................................................................... 207
Diffision of DCM through Halton till (a) source SS and receptor .................................................................. RS solutions (b) depth profiles 208
Diffision of 1,2-DCA through Halton till (a) source SS and ..................................................... receptor RS solutions (b) depth profiles 208
Diffision of TCE through Halton till (a) source SS and receptor RS solutions (b) depth profiles .......................... ....... ..................... 208
Diffision of benzene through Halton till (a) source SS and ..................................................... receptor RS solutions (b) depth profiles 209
Diffision of toluene through Halton Till (a) source SS and ................................................... receptor RS solutions @) depth profiles 209
Diffision of ethyl-benzene through Halton till (a) source SS and ..................................................... receptor RS solutions (b) depth profiles 209
Diffision of m&p-Xylenes through Halton till (a) source SS and ..................................................... receptor RS solutions (b) depth profiles 2 10
Diffision of O-Xylene through Halton till (a) source SS and ..................................................... receptor RS solutions (b) depth profiles 210
Concentration of VOCs in solution: monitoring stability of ..................... dissolved DCM, 1 ,ZDCA and TCE in time ................ . . . . . 2 1 I
Concentration of VOCs in solution: monitoring aability of dissolved benzene, toluene and ethyl-benzene in time ................................ 2 1 1
Concentration of VOCs in solution: monitoring stability of ........................................... dissolved m&p-Xylenes and O-Xylene in time 2 1 1
Intnnsic degradation of DCM through Halton till: 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles .................. ... ........................................................................... 212
..* Xlll
Fig. 5.1 5.1 Intnnsic degradation of I ,2-DCA through Halton till: 3 cm compacted plugs; (a) source & receptor solutions @) depth
...................................................................................................... profiles 2 13
Fig. 5.16.1 Intnnsic degradation of TCE through Halton till: 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles ...................................................................................................... 2 14
Fig. 5.17.1 Intrinsic degradation of benzene through Halton till: 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles ...................................................................................................... 2 15
Fig. 5.18.1 Intrinsic degradation of toluene through Halton till: 3 cm compacted plugs; a) source & receptor solutions (b) depth profiles ...................................................................................................... 216
Fig. 5.19.1 Intrinsic degradation of ethyl-benzene through Halton till: 3 cm compacted plugs (a) source & receptor solutions (b) depth profiles ...................................................................................................... 2 17
Fig. 5.20.1 Intnnsic degradation of m&p-xylenes through Halton till: 3 cm compacted plugs; (a) source & receptor solutions @) depth
...................................................................................................... profiles 2 18
Fig. 5.2 1.1 Intrinsic degradation of O-xyiene through Halton till: 3 cm cornpacted plugs; (a) source & receptor solutions (b) depth profiles ............................. .. ....................................................................... 219
Fig. 5.14.2 Intrinsic degradation of DCM through Halton til1: 5. cm compacted plugs; (a) source & receptor solutions (b) depth profiles .................................................................................................... 220
Fig. 5. 14.2.1 Variation of intnnsic degradation parameters: (a) effea of lag penod on DCM impact elaborated for cells U- 1 & IV-2 show in Fig. 5.14.1 (b)(4); (b) effect of lag period and half-lives on DCM
..................... impact elaborated for the ceIl 3- 1 shown in Fig. 5.14.2@)(2) 22 1
Fig. 5.15.2 Intrinsic degradation of 1,2-DCA through Halton till: 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles ...................................................................................................... 222
Fig. S. 16.2 Intrinsic degradation of TCE through Halton till: 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles ............... .. ............................................................................... 223
Fig. 5.17.2 Intrinsic degradation of benzene through Halton till: 5 cm compacted plugs; (a) source & receptor solutions @) depth profiles ................................................................................................. 224
Fig. 5.18.2 Intnnsic degradation of toluene through Halton till: 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles ...................................................................................................... 225
xiv
Fig. 5.19.2 Intrinsic degradation of ethyl-benzene through Haiton tiU: 5 cm compacted plugs; (a) source & receptor solutions (b) depth
...................................................................................................... profiles 226
Fig. 5.20.2 Intrinsic degradation of m&p-Xylenes through Halton till: 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles ...................................................................................................... 227
Fig. 5.2 1.2 Intrinsic degradation of O-Xylene through Halton till; 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiies ...................................................................................................... 228
CHAPTER 6
Fig. 6. 1
Fig. 6 . 2
Fig. 6. 3
Fig. 6 . 4
Fig. 6. 5
Fig. 6. 6
Fig. 6 . 7
Fig. 6. 8
Schematic of hypothetical landfill with leachate collection systems and compacted clay liner .............................................................. 247
Impact of DOC in hypotheticd aquifer: Variation oflag penod: DOC degradation considered only in the waste fill: (a) 2 year lag
............................................................................................. (b) 5 year lag 248
Impact of DOC in the hypothetical aquifer: Variation of half- lives in the waste fill and soil layers: (a) 5 year lag (a) 10 year
............................................................................................................ lag; 249
Impact of DCM in the hypothetical aquifer: Variation of lag period: DCM degradation considered in the waste fill only: (a) 5 year lag (b) 10 year lag; .................................................................. 250
Impact of DCM in the hypothetical aquifer: Variation of half- lives in the waste f i i l and soi1 layers: (a) 5 year Iag (bl 10 year
............................................................................................................ lag; 25 1
Impact of benzene in the hypothetical aquifer: Variation of lag period: Benzene degradation considered in the waste fill only: (a) 10 year lag (1) 50 year lag;.. ................................................................ 252
Impact of benzene on the hypotheticd aquifer: Variation of half- lives in the waste fill and soil layers: (a) 10 year lag (b) 50 year lag; ............................................................................................................ 253
...................... Impact of xylenes contamination in the hypothetical aquifer 254
LIST OF APPENDICES
................................................. APPENDIX 1 . SUPPLEMENT TO CBAPTER 1. 267
APPENDLX 2 - SUPPLEMENT TO CHAPTER 2.... e........... . . ~ * . ~ ~ ~ ~ m ~ . . . a ~ a ~ ~ e ~ . a e œ 2 6 9
....................................... APPENDIX 3 - SUPPLEMENT TO CHAPTER 3 e e 2 7 3
APPENDIX 4 - SUPPLEMENT TO CEAPTER 4 .................. ............~......m.....œ..277
APPENDIX 5 - SZTPPLEMENT TO CHAPTER 5 .................................................. 287
LIST of ABBREVIATIONS and SYMBOLS
a
ATP
B A R F
BTEX
CCL
cfu
1,2-DCA
1 -D
DCM
DOC
Eh
HAB
KVLL
MOE
MSWL
(0)DWO
O. Reg 232198
PH
M U
PLCS
PLP
SLCS
SPME
SRB
TCE
TEX
TRIS -EDTA
annum, year
Adenosine-5-trip hosphate
Biological Activity Reaction test
Benzene, toluene, ethyl-benzen, xylenes
Compacted clay liner
Colony forming units
1,2-DichIoroethane
one-dimensional
Dichloromethane
Dissolved organic carbon
Redox (oxidation-reduction) potential
Heterotrophic aerobic bacteria
Keele Valley Landfill leachate
Ministry of Environment (Ontario, Canada)
Municipal solid waste landfill
(Ontario) Drinking water objective
Ontario Regdation 232/98
- log concentration of H* ions
- log dissociation constant
Relative light units
Primary leachate collection system
Possible log population
Secondary leachate collection system
Solid phase micro-extraction
Sulfate reducing bacteria
Tnchloroet hylene
Toluene, ethy 1-benzene, xylenes
Tns~ydroxymethyl]arnino-methane
ethylenediarninetetraacetic acid
Total organic carbon TOC
USEPA
VC
VF As
vocs
US Environmental Protection Agency
Vinyl chionde
Volatile fatty acids
Volatile organic compounds
Microbial population density
Biochemical oxygen demand
Concentration of contaminant in solution
Concentration of contaminant on soil particles
Solubility of a compound in water
Chernical oxygen demand
Coefficient of hydrodynamic dispersion
Effective diffision coefficient in soil pore water
Diffision coefficient in glass porous disc
Coefficient of mechanical dispersion
Difision coefficient in water
D, = &IR, retarded diffision coefficient
Henry's law constant
Organic carbon content in soil
Hydraulic conductivity in porous medium
Zero order reaction rate constant
First order reaction rate constant
Reaction rate for exponential growth at high substrate
Reaction rate for Monod without growth
Reaction rate for exponential growth at low subarate
Linear sorption coefficient
Half-saturation constant for enzymatic reaction
Organic carbonlwater partitioning coefficient
n-Octanollwater partitioning coeffcient
Half saturation constant for the microbiai growth
Porosity of soil
Vapor pressure
R = 1 + p Kd n, retardation coefficient
Coefficient of determination
Concentration of growth limiting substrate
Rate of infiltration
Time
Groundwater (seepage) velocity
Darcy velocity in cornpacted clay liner
Darcy velocity in SLCS
Darcy velocity in till (natural confining layer)
Darcy velocity in aquifer at dom-gradient edge
Liquid molar volume
Depth
Dispersivity
Dielectric constant
Specific growth rate (Monod equation)
Dipole moment
Maximum specific growth rate
Dry density of the soil
CHAPTER 1 INTRODUCTION
Significant developments and improvements have been made during the last three
decades in the practice of landfiiî design and waste treatment technologies. With growing
awareness about the consequences of poliution and stringent regdatory requirements on
maximum permissible emissions, there is a demand for reliable and safe disposal
facilities. Considerable research has been directed at design of various sophisticated
bamer systems, which would secure against unwanted spreading of contamination into
the surrounding environment. The simplest case of barrier design in landfill engineering
is a blanket leachate collection system over a compacted clay liner. Based on a low
hydraulic conduaivity of the liner and a control of the gradient due to the operation of the
leachate collection system, the pnmary mechanism of contaminant transport through the
liner is diffision. The diffusive characteristics of this engineered barier have been
examined both in the laboratory and in the field (Rowe et al., 1995).
Generally, the rate of diffision, has been deduced from monitoring concentration
of contaminants in controlled experiments (Barone, 1990), and subsequently solving mass
transport equation with coupled processes of sorption and firn order reaction, employing
Fickian diffision (e.g. Rowe et al., 1995; Rowe & Booker, 1999):
where:
n is the effective porosity of the soi1 [-1; C is the concentration of tested contaminant at
depth z and time t wJ]; D = D.+Dmd is coefficient of hydrodynamic dispersion p2l?], v is the seepage or (average linearized) groundwater velocity, and m, is Darcy (or
discharge) velocity [LT']; p is the dry density of soil w-~]; K d is Linear sorption
(partitioning) coefficient K~M'] and A is the first order reaction rate constant [T'].
If the influence of advection is eliminated as in pure diffusion experiments, i.e. v =
0, D is taken as De, effective diffusion coefficient, since mechanical dispersion, D,d = v a.
(a, dispersivity CL]) is negligible due to negligible velocity. For non-reactive a d o r non-
degradable species such as majority of inorganic contaminants, the reaction can also be
neglected (A = O), thus fùrther simp&ng the test procedure and equation 1.1. Employing
the test with appropriate boundary conditions (Rowe et al., 1995) both De and K . could be
deduced, or even lumped in so-called "retarded diffision coefficient", D,, = DJR
(retardation coefficient [-1, R = 1 + pwn). De for a particular contaminant species is
inferred through iteration (matching the experimental data collected for concentration vs.
time and or vs. depth with a cornputer generated theoretical curve) using appropriate
program devised to solve the transport equation 1.1. This difision coefficient is
accordingly used as modeiing parameter in engineering simulations of contaminant
impact for large-scale settings.
There is significant body of iiterature compiled on various aspects of contaminant
transport coupled with sorption and degradation mechanisms in porous media, however
lack of research on degradation of organic contarninants, pariicularly in clayey soil and
natural confining deposits with low hydraulic conductivity is still striking. Perhaps the
conventional belief that such degradation would be marginal and impractical to pursue
given the complexities of the soi1 system, uncenainties about microorganisms and
analytical difficulties deterred researchers fiom investigating. Yet many organic
contarninants generated fiom huge municipal solid waste landfills are inevitably although
slowly difising out of the waste fil1 through the engineered clay liners and confining
deposits towards clean ground water, hence it becomes imperative to explore every
chance for their irreversible reduction. It was the intention of this study to examine the
possibility of biodegradation of selected few organic contaminants under such adverse
conditions in a controlled laboratory experirnent. Three most comrnon volatile fatty acids
(VFAs), (acetate, propionate and butyrate), generated fiom organic waste fermentation
were chosen to represent bulk organic contamination, while eight volatile organic
compounds (VOCs), (BTEX and three chlorinated aliphatic), were selected as
representative of priority "micro-pollutants" found in leachates fiom municipal landfills.
A summary of these chernicals and their physical-chemicai properties is given in the
Table 1.1, and their chernical formulae in Fig. Al . 1, Appendix 1.
Acetate, propionate and butyrate are miscible with water, dissociate extensively at
environmental pH, and are generally considered short-lived contaminants (Mackay et al.,
1992). They rnight, however be discharged fiom landfills at relatively high levels and
although assumed readily biodegradable, very little is known on their fate in compacted
clayey soil. Permissible Iimits for this class of contaminants in drinking water are not
currently available in Ontario regulations.
Volatile organic compounds (VOCs) have been recognized as ubiquitous
pollutants in fuel spills, abandoned gas stations and closed rnilitary bases, discharge
effluents fiom refineries, pharmaceutical and chernical factories, landfill leachates and
migrating contaminant plumes in aquifers. Water solubilities greater than those of other
aliphatic, alicyclic and polycyclic hydrocarbons, make DCM (dichiorornethane), 1,2-
DCA (1,2-dichloroetane), TCE (trichloroethene), and BTEX (benzene, toluene, ethyl-
benzene, m, p & O - qlenes), among the many VOCs alike, more available and mobile in
the subsurface environment. Although volatile and prevalent as atmospheric poilutants,
their evaporation is impeded when released in deep and saturated soils, where they
become subject to advection (Mackay et ai., 1992). The exposure to the selected VOCs
poses high risk to health, particularly because of carcinogenic properties of DCM, 42-
DCA TCE and benzene and toxicity of TEX (USEPA, NPDWR), thus making their
concentrations in drinking water subject to very stringent regulations (See Table 1.2).
Biodegradation of selected VOCs has been studied extensively under vanous
experimental and field conditions and in predorninantly permeable soils and sediments
(Aronson & Howard, 1997; Wiedemeier et al., 1999) however no information was found
regarding their fate in compacted clayey soii under environmental conditions relevant to
their release fiom Iandfills.
1.2 Research objectives
The principal objective of this research study was to examine the potential for
intrinsic degradation of selected group of organic contamînants under dominant diaisive
transport through the laboratory compacted clayey soil. In order to accomplish this goal it
was necessary to:
consider the dominant mechanisms influencing contaminant migration through
porous media including diffusion, sorption and biological reaction associated with
selected contarninants, and identify the relevant parameters to be examined
develop a test methodology including apparatus design and media seleaion,
analytical protocols and monitoring which could be used to examine diffision,
sorption and biodegradation in compacted soil
estimate the first order rates of biodegradation for the selected group of chemicds
in the compacted clayey soi1 examined, while delineating the influence of
diffision and sorption
estimate the sorption and diffision parameters, by performing separate ancillary
batch and short terni tests for the same group of chemicals and soi1 in order to
validate these parameters in the absence of significant biodegradation
examine the potential implications of biodegradation on the impact of selected
organic contarninants migrating under a hypotheticai landfill by the means of
cornputer mode1 using parameters based on the laboratory study
examine the likelihood of dichioromethane degradation in the various media and
subsequently estimate biokinetic degradation parameters.
1.3 Thesis outline
This thesis is divided ïnto seven chapters. Each of them is conceptualized and
written as a self-supporting unit independent of other chapters.
Chapter 2 examines the fate of dichloromethane in leachate. Batch degradation
tests are performed, with both synthetic and Keele Valley LandW leachate from
municipal landml and local clayey soil. Biokinetic parameters for dichloromethane are
pro posed based on noniinear regression analysis.
Chapter 3 focuses on the fate of dichloromethane from synthetic and Keele Valley
Landfill leachate under dominant difisive transport through compacted clayey soil
obtained fiom the two local sites. The details on experimentation concept with
preparation procedures and apparatus set-up are developed and preliminaiy intrinsic
degradation tests are initiated with this project.
Chapter 4 presents the principal experiment devised to simulate intrinsic
degradation of selected organic contarninants through compacted clayey soil. These
contaminants are tested as a major organic component of a synthetic leachate prepared to
resemble real effluent generated in a local municipal landfill. Degradati0.i testing
apparatus is adjusted and analytical procedures expanded in order to test clayey soil and
consequently extract information on the examined coupled processes (in the clayey soil).
Results regarding sorption, difision and degradation of the three volatile fatty acids
(VFAs, acetate, propionate and butyrate) are presented and the conditions of processes
interaction and outcome are discussed.
Chapter 5 provides the relevant aspects and results on intrinsic degradation of
eight volatile organic compounds (VOCs: DCM, 1,2-DCA TCE, BTEX) based on
experiments descnbed in Chapter 4.
Chapter 6 deals with the prediction of the impact of selected organic contarninants
on aquifer under the hypothetical Iandfill. Coefficients of sorption and difision as well as
degradation rates obtained from experimental work are used as parameters in computer
modeling.
Chapter 7 gives an outline of the conclusions for this research as well as
recornmendations for fùture work.
1.4 References
Aronson D and PH Howard, 1997, Anaerobic biodegradation of organic chemicals in
groundwater: A sununary of field data and laboratory midies, Environmental
Science Center Syracuse Researc h Corporation,
hiip://esc. syrres. condAnaero bicRpt. hm
Barone, FS, 1990, Determination of difision and adsorption coefficients for some
contaminants in clayey soil and rock: laboratory determination and field
evaluation, Ph. D. thesis, University of Western Ontario, Faculty of Engineering
Sciences
Danbert, TE, RP Danner, BM Sibul and CC Stebbins, 1995, Physical and
thennodynamic properties of pure chemicals data compilation, Design Institute for
Physical Property Date, Amencan Institute of Chernical Engineers
Fiick, EW, 1985, Industrial solvents handbook, 3d ed., Noyes Publications
Mackay, D, MY Shiu and KC Ma, 1992, Illustrated handbook of physical-chernical
properties and environmental fate for organic chemicals, Volume 1 Monoaromatic
Hydrocarbons, Chlorobenzenes and PCBs; Volume iTI Volatile Organic
Chemicals; Volume N Oxygen, Nitrogen and Sulfur containing Compounds,
CRC Lewis Publisher
National Prirnary Drinking Water Regulations (NPDWRs or primary standards),
USEPA, Office of Ground and Drinking Water,
hip:i,iww. epagov/s~fewarer/mc1. hm!
Ontario Dnnking Water Objectives, revised 1994, Ontario Ministry of Environment,
O Queen's Printer for Ontario
Ontario Regulatioo 232/98, made under the Environmental Protection Act, Exîract fiom
the Ontario Gazetîe, vol. 13 1-22, O Queen's Printer for Ontario
Rowe, RK and JR Booker, 1999 POLLUTE v. 6 . 5, 1-D-pollutant migration through a
nonhomogeneous soil, 1983, 1990, 1994, 1997, 1999. GAEA environmentai
Engineering Ltd.
Rowe, RK, RM Quigley and SR Booker, 1995, Clavev banier ?stems for waste
disposal facilities, E & FN Spon, An hprint of Chapman & Hall
Wiederneier, TH, ES Rifai, CJ Newell and JT Wüson, 1999, Natural attenuation of
fûels and chlorinated solvents in the subsurface, John Wiley & Sons, Inc.
Yaws, CL, 1995, Handbook of transpon property data, Gulf Publishing Company
d n
7 ? ? ? 3 O
b 00 C rn G
. * CI d
Nu2 =? . = -
F m =!z
3 3 2:
3 ". II
3 hi 2 2 o.' 5- 2-
. C
2 - a 0 s : O V1 ., . -1 cr a C a -
x = m m in r, N IT 55
s 2 a ='. C? S -
U
2 d z 3 .r)
0 e ~C 2 1; 5 m G
I -
Table 1. 2 Drinking water objectives for tested organic chernicals
Organic chemicai
d / Dichlommethane (DCM)
Dissolve. Organic Carôon
Ontario Drinking Water
Objectives (ODWO).
I 5.0 not available
MAC
b%LI
USEPA
NPDWRS'
NPDWRs National Primary Dnnking Water Regulations (USEPA) legally enforcable standards that appiy to public water systems
MAC - maximum acceptable concentration (a health-related objective)
IMAC - interim maximum acceptable concentration (a health-related objective)
MCLG - maximum contaminant level goal (a non-enforceable public health goal)
MCL - maximum contaminant level (an enforceable standard)
TT - treatment technique (an enforceable procedure)
A 0 - aesthetic objective (established for non-health related parameters)
NSDWRr National Secondary Drinking Water Regulations (LJSEPA)
non-enforceable guidelines; states rnay choose to adopt hem as enforceable standards
IMAC
[Wm MCLG' MCLsor
Tli
CELAPTER 2 BATCH DICHLOROMETBANE DEGRADAïION TESTS'
2.1 Introduction
Dichloromethane (CH2Clt; also known as DCM or methylene chloride), is a
chlorinated aliphatic hydrocarbon that belongs to one of the most important categories of
industrial chernicals with respect to their production, use, dispersion in the environment,
hazard and population exposure. It has been fiequently detected in surface and
groundwater (Ramamoorthy Br Ramamoorthy, 1997) as well as in the municipal solid
waste landfill leachate (Rowe, 1994).
Dichloromethane @CM) has the highest water solubility among ail of the chloro and
halo-methanes, (13.0-19.4 gLL at 2S°C), and is regarded as highly mobile organic
contaminant (Roy & Griffin, 1985). It is classified by the US EPA as a probable human
carcinogen based on its ability to cause lung and liver cancer in mice (Nat. Tox. Prog.,
1986).
Dichloromethane, like any of the related haiogenated aliphatics is quite susceptible to
volatilization from solution into open space (Dilling, 1977) as well as chernical reaction
such as hydrolysis and oxidation (Vogel et al., 1989). The reported half-life for DCM
hydrolysis varies widely from 1.5 years (Dilling et al., 1975), 704 years (Mabey & Mill,
1978) to a maximum of 4600 years (Edwards et al., 1982). Although initially considered
biologically non-degradable based on negligible oxygen consumption during standard
biochernical oxygen demand (BOD) test (Klecka, 1982), a number of expenmental
studies conducted to date have confirmed that dichloromethane is biodegradable. Past
research has been pnmady focused on aerobic oxidation involving both mixed
unidentified facultatively methylotrophic bacteria and pure bacîerial cultures, such as
strains fiom genera Methylobacterium, MethyIophilus, and H'homicrobitrrn, derived
fiom activated sludge or sewage (Rittmann & McCarty, 1980; Stucki et ai., 198 1; =Ili &
Leisinger, 1985; Kohler-Staub et ai., 1986; Wackett et al., 1992). Dehalogenase, a strictly
i This manuscript is in preparation for pubiishing
dichloromethane inducible enzyme f?om the Theta-class of glutathione-S transferases,
catalyzes nucleophilic displacement of CI from DCM (Kohler-Staub & Leisinger,
1985;La Roche & Leisinger, 1990; Leisinger et al., 1995), yielding formaldehyde,
inorganic chloride and reduced glutathione, analogous to a DCM conversion described in
rat liver cytosol (Ahmed & Anders, 1978).
Markedly less information is available on anaerobic degradation of DCM. Leisinger
(1983) suggested that indigenous bactena were actively rernovhg halogenated
contarninants, including DCM, fiom several polluted aquifers in Gerrnany. Wood et al.,
( 1 98 1 ; 1985) reported a relatively shon half-life of 10- 1 1 days for DCM degradation in
laboratory tests with rnuck-water samples denved fiom contaminated aquifer in Florida.
Gossett (1986) reported DCM oxidation to COÎ in the mixed anaerobic culture originally
seeded from municipal digester sludge. Later studies of Freedman & Gossett (1991) and
Stromeyer et al., (1991) revealed that two separate strictly anaerobic mixed cultures were
able to grow consuming DCM as the sole carbon and energy source. For the both of these
cultures. acetate and methane were the degradation products, suggesting the reactions of
the acetyl-coezyme-A pathway in DCM dissimilation.
Freedman & Gossett (1991) proposed the disproponionation of DCM by mixed
culture under methanogenic conditions. Two modes of degradation were effective, the
principal one was oxidation of DCM ( 1 4 ~ ~ 2 ~ 1 2 ) to 14c02 by so-called DCM-oxidizers
likely belonging to eubacteria. The other mode was DCM ("cH~c~~) fermentation
(partial oxidation) to acetic acid ( i 4 ~ ~ 3 ~ ~ ~ ~ ) carried out by eubactena as well, possibly
by some of the acetogens. It is hypothesized that the methyl C ( 1 4 ~ ~ 3 - ) came directly
from DCM ( 1 4 ~ ~ 2 ~ 1 2 ) , while the carboxyl C (-COO.) came fiom unlabeled CO2 supplied
From the large pool of carbonates in the basal medium and the hydrogen equivdents were
supplied from the oxidation of another mole of ' 4 ~ ~ 2 ~ 1 1 to I4c01. The products of the
DCM degradation were converted to methane, possibly by both COrreducing and
aceticlastic methanogens, however the involvement of methanogens in the direct
breakdown of DCM was niled out (Freedman & Gossett, 1991). Stromeyer et al., (1991)
came to the same conclusion that metbanogenesis was not an obiigatory step in the DCM
degradation. Braus-Stromeyer et al., (1993) sirnplified their original mixed culture
(Stromeyer et al., 1991) to three-component acetogenic mix called culture DM. While
growing on DCM, this culture converted it to acetate, releasing chloride steadily and, to
lesser extent, transient amounts of formate. Two distinct anaerobic microorganisms were
isolated from the culture DM, strain DMA and strain DMB, but neither of them aione,
could degrade or grow on DCM. However, when strain DMA was in the solidified
medium in CO-culture with either MetIranospin'IIuum hungatei or with the strain DMB,
degradation of DCM proceeded as observed eariier with the culture DM. Furthemore,
Braus-Stromeyer et al., (1993) offered some clarification of the anaerobic metabolism of
dichioromethane. By eliminating oxygenation from the strict anaerobic culture, and
reduaion through the negligible chloromethane production, it was speculated that the
dehalogenating mechanism could be a fomd hydrolysis, catalyzed by the strain DMA to
yield free or bound intermediate at the oxidation state of formaldehyde. The next
metabolic step is the oxidation of the intennediate to formate, followed by formate
transformation to acetate. Based on the observed 50% reduction in the specific
radioactivity of ' 4 ~ ~ 2 ~ 1 2 in the rnethyl-C position of 1 4 ~ ~ , ~ ~ ~ * , Braus-Stromeyer et al.,
(1993) concluded that interspecies transfer of formate was prevalent in the conversion of
DCM to acetate by the DCM-fennenting stain DMA and its syntrophic acetogenic
partner, strain DMB. Magli et al. (1995) continued the research on acetogenesis, by
subcultivation of rnixed culture DM with DCM as the sole carbon source. The new fast
growing two-component mixed "culture DC" emerged, consisting of the strain DMB and
a new anaerobic organism, strain DMC. Growth on the selective medium combined with
16s ribosomal DNA anaiysis, revealed that the strain DMB was not a homoacetogen as
believed earlier (Braus-S tromeyer et ai., 1993), but a sulfate-reducing (Desziljovibrio
species) bacterium. The new organism, strain DMC closely related to Desirlfotornaculzïm
orientis and Desulfitobacterium dehalogenans could not be isolated in pure culture on
DCM, however when CO-cultured with one of the three metabolically different parinen
(acetogen, sulfate-reducer or methanogen) it grew, degrading DCM to acetate. Magli et
al., (1995) determined that the new strain DMC alone canied out both dehalogenation and
the acetogenesis from DCM, aithough it depended on an unidentified growth factor
produced by the partner.
These findings are of particular importance for any naturd or man-made fermenting
waste effluents and environrnents, which allow for intrinsic development and syntrophic
association of the mentioned metaboiic microbial groups. It is certain that DCM is
degradable (i.e. femented under anaerobic conditions). It cari even serve as a growth
substrate, since a supplementary electron donor is not required for the culture growth. The
overall DCM transformation seems thermodynamicaily favorable, but most of the energy
is consumed in the dehalogenation sep (hydrolysis) and very iittle is conserved for ceil
reproduction and maintenance (Braus-Stromeyer et al., 1993). In case of an organic waste
effluent such as landfil1 leachate, nch in many fennentable substrates (and potential
electron donors), it is very difficult to assess if a heterotrophic anaerobic population could
grow at the expense of low levels of man-made organic pollutant(s), such as
dichloromethane. As suggested by many researchers (Alexander, 1994), it is likely that, in
a polluted environment, a substrate like dichloromethane rnight be also (or rather) CO-
metabolized ("fortuitously metabolized"), Le. degraded but not used for the growth as a
carbon and energy source by the active microbiai population.
Despite the unknowns about the dehalogenating mechanism eeisinger et al., 1995)
and extreme substrate selectivity of the fastidious DCM-fermenting isolates (Magli et al.,
1995), it is reasonable to hypothesize that a mixed anaerobic population fermenting
organic waste or landfill leachates should be able to acclimate to dichlorornethane and
consume it readily.
In this chapter, the fate of dichloromethane and the kinetics of its rernoval fiom real
and synthetic leachates were exarnined for a range of different test conditions. The batch
tests were performed with the objective of estimating the biokinetic parameters and the
rate of degradation in the expectation that this information may be useful in predicting the
fate of dichloromethane in the polluted environment.
2.2 Materials, methods, theoretical considerations and data analysis
2.2.1 Materials: Keele Vaiiey Landf3I leachate, synthetic leachate and soi1
The leachate used for batch experiments was collected from the Keele Valley
Municipal Landfill site at Maple, Ont., Canada. This leachate (KVLL) was typicdy
anaerobic (Eh < -150 mY), süghtly acidic to neutral (6.3<pHcl.3), contained a variety of
dissolved inorganic species and volatile fatty acids (VFAs) cornrnon to the organic waste
fermentation processes (Table 2.1). The arnounts of acetic (ethanoic), propionic
(propanoic) and butyric (butanoic) acids were equal or less than 3000, 2600 and 480
mg2, respectively, based on the 50% of the times sarnpled. Data compiled on the selected
volatile organic micropollutants (Rowe, 1994) indicated consistent presence of benzene,
toluene and xylenes, as well as relatively high levels of dichloromethane in the early
years of leachate production, but very low levels subsequently. Using BART@ (Biological
Activity Reaction Tests) biodetedors (Cullimore, 1993) the heterotrophic aerobic (HAB)
and sulfate reducing (SRB) bacteria were found to be the dominant groups (Table 2.1).
Other characteristic groups such as iron related bacteria (RB) and slime forming bacteria
(SLYM) usually had a lower count. In sumrnary, this leachate was rich in readily
biodegradable carbon (VFAs) and had a sizable rnicrobial population in an active stage of
fermentation and gas production.
The synthetic leachate was a blend of volatile fatty acids, inorganic nutrients and trace
rnetals, and pH adjusted to 6.8, as shown in Table 2.1 The trace metal solution was made
up using sulfate salts to replace chloride salts in the solution originally proposed by
Kosaric (1988), while still maintaining sulfates in the synthetic Ieachate at levels sirnilar
to those observed in KVLL. This synthetic leachate is considered to be a suitable medium
for the growth and maintenance of the acetogenic, rnethanogenic and sulfidogenic
bacteria, which are engaged in the mineralkation stage of anaerobic degradation. Also,
the mix containing only the three volatile fatty acids (acetic, propionic and butync),
adjusted to pH = 6.8 was used as a possible variation on landfill leachate (Blakey et al.,
1988).
Halton till (fiom the Halton Waste Mangement Site, Southern Ontario, see Rowe et al.
1993) was studied as a representative soil material used for the construction of compacted
clay liners. The mineralogical characteristics of this soil, together with ionic content of
soil pore water and soil bacteria counts are given in Table 2.2.
2.2.2 Methods
160 mL serurn bottles were purged with C h : CO2 gas mixture at 60:40 ratio just
before filling with leachate. Approximately 130 mL was poured into each bottle so that
sufficient headspace (about 30 mL) was available for unhindered biological activity (gas
generation). After filling, the sensm bottles were quickiy capped with butyl rubber septa
and tightened with an aluminum crimp and stored in the dark. Unless othenuise noted,
senim bottles were incubated at a temperature of 24' i 2' C. The change of
dichloromethane concentration was monitored in time. The following describes the five
series of tests that were performed:
1. DCM in Keele Valley Landfill Leachate KVLL)
These tests examined the potential degradation of DCM initially present in the KVLL as
well as DCM added to the KVLL under anaerobic conditions. The leachate batches were
randornly selected so that they represented seasonal variations in leachate composition.
Selected bottles were spiked with known amount of dissolved DCM, sufficient to yield an
initial concentration of 4-7 rng/L. For respiking of leachate, concentrations were increased
up to about 30 m g L to identiQ whether these levels could inhibit degradation.
2. DCM in distilled water lDCMlWater Controls)
These tests were intended to demonstrate XDCM is stable in water solution. Several of
the semm bottles were filled with approximately 130 mL of distilled de-ionized water
(rather than leachate) and subsequently spiked with a known amount of DCM.
3. DCM in distilled water with soil (SoiVWater Controls)
Samples of Halton Till were air-dried under nitrogen atmosphere in a glove box. Soil was
passed through the US. No. 4 sieve (4.75 mm openings), pulverized and 40 g of this soi1
was placed into the semm bottles prior to adding distilled water to give a total (soil +
water) volume of 130 mL. Once the senim bottle had been seaied, the soiVwater mixture
was spiked with a known amount of DCM. Soil was dispersed and allowed to equilibrate
and settle (2-3 hours) before the DCM monitoring began. These tests were performed to
evaluate the extent of DCM degradation when the source of bacteria and nutrients was the
air (NZ) dned soil and DCM was a sole carbon source.
4. Soil mixed with KVLL
Soil was prepared in the sarne manner as described above, prior to combining with - 130
mL KVL leachate. The bottles containing leachate batches without DCM were spiked
with known amount of dissolved DCM. These tests examined the effect of soil on the
DCM degradation in KVLL.
5. DCM in the mthetic media mixed with soil
In this test series the KVLL was replaced by synthetic media which contained either only
organic (VFA-mix) or both organic and inorganic nutrients (synthetic leachate). However,
none of the synthetic mixtures contained bacteria indigenous to red leachate. The
composition of the synthetic leachate is given in the Table 2.1, while the VFA-mix
contained only the three carboxylic acids in de-ionized water adjusted to pH = 6.8. The
soi1 was prepared as descnbed above, mixed with the solutions (at the same soil: liquid
ratio) and then spiked with DCM. These tests examined whether the indigenous soil
population could potentially metabolize DCM in presence of volatile fatty acids as
pnmary carbon and energy source.
Dichloromethane presence and concentration was determined on a Shimadm 9-A
gas chromatograph with a flame ionkation detector. A fused silica wide bore capillary
column (30 rn x 0.53 mm ID SPB-5, 3 mp film thickness, Supelco, Belfonte, PA USA) at
a flow rate of 5 mL of Hdmin, was programmed kom initial 1 min hold at 50°C to final
150°C at 10°Clmin temperature increase rate. 1 p5 aliquots were injected directly into the
injection port using a gas tight glas syrînge.
Possible log population (PLP) [based on most probable number (MPN)] of colony
forming units (cfu) was estimated using the Biological Activity Reaction Test (BART?
(Cullimore, 1993). The test tubes i.e. "BARTs" contained the selective culture media
conducive for the growth of the direrent rnicrobial groups. The tirne lag (or days of
delay, User Manual, 1999), before the occurrence of characteristic reaction patterns may
be used to estimate the PLP for the population of interest. For this research the
heterotrophic aerobic bacteria @hW) and sulfate reducing bacteria (Sm) were selected
and monitored as representative dominant groups of the bacteria capable of fermentation
and degradation of many organic compounds. The activity of viable biornass in the
leachate and soi1 sarnples was also checked by measuring the adenosine-Striphosphate
( ATP) using a LumacJM Biocounter mode1 20 10.
2.2.3 Theoretical considerations
The Monod growth equation is generally used to describe the saturation kinetics
of the single substrate utilkation by the active microbial biornass. This equation is known
in its differential forrn as:
where p is the specific growth rate, B is the population density, pma is the maximum
specific growth rate, S is the concentration of growth limiting substrate and K, is the hdf-
saturation constant for the rnicrobid growth, and t is time. The analogous concept of
saturation catalysis kinetics can be applied for the case where the active bacteria are
capable of utilizing any test compound while growing at the expense of another organic
compound. This is usually expressed using the Michaelis-Menten equation:
where. v,, is the maximum specific reaction rate, and K, is replaced with Km. This
designation is used to accentuate the fact that the biomass catalyses the reaction but does
not Vary and as such funaionally resembles enzymes as postulated in the fundamental
enzyme kinetics reactions (Alexander & Scow, 1989). In order to define population
growth it is necessary to employ another equation in which the microbial population B, is
a function of tirne rather than of the test substrate concentration. The expression generally
applied and validated for many cases (Simkins & Alexander, 1984; Schmidt et al., 1985)
is the logistic growth equation, given here in its differential fom:
where, r is maximum specific growth rate and B., the maximum possible population
density that cm grow in that environment.
The need to use a model with the minimum parameters necessary to describe the
collected data as well as the availability of straightfotward analytical solutions have led to
useful and justifiable simplifications (Simkins & Alexander, 1984; Alexander & Scow.
1989; Schmidt, 1992). The first one is applied to the logistic growth equation. For the
case where B,, greatly exceeds actual population B (B,, » B), the rate d3Ai.t -- rB and
growth becomes exponential. Conversely, if the actual population B approaches the
maximum (B = B,,), microbial growth in Eq. 2.3 becomes negligible. Secondly, it is
possible to sirnplifi the Michaelis-Menten equation in case of extreme ratio of initial
substrate concentration S, to half-saturation constant K,,,. For S, >> K,, the rate dS/dt
becomes linear and constant, following the "zero-order" kinetics. If S, « Km, the rate
dIi dt decreases continuously with the decrease of the substrate, resulting in the "first
order" kinetics relationship.
The expressions for the test substrate utilization by microorganisms that do not
grow at the expense of that test substrate, as presented above (Eqs. 2-1-23), can be
classified in two categories. The first refers to the case where there is no growth of the
population at d l , i.e. B z B,,, and B is treated as a constant in the Michaelis-Menten Eq.
2.2. The three speciai cases of the models defined by Eq. 2.2, also referred in literature
(e.g. Schmidt et al. 1985) as Monod family of models, which are considered to be
relevant to this research, are listed in their integrated forms, as foilows: (It is noted that,
most ofien the datasets on substrate or produa change in time are collected, thus
prompting one to use integrated forms of the model equations because they define either
substrate degradation or product formation curves in tirne.)
1. No growth accornpanied with low concentration of test substrate (First-order,
also referred as pseudo-fkst order):
S = ~ , e - " ' , applicable if S, << Km and B s B,,, where kl = vm&&m .
(Characteristic feature of this equation is a substrate half-life, tl.12 = In 2/kl).
2. No growth accompanied with high concentration of test substrate (Zero-order):
S = S, - k,t , applicable if S, >> Km and B E Bm=, where ko= v,,B, and
3. No growth accompanied with intemediate concentration of the test substrate
(Michaelis-Menten, sometimes cailed Monod without growth):
S K m In - + S + Su = -kt, applicable if B z B, and S, = Km, where k = v,,B, Sc?
The second category includes simplified Michaelis-Menten expressions
incorporating rnicrobiai growth as various functions of time. It has been maintained that
the metabolism of one (test) substrate mediated by an active and large microbial
population growing on non-lirniting levels of another (second) substrate is affected by the
growth on the second substrate (Schmidt et al., 1985). Many of the kinetics models have
been proposed and consequently subjected to critical reviews by microbiologists and
ecologists (Schmidt et al., 1985; Alexander & Scow, 1989; Schmidt, 1992; Alexander,
1994; Robinson, 1998; 1985; 1983). Only two special cases, also known as the Dual
substrate models (Schmidt, 1992), are given here in their integrated forms (Schmidt et al.,
1985):
1. Exponentiai growth accompanied with low concentration of test substrate:
S = ~,e'-~~""'-', applicable if's, c< Km and B << B,,, where ki = v-Bn,,
and
2. Exponential growth accompanied with high concentration of test substrate:
S = S, - k, (e" - 1) 1 r , applicable ifs, » Km and B << B,,, with kh = vmaB0.
In addition to the listed substrate utilization equations, there are rnany othen,
based on more complex biological theories and defined with multi-parameter
mathematicai fundions. The work of Simkins & Alexander (1984), Robinson (1985),
Schmidt et al. (1985), Alexander & Scow (1989) and Schmidt (1992) is recommended to
the interested reader.
2.2.4 Data analysis
Al1 data collected for the DCM disappearance in time were processed using a
nonlinear regression anaiysis with cornrnercially available software GraphPad ~rizm"
Version 2.0 (GraphPad Software Inc., San Diego, CA USA). The best fit was found
through iteration by rninimizing sum of squares of the vertical distances of the data points
from the mode1 curve (sum of squared residuals). In its output, the GraphPadTM provides
the basic statistic (standard error and 95% confidence interval of the estimated
parameters, coefficient of detemination @ as well as information whether the analyzed
dataset deviates From the selected model, based on the random distribution of the
residuals around the predicted curve.
The cornputer program SMGLEQU developed by Sirnkins (personal communication,
1996) was used to fit the data to the integrated Michaelis-Menten (Monod without
growth) equation. This program also employs the Marquardt technique to rninimize the
sums of squared residuals, rearranges the integrated Michaelis-Menten equation to the
general fom f (SJ) = 0, and then solves it numericaily by Newton-Raphson iteration.
2.3 Results and discussion
2.3.1 DCM degradation in the Keele Valley Landfiil leachate (KVLL)
The change of DCM concentration with time in the KVLL is shown in Figures
2.1-2.5. Figure 2.1 shows the disappearance of dichloromethane either found originally in
the KVLL (lefl side) or upon a single addition to the leachate (right side). The effect of
subsequent additions of dichloromethane on its degradation rate for the two selected
leachate batches at two different temperatures is presented in the Figures 2.2 - 2.3. The
results of the temperature impact on the degradation rate are given in Figures 2.4 and 2.5.
Despite some scatter in the data (Figs 2.1 - 2.3), there is a consistently rapid
disappearance of DCM, dropping below 0.35-0.5 mg4 within a period of 20 days or less.
In contrast, the control tests performed for distiiled water spiked with DCM in the Figure
2.7, showed no signifiant change in concentration over a penod in excess of 400 days.
There was no measurable decline of dissolved DCM fiom the solution, and hence no
evidence of sorption on the glass or abiotic degradation. When kept in tightly closed glass
botties, dichloromethane appears to be stable in distilled water for a relatively long period
of time. Cornpanson of the results displayed in the Figures 2.1-2.5 with those in the
Figure 2.7 indicates that the decrease in concentration may be attributed to reactions
taking place in the leachate.
Bacterial counts of the leachate confirmed high aaivity ("aggressivity", B A R T ,
User Manudo, 1999) of both heterotrophic aerobic (HAB) and sulfate reducing (SRB)
bacteria, which ranged at (0.2-6) x 1 o7and at (6, 15 and 150) x lo5 c f u d , respectively.
Part of the variation in count of the colony Fomiing units (ch) can be attnbuted to the
modified counting technique of the moa probable number of cfu ( B A R V s ) , which is
known to have lirnited precision. This variation, panicularly for the SRI3 was, however,
prominent oniy for the separate batches of leachate, supplied at different times of the
year. It is believed that the seasonai fluctuation in quaiity and suspected aeration of the
leachate during collection may have afEected the count. Not withstanding any
shortcomings of the method and the complexity of the medium, the growth of the
microorganisms in the KVLL was not observed in these batch tests. The counts for one
particular batch were quite consistent; for example, the bactena in the batches 4 and 5
were stable at 1 x 1o7and 1.5 x 1 o6 cfU/mL for the HAB and the SRB, respectively during
the 40 day test.
These rnicroorganisms had been cultivated and maintained at the expense of high
levels of readily fermentable organic substrates (-3 - 5 g/L VFAs, as well as entire pool
of uncharacterized dissoived organic carbon) and not at sporadic and low levels of DCM
found in the waste and KVLL. In their lab scale study, LaPat-Polasko et al. (1984), have
already dernonstrated that low levels of DCM can be easier metabolized if acetate is
added as primary substrate. Careful examination of DCM disappearance pattems in the
Figures 2.1, 2.2 and 2.3 confirms this observation. Data points form concave-up to almost
linear patterns, which are typical shapes for the Michaelis-Menten (Monod without
growth), first-order and zero-order equations, ail used for modeling the substrate removal
kinetics without microbial growth (Schmidt et al., 1985; Alexander & Scow, 1989). But,
some of the data patterns ai the early stage of re-spiking in the Figures 2.2 and
panicuiarly the pattems in the Figure 2.4 are visibly concave-down suggesting DCM
decay linked to the microbial growth (Schmidt et al., 1985; Schmidt 1992). For that
reason, and in addition to evaluating the grounds for applicability of the discussed
models, al1 data were fitted to selected decay models linked to microbial growth as well
as to the models without growth.
The collected datasets were fitted first to the Michaelis-Menten (Monod without
growth) mode1 to solve for the parameters & and Km. Once the ratio between the two is
known, it becomes clear whether the simplification toward the first or the zero order
kinetics models could be made at all. The results of the nonlinear regression analyses for
the batches with KVLL are summa~ed in the Table 2.3 and shown in the Fig. 2.1 and
2.2a. (For more details regarding the outlook of the various fit lines, see Figs A2.1, A2.2
and A2.3 in the Appendix 2.)
The zero-order, first-order and two selected growth-related models were aiso
tested by noniinear regression analysis (see Figs 2.2b and 2.3) and the values of the bio-
kinetic parameters and basic statistics are enclosed in the Tables 2.3 and 2.4. The sums of
squares for dl of the tested models are low and almost dl of them generate fit lines with
high coefficients of determination (R). Based on the sirnilar (and high) values of R ~ S
alone (Robinson, 1998), one could
tested models relative to others for
not be definitive about the superiority of any of the
explaining the observed variance. It is recommended
that the model be rejected if the standard errors of the estimated parameters exceed 50 %
of the estimates (Robinson, 1998). It seems that the zero-order model is statistically
superior to the others, because the standard enors (available only with the GraphPad
Primm outputs) for So and ko were rarely higher than 10 % of the estimates, while the
mors for the parameters of the other models, particularly the first-order were between 25-
40 % or more, al1 at the listed high R's.
If guided by the critenon of simplicity (Schmidt, 1992) and encouraged to choose
a model wit h "close-to-linear behavior" (Robinson, 1 W8), the preference should be given
to the models with less parameten provided that they give a reasonable representation of
the data. In an attempt to evaluate the grounds for the use of a mode1 with less
parameters, (Le. the zero or the first order), it was observed that the half-saturation
constants Kms decreased evidently lower than So (2.5 > Km > 0.02, i.e. S, » Km), when the
S, increased fiom -5 or 3 mg/Z to -30 mg/L (7 to 10 times). Thus, if one seeks to
justifiably apply a model simpler than Michaelis-Menten (Monod-without-growth), a
necessary condition for the use of the zero-order rather than the first-order kinetics of
substrate removal is fulfilled. This observation is based on the analysis of the DCM
disappearance cuives only, and at the presence of the other substrate and active mixed
non-growing microbial population. The Michaelis-Menten (Monod-without growth)
model is, however, superior for the w e s when the culture is not acclimated to steady
supply of DCM, i.e. So * Km. For the cases with DCM originally present or added just
once to the leachate, the values of the half-saturation constants Km, are close in magnitude
to the values of S.. M e r the microorgm*sms' exposure to DCM the rate of the reaction
became very fast, and levels as hi& as 30 mg/L were removed within less than 48 hours,
which was rnanifested as high affinity of the involved enzymes for DCM (i.e. So » Km).
This justifies the use of the zero-order model for sorne of the tests, besides the fact that
the zero-order model resulted in good fit lines and satisfactory statistic. However, caution
is required in the application of this model for the prediction of the DCM fate in more
cornplex namral environment. Reported to be valid for many environmental samples
(Alexander, 1994), the zero-order rnodei does not by itseif provide information about the
rnicroorganisms carrying the reaction and it can generate negative values for
concentrations (Bekins et al., 1998).
The first-order model produced acceptable fit lines for dichioromethane
degradation, but the conditions (S, should be I<,) for its use are not filfilled. The
frequent and preferential use of the firstsrder kinetics model has been criticized
(Alexander & Scow, 1989; Bekins et al., 1998) as mere mathematical convenience rather
than a result of appropriate model discrimination, which provides the ground for models
use for the real environmental settings. Two growth-linked rnodels employing
exponential growth also produced satisfactory fit h e s in addition to providing the
information about rate of microbial growth. To validate the accuracy of the parameters for
any of the tested models, the independent estimates should be inferred from the separate
experirnents such as measurement of the population growth. Unfonunately, such an
experiment would require observation of the growth in a strictly controlled lab-scale
landfill reactor dunng a long period of testing. Although the landfill rnicroorganisms have
been studied thoroughly (Barlaz et al., 1989; Senior, 1995), there is very little infonnation
on the biokinetic growth parameters of the population. It was inferred that the growth of
three most prominent groups of landfill bacteria (cellulolytics, acetogens and
rnethanogens) could be fitted (satisfactorily) to exponential growth, upon the analyses of
the data available from Barlaz et al. (1989). In case of the genuine field effluent, such as
KVLL, nch in readily degradable organic chemicals, where the growth of
microorganisms could not be clearly differentiated or characterized in a short period of
observation, it is reasonable to assume that a substrate such as dichIoromethane is CO-
metabolized. The fact that S, becomes » K,,, without evidence of microbial growth,
together with lack of inhibition of DCM (CO-metabolic) degradation at high &s (-30
m g i ) given the ample (1000 to t O0 times DCM) supply of VFAs from leachate, supports
this assumption (Schmidt et al., 1985). Considering that, any of the seleaed models
which are not linked to growth, i.e. both zero-order (S, n Km) and first-order (S, Km)
rnight be used to predict the disappearance of non-growth substrate. If the concentration
of DCM is lower than tested, which is very likely to occur in municipal solid waste
landfill, it is expected that the first order mode1 would be well suited for the prediction
analysis.
Another feature apart from kinetics is the response of the leachate microorganisms
to the subsequent addition of DCM. Batches 4 and 5, originally without detected DCM,
repeatedly received DCM after the previous supply had been depleted. DCM decreased
below detection afler about 4.7 to 7.1 days, and batches were re-spiked and the DCM
decrease monitored again (see Figure 2.2ah). M e r the initial bacterial acclimation to
DCM, noticeable upon the first spiking, there proceeds a rapid decrease in concentration
(at a relatively constant rate) for subsequent additions of DCM. This is consistent with the
reported 6-9 day-lags (Gossett, 1986), which vanished d e r re-spiking, and were followed
by rapid degradation within three days, suggesting that microbial consortia had
acclimated and some inducible enzyme system had been activated. Brunner et al. (1980)
indicated that dehalogenase had to be inducible by substrate, since the enzyme activity
become measurabie d e r the methanol grown cells were switched to DCM,
dibromomethane and diiodomethane feed. Braus-Stromeyer et al. (1993) have also
observed consistent lag phase between 5-15 days (at 30°C) when fiesh medium was
inoculated with growing culture of enriched DCM-utilking bacteria (culture DM). It
appears that the KVL leachate is inhabited with species which exhibit similar response.
Considering the metabolic processes taking place in municipal landfills and leachates, it
is not surprising that the degradation of DCM parallels that obsewed with selected
fermentative and hydrogen utilizing species examined by previous investigators.
The effect of temperature on DCM removal was examined with a few ieachate
batches. Following the subsequent addition of DCM at 24"C, the same semm bonles with
leachate batches 4 and 5 were re-spiked but were placed in an incubation room at 10°C.
This temperature is likely to prevail in the bmier system beneath the Iandfill and in
underlying soil. The DCM disappearance curves for 10°C are shown in Figures 2.2 and
separately in the Figure 2.3. Comparing the results, it is evident that the rate of decrease
in concentration is slower at 10°C but that the general trends are the same.
When leachate batches 7 and 8 were spiked with DCM and incubated both at 24°C
and 10°C different patterns were encountered, as seen in the Figure 2.4. The only
prominent difference between these two batches was the storage age; batch 7, was aored
for 6 days before testing, while batch 8 was less than one day old and was used upon
amval. As would be expected based on previous tests, degradation at 24°C was rapid for
@oth batches 7 and 8) without any lag pior to DCM disappearance. The inductive
mechanism has been virtually engaged fiom the start. The rate of DCM removal from the
corresponding batches incubated at 10°C was much slower and appeared biphasic with
concave down shape. Just for this single cornparison between the two batches, it may be
inferred that at 10°C, DCM-fermenters from the fresh leachate (batch 8) facilitate the
degradation faster than those present in "stale" leachate (batch 7). It is considered that this
lag, long relative to the one at 24°C is a direct consequence of lower incubation
temperature (10°C) in these tests. It is also possible that the certain inhibitory effeds
become amplified in the association with low temperature for the case of these two
leachate batches (7 and 8) subjected to a single addition of DCM. Batches previously
exposed to several additions of DCM at 24°C were not drastically retarded at 10°C (see
Figure 2.3).
The regression analyses for the batches 7 and 8 could not generate satisfactoty fit
lines to the collected datasets as show in the Fig. 2.5. The growth linked models seem to
be fitting better (concave down curves), than the other tested models, yielding relatively
high R' statistic (see Tables 2.3 and 2.4 and Fig. A2.2 Appendix 2), yet producing the
curves which deviate fiom the mode1 based on distribution of the residuals. These tests
were not repeated, thus no observation could be made as to evident suitability of growth
linked rnodels under particular testing conditions. It is noted that the fit lines for the
batches 7 and 8 were poor for the disappearance curves at both 24" and 1OoC, which leads
to the observation that factors other than temperature iduenced the DCM removal
kinetics.
Funher investigation of the inhibitory factors was not undertaken, however this
retardation of DCM removal does deserve more attention.
2.3.2 DCM degradation in a KVLUSoil suspension
The results for the tests performed for a mixîure of soil as a resting suspension
with the KVLL are s h o w in Figure 2.6. AU of the tests, with the batch 1 (originally with
DCM) and with the batches 7 and 9 (spiked) indicated a relatively rapid decrease in
concentration. Tests with batches 7 and 8 were conducted at 10°C as well. The results
From soil-leachate suspensions at 10°C show a distinct lag phase prior to degradation, as
evident from the Figure 2.6. This is sùnilar to the lag periods observed in the tests with
leachate alone for the same batches (see Figure 2.4). Comparing Figures 2.4 and 2.5, it is
evident that the rate is slower in the soil-teachate medium than in leachate alone for the
fresh stock (batch 8) at 24°C. For the stale stock (batch 7) there appears to be an almost
negligible difference between the lag phases in the two media. However, the removal rate
of DCM in the batch 7 ("stale7') was faster than in the batch 8 (fiesh) in the presence of
soil as opposed to the leachate alone. Nonlinear regression analyses indicated that at
240C7 some of the listed models (growth linked) could be potentially suitable for
interpreting DCM removal kinetics, while none of them could produce satisfactory curves
for the 10°C data patterns (see Tables 2.3 and 2.4 and Fig. A2.3 Appendix 2).
Tests perfonned at 10°C provide another example of slow initial degradation.
Figure 2.5 shows a prolonged initial lag phase, followed by the fast removal phase. Apan
from the temperature effect, there is no evident explanation for the diferences in the
DCM removal between the two batches tested. The results of these tests, however, do
indicate that the soil-leachate suspension at 10°C provides a suitable environment for
dichlorornethane degradation.
2.3.3 DCM degradation in SoiWater controls
The tests involved soi1 in distilled de-ionized water with DCM. Independent
equilibnum sorption tests for the low concentration range (< 5 mg/L) indicated that DCM
expenences negligible to slight sorption ont0 the Halton Till with partitionhg
coefficient of 0.05 to 0.6 mUg. Thus, it is likely that the effect of sorption of
dichloromethane ont0 the soil is included in the few initial measurements of DCM
concentration. The results obtained from these tests, as presented in the Figure 2.7, did
not show any significant decrease in concentration over a penod of more than 400 days.
This indicates that DCM was stable in soil/(de-ionized) water resting suspension. The
degradation is negligible under tested conditions where DCM was the only source of C
and the soil the only source of bactena and inorganic nutnents.
2.3.4 DCM degradation in the Synthetic 1eachatdSoil suspensions
This testing scenario utilized essentially microorganism-fiee solutions expected to
be conducive for fermentation when in contact with sparse bacteria detected in the air-
dned Halton till. Using BARTM tests. the total population of soil heterotrophic aerobic
bacteria (HAB) and sulfate reducing bacteria (SRB) were estimated to be about 10'
c h r mL, while ATP readings indicated a low total viable biomass of 0.5 ng/g. The results
of the DCM-synthetic leachate/soil incubations are presented in the Figure 2.8. Anaiysis
of the data for the growth medium utilizing synthetic leachate (mix of the three short
chah volatile fatty acids combined with inorganic nutnents) and (a poor) fit to the first
order kinetics with a half-Iife of 457 days does not suggest strong likelihood of fast DCM
degradation. The sarne holds for the tests perfonned with the synthetic mix composed of
the three VFAs only. The rate of DCM disappearance from this solution is low as well,
resulting in half-life of 379 days. Although the synthetic leachate had additional
ammonium nitrogen and phosphate and an extra 1 mLZ of TMS it appears that these
nutrients do not enhance the rate of DCM degradation. It is unlikely that the synthetic
leachate (with inorganic nutnents) exerts any inhibition on the indigenous soil population
since, in itseif, it cm be used as a growth medium. The presence of these extra nutnents
as much as the soil's own inorganic nutrient content does not affect the rate of DCM
removal. It can be argued that the difference between the estimated hypotheticai rates is
marginal at this stage of incubation. n u s , this could be characterized as a very long lag
phase of the potential DCM rernoval, given that the conditions for fermentation and
anaerobic metabolism exist. The soil in the bottles filled with both media had scattered
black Stains which confirmed the beginning of sulfate reduction. This blackening in itself
cm serve as a proof of the growth of SRB (Postgate, 1979; Cuilimore 1993), which is not
surprising aven the ample levels of sulfates originally present in this soil (-5 g/L in the
pore water) coupled with a generous batch of readily degradable C source (VFAs). In
parallel to the burgeoning of SR., seen by the naked eye, it is expected that the bacteria
from metaboIically faster facultative aerobes would also start growing at the expense of
the available carbon (VFAs). Neither of the two synthetic media was chemically reduced
for these tests, since the media redox-potentiai of + 80 mV is quite suitable for the wide
range of soil bacteria. It was the intention of this test to simulate the response of the
uncontarninated (oxidized) soil with aarved bactena to the "organic pollution" of
fermentable substrates such as VFAs and DCM. Bearing this in mind together with the
kinetics of the microbially mediated reactions in the soil, a 180 day lag (or hypotheticai
half-lives longer than 400 days) observed in these tests are quite redistic for the soil
bacteria with long history of carbon starvation and test induced disturbance due to air
drying. It is very likely that such long lag periods, as created in these experiments, will
persist before any of the DCM degrading mechanisms are activated.
2.4 Conclusions
The batch degradation tests have show that dichloromethane @CM) degrades
rapidly under anaerobic conditions in the municipal solid waste (MSW) leachate supplied
from the Keele Valley Landfill (KVL) site. There were no inhibitory effects observed at
the tested DCM concentrations ranging from (1.6-7.6) mg% up to 34 mg/L. The
indigenous microbial population of this leachate appears to readily acclimate to the
presence of dichloromethane. Upon the subsequent additions of DCM to the leachate, the
rate of degradation uicreased, resulting in steady removal of - 30 mg/L within less than
48 hours. This general finding is consistent with degradation observed in batch tests
conducted using a culture isolated fiom lab scaie anaerobic digester, as reported by
Freedman & Gossett (1991). Control tests involving diailled water spiked with DCM did
not show any evidence of degradation or Ioss of DCM over 300 day period, thus the
removai of DCM in the KVL leachate can be assigned with confidence to the biological
processes.
Reducing the temperature from 24OC to 10°C slowed down the DCM rernoval but
the rate of removal, generally remaineci fast. Batch degradation tests conducted with
(Halton till) mil-leachate resting suspensions indicated simiiar behavior to that observed
with leachate alone and suggested that the presence of the soi1 did not affect DCM
degradation.
Nonlinear regression analyses of the collected datasets indicated that Michaelis-
Menten (Monod without growth) equation is generally most appropnate for the modeling
and prediction of the DCM degradation rate. For the cases where the microorganisms are
not entirely acclimated to DCM and initial concentrations 1.6 - 7 mg/L, the best estimate
of biokinetic parameters are: K. = 0.85 - 8 mg/L (Le. Km -- S,) and v,, = 0.5 - 1.4 mgiL
day. At higher DCM concentrations (and subsequent exposure, when the microorganisms
becorne acclimated), v,, (max specific rate of reaction) increases to 20- 4 mg/L per day,
while the half saturation constant decreases to 1.2 to 0.02 mg/L. For such case the zero-
order kinetics can be potentially used for prediction of the DCM fate in anaerobic
environment as well. The use of the first-order kinetic could potentially be appropnate for
the concentrations lower than those examined in this study. Apart from being justified for
a particular case, this frequently preferred mode1 should, however, be used with caution.
If the bias due to approximation (and sometimes necessary mathematical convenience)
results in significant over-prediction of the removal (i-e. higher degradation rate, or less
DCM remaining in the system) relative to the other models, then the use of the first-order
kinetics should be avoided.
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symposium on microbial growth on Ci compounds, San Diego, US4 27 Aug-1
Sept.
Miigli, A, FA Fainey and T Leisinger, 1995, Acetogenesis fiom dichloromethane by a
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Schmidt, SK, S Simkins and M Alexander, 1985, Models for the kinetics of
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Senior, E, 1995, (ed.), Microbiolow of landfill sites, Second Edition, Lewis Publishers
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Table 2.1. Composition of Media Used for the Batch Degradation Tests
Keele Valley Landfili Leachatc WL)'
Component
Generd Indicaton COD (Chemicai Ovgen Demand) BO& (E3iochemicai Ovgcn Demand) TSS (Total Suspended Solids) VSS (Volatile Suspended Solids) DOC @issolved Organic Carbon) ORP (Oxidation-Reduction Potentid), [mv Alkalinity (as CaCO,) Temperature, ["Cl PH (-1 (-log H3
O rganic Compounds Acetate CH3Cûû- Propionate CH3CH2Cûû- Butyate CH3 CH2CH2COO' Phenolics. total Total Organic Nitrogen (Kjeldai) Vinyl chloride Chlorocthane Dichloromcthane DCM 1 . 1 -Dichloroethane ( 1.1 -DCA) cis- \ .2-Dichioroethene 1.1.1 -Trichloroethane 1 A-Dichlorobcnzene BTEX
Inorganic Anions Bicarbonate HCQ' Chloride Cl- Sulphate SO," NitntcsMitrites Nitrogen
Inorganic Cations Ammonium Mi.,- Calcium ~ a " Iron. total Fe Magnesiun ~ g ' Phosphorus, total Potassium K" Sodium Na*
Bacteria (du/mL) and ATP (agig; ng/mL)
H M (Heterotrophic Aerobic Bacteria) SRI3 (Sulphate Reducing Bacteria) ATP (Adenosine-Tri-Phosphate)
Synthetic LeachatcL
Component
Generai Indicators COD TSSNSS 0R.P [mVI Alkalinity (as CaC03) Temperature, ["Cl PH [-1 (-log Hl
Volatile Fatty Acids Wh)
Acetic Propionic Butyic
1norga.uk Nutrienb
Cas04 &CO3 K 2 r n Na?IC@ W C 0 3 (Ntt)2S04 (Trace Metai Solution) NaOH (pH adjustrnent)
Trace Metal Solution
Ai(S04)3 x 16 H20 Cos04 x 5 H20 CUSO~ x 5 H20 FeS04 .u 7 HzO Ha03 MnS04 x H20 (NKr)sMot O.,S x 4 Hz0 NiSOs x 6 HzO ZnSOJ x 7 H.,O 96% concentrated H2S04 Distilleci water
search Centre; rmodined from Kosaric,
Table 2.2.Characteristics of the Soü malton Tü1) Used in the Batch Degradation Tests
Pararne ter
Minerai Specicr (%) ' clay minerais q- totaI carbonates feldspar organic matter
Clay Minerils ( % ) '
mavimum dry density (M& relative density (-) optimum water content (%) water content (VO) plastic limit (Yo) organic carbon content C/, ) (%)
Iaorganic Ions in the Pore Water [mg/L] at -1 5 % moisture content
Bicarbonate HC03- Chioride CI' Nitrate NO3' Suiphate SOI" Calcium ca2+ Magnesiun ~g Potassium K" Sodium Na' PH [-1
Bacteria (cfdmL) and ATP (nglg)
at -15 % moisture content HAB (Heterotrophic Aerobic Bacteria) SRB (Sulp hate Reducing Bacteria) ATP (Adenosine-Tri-Phosphate)
at -2. J % moisture content (air dried) HAB (Heterotrophic Aerobic Bacteria) SRB (Sulphate Reducing Bacteria) ATP ( Adenosine-Tri-Phosphate)
l data h m Quidey, 199 1. at dqth 5.5 m below surface
Table 2.3
DCM lm@]
r n f d d t d once 2 &C Batch 4 Sa = 5.6 Ba th5 Sa=5.1 Batch 6 S. = 7.6 Balch 7 S. = 5.2 Batch 8 S, = 5.2
once 1 b C Batch 7 Sa = 5.0 k h 8 S. - 5.3
38
Cornparison o f Monod without growth, Zero and First-order models
SSR 0.44 0.09 0.07
3.40 1.45 7.69 4.20 4.92
1.44 1.61
1.01 0.40 14.1 0.02
6.34 4.19
3 -46 8.64
4.16 6.59
15.6 0.61 13.6
57.5 5.07 2.52
22.7 23 3 6.15
9.54 6.24 7.68 6.99
SSR = 5 (yo- - y-)', sum of quared re determination; * significant deviation h m model; '
R2 0.708 0.88ï 0.89 1
0.792 0,889 0.583 0.922 0.877.
0.671' 0.750.
0.595' 0.889'' 0.693 0.97 1
0.92 1 ** 0.760''
0.994 0.987
0.984 0.980
0.942 0.993 1 .O00
0.888 0.988 0.996
0.986 0.986 0.987
0.996 0.988 0.989 0.961
: 1- SSF
SSR 5.52 O. 16 0.47
8 . 4 2.54 24.2 4.20 5-23
23.6 15.1
5.61 8.67 29.1 0.04
7.89 26.5
3.82 7.38
12.9 8.66
1 S. 1 1 .O6 O. 19
49.0 8.38 1.28
9.34 9.13 7.93
2.42 7.4 1 7.96 6.78
SSR 2.43 0.09 0.02
5.12 1.39 9.09 11.1 10.1
25.9 19.1
1.01 t 8.2 37.9 0.12
17.6 32.9
35.3 72.8
56.4 14.0
43.6 1.11 19.5
94.6 47.6 27.6
72.0 59.7 19.3
34.1 27.0 47.1 13.6
Table 2.4
Batch #
D a 1 ln KYLL B 1 S. = 3.0 B 2 S . = 1.6 B 3 S. = 3.0
rddd once 2EC B4 S. = 5.6 B 5 Sa =5.1
B 6 S. = 7.6 B 7 S, = 5.2 6 8 S. = 5.2
once 1 bC B 7 S, = 5.0
B8 S. = 5.3
r n f + S o U 2J*C
B 1 S, =3.0 B 7 S, = 7.4
B8 S. =6 .1 B 9 So = 0.9 DC31+So(l
1 09c B 7 S 0 = 6 . 1
B 8 S, =6.5
Respüud 24*C BI Bonle 1
2" S, = 28.4 3* S. = 30.9 4" 5, = 26.8 5" Sa = 29.0 6@' Sa = 27.3 -fh S, = 27.4
B 5 Boule 1 2* Sa =19.1 3* S. = 28.4 4" S. = 34.0 B 5 Boale 2 2" So =25.O 3* Sa = 31.5 4" S. = 25.8
Rapiked 10'C B 4 Bottlt 1 8& S. = 23.7 9" S, = 32.6 10" S, = 33.7
BSBts 1&2 5" Sa = 27.4 6@' Sa = 32.2 7@' S. = 30.7 8" S. = 28.5
39
Comparison of Monod without growth and growth linked models
Monod without growîh B r B , , S o z K m
- SSR
0.44 0.09 0.07
3.40 1.45 7.69 4.20 4.92
1.44 1.61
1.01 0.40 14.1 0.02
6.34 4- 19
3.46 8.64
4.16 6.59
15.6 0.6 1 13.6
57.5 5.07 2.52
227 25.3 6.15
9.54 6.24 7.68 6.99
- SSR
4.77 0.15 0.24
5.m 1.80 13.1 0.79 0.59
12.0 9.85
5.62 2.73 12.2 0.04
2.12 14.1
6.14 1.35
12.2 13.8
6.16 2.18 0.42
1.77 14.4 6E-4
5.32 6.39 8.78
238 6.30 14.9 4.09
Esp. Growth & low DCM B « B,,- Soi( K,
kl [l/dayl; r [day]
- SSR
233 0.08 0.02
5.09 1.41 8.48 1.38 0.72
12.4 11.4
1.62 4.1 9.82 0.04
16.1 32.8
10.3 1.55
2.92 1 3 4
4.80 0.59 8-19
38.6 18.9 4.92
2.55 3.77 17.4
5.60 6.0 1 2.89 3.02
SSR = 1 (y- - nmi of squared residualrg R' = 1- SSW r(y-)2 -2(yd,-din] coeficient ol detenninatioq * sipificant dwiation h m model; - too few points, not fitted
O 7 14 2 1 28 35 42
Time [days]
e L W C ~ 4 - WICS I , 2 8 ~ 1' spikc
- - Bacdi 4 - 1" spikt - M-M fit linc
Ez.tcti4-Boule 1 spikd 7iimCr@24"C, 3 h@ 1 0 ~ ~ ( 1 0 ~ i n t a l . l )
-- b k h 4 Botîlc 1- M-M fil lin~
BatchS-üatllea1&2 1"spike
- - Bilch 5 - 1' rpikc - M-M fil line
A Baich 5 - Boîllc Ispikcd 4 timcr @ 24°C
-- Bakh 5 Doulc 1- M-M fil l i n ~
A 1 3 r l c f i 5 - B a U l c 2 s p i k c d 4 ~ @ 2 4 * ~
.,... Dach 5 Boitle 2 - M - M fil linc
A Birtch5BdllCsl&Z spikd4iimcs@ IO0C(8timginLol*l)
-- &tch 5 - b~tilts 1&2 @ 10°C - M-M fit lk
Fig.2.2 EFECT of SUBSEQUENT ADDITION of DCM on ITS DEGRADATION RATE :Batches 4 & 5 at 24" and 10°C with fit lines to the Michaelis-Menten (M-M) kinetics
2 1 2 8
Time [days]
- - 1" spikt: Exp. p w i h & low DCM
Biidi 4 - W l c 1 rpikcd 7 rimes @J 24'~. 3 timc @ 10°C (10 iimcs in -1)
- 3' - rpikcs @ 24°C fittcd (o the ZcrCHubcr W c r ; @ lOaC : 8*spikc: Eq. growlh & hi& DCM; 9% 1 0 ~ ~ i k t : ihc Za& kinetica
- - 2*& 4& apikc: Exp. growth & hi& K M ; 3" spikc: the Zerwmh kimtica
..... 2"' & 4"rpikc: E q . p w t h & hi$ DCM; 3* rpikt: chc Ztroorda kllvtia
-- @ 10°C il1 fitfuî to the ihworda kinetics
Fig.2.3 EFFECT of SUBSEQUENT ADDlTlON of DCM on ITS DEGRADATION RATE : Batches 4 & 5 at 24Oand l o O C with the lines fit to Zero-order and Growth-linked models
b . , O 1 I h A . . - -
I 1
O 7 14 2 1 28 3 5 42
Time [days]
Balch 4 - BottJc 1 spikcd 7 tima @ M°C, 3 t h e @ 10°C(lOiimcsinto(rl)
- &ich 4 - the Zkrwxder kindics
--- Baich 4 - the Firduda kirvtia
m . . . . htch 5 - Lhc Fimi-order kine tics
Fig.2.4. EFFECT of SUBSEQUENT ADDlTlON of DCM on ITS DEGRADATION RATE : Batches 4 & 5 at 10°C - Cornparison between the Zero-order and the First-order kinetics
O 14 28 42 56 70 84 98
Time [days]
Fig.2.6. DEGRADATION PATTERN and KINETICS LlNES upon SINGLE ADDlTlON of DCM in SOIL-KVLL SUSPENSIONS -
Batches 1 J.8 & 9 at 24OC and batches 7 &8 at IOOC
Time [days]
Fig.2.7. MONITORING DCM CONCENTUTION in DCM
Tune [days]
Fig. 2.8. DCM DEGRADATION fiom SOIUSYNTHETIC LEACHATE SUSPENSIONS
CHAPTER 3 LABORATORY TESTING OF ANMROBIC DEGRADATION OF
DICHLOROMETHANE ZTNDER DlFFUSIVE TRANSPORT THROUGH CLAY '
3.1 Introduction
Even in well-designed landfills with a low penneability clay barrier on a
composite liner (geomembrane over compacted or intact clay), there is significant
potential for difisive transport of organic compounds present in landfill leachate through
the bamer system. One potentiai contaminant of particular interest is dichlorornethane.
Dichloromethane has been show to readily difise through high-density
polyethylene (HDPE) geomembranes (Rowe et ai., 1995a; 1996b) and through compacted
clay (Rowe & Barone, 1991). However, the work of Gosset (1985), Freedman & Gossett
(1991) and Braus-Stromeyer et ai. (1993) indicate that DCM could be anaerobically
degraded by a mixed bacterial culture. These studies approached degradation from the
perspective of wastewater treatment. Nevertheless, they raise the possibility of anaerobic
degradation of DCM occumng as contarninants migrate through a clay liner and into a
receptor aquifer. The work reported in Chapter 2 provides additional support for this
hypothesis by showing that DCM degrades in MSW leachate. However, prior to this
study there has not been any systematic study of biodegradation of DCM as it diffuses
through clay similar to that used in a liner. Thus, the objective of this chapter is to
examine the potential breakdown of DCM in a senes of laboratory difision experiments.
The use of laboratory diffusion tests to assist in the understanding of contaminant
transport mechanisin and to obtain an estimate of diIfiision and sorption parameters
relevant to landfill applications has been described by a number of investigations (e.g.
Rowe et al., 1985; 1988; Shackefford & Daniel, 199 1; etc.). Work on the diffusion of
organic contaminants through both geomembranes and clay has been surnmarized by
' Significant priions of this chapter were published in the the Journal of Geotechnical and
Geoemironmental Engineering, ASCE, 1997, 123 (12). p 10854095. tiIted 'Anaerobic degradation of
DCM diffusing through cIay" Reprinted with permission from the pubiisher (ASCE), 09. Feb. 200 1
Rowe et al. (1995b). In the present chapter, similar diffusion tests wili be performed, but
in this case, ernphasis wiU be placexi on examining the potential degradation of DCM.
3.2 Test program
Two series of tests were performed. The first series involved exarnining the
potential degradation that could occur using a synthetic leachate (Table 2.1) in contact
with an intact "undisturbed" sarnple of silty clay Sarnia till taken h m a depth of about 9
m (Table 3.1). In these tests, the synthetic leachate provided a source of nutnents,
whereas the "undisturbed" clayey till was the pnmary potential source of bacteria. The
second series of tests involved the use of leachate from KVL (Table 2.1) in contact with a
compacted clayey silt, Halton till (Tables 2.2 and 3.1). In this case, the leachate was both
a source of nutrients and anaerobic bacteria. The Halton till obtained from a depth of
about 6 m was air dried, pulverized to pass the US. No 4 sieve (4.75 mm), rewetted to a
water content of 14% (i.e. 3% wet of optimum), and compacted using standard Proctor
compaction to simulate a clay liner. The compacted clay was sampled using a Shelby tube
and extruded from the Shelby tube into a glass testing cell.
For al1 tests, an attempt was made to maintain a constant Ieachate source
concentration; however, for the KVL tests, the original leachate, sarnpled at a number of
different times, was variable (Table 2.1). To minimize the ef'fea of this variability,
additional DCM was added to the KVL Ieachate to bring the concentration to the desired
target values, to be discussed later. The DCM and nutrients could difise through the clay
plug and into a receptor solution below the clay plug. The receptor solution had been
artificially prepared to have a similar ionic content to the pore water of the clay. The
concentration of DCM in the receptor solution was then rnonitored with time.
Some of these tests were performed with the leachate dùealy on top of the clay
plug, some were performed with a coarse sand layer above the clay plug where it was
hypothesued that the sand would provide a suitable location for bacterial growth, and
some were performed with a layer of sand and UASBR (up-flow anaerobic sludge blanket
reactor) granules above the clay plug. The UASBR granules used for seeding the
anaerobic reactors are an engineered rnicrobiai biornass containing a variety of strict and
facultative anaerobes (Kosaric, 1988). They were used to provide a source of bacteria in a
number of tests, and dowed a cornparison to be made between results when the initial
bacteria population had been enhanced by the addition of UASBR granules.
The background rnicrobial data obtained corn this study are summarized in Table
3.2. Both the possible colony forming units (cfw'g) and ATP concentration for both the
intact Sarnia till and the compacted Halton till were similar.
3.3 Analytical methods
Dichloromethane concentrations were routinely obtained by gas chromatography
using Shimadm GC 9-A equipped with fused silica wide-bore capillary column, SPB-5
(Supelco), 30 m by 0.53 mm [D, 3 pz film thickness and flame ionkation deteaor (FID).
Separations were performed at 5 m h i n He, colurnn T = 50°C hold 1 min, to T = 120°C
hold 3 min at 10°/min. Aqueous samples of 1 pL were directïy injected with method
detection limit (MDL) of 0.35 - 0.8 mg/L DCM.
Bacteriological studies carried out as part of the cunent study focused on bactenal
species present in clay and leachate that are Iikely to participate in biodegradation. An
indication of the viable bacterial population was obtained by evaluating the levels of
adenosine-Striphosphate (ATP) in the samples using a LUMAC-3M Biocounter mode1
2010 and assay reagents according to the protocol recornmended by the device
manufacturer. The size of the bactenal population of sulfate-reducing, slirne forming, and
iron-related bactena was estimated based on the possible log population of colony
foming units per milliliter, using biological activity reaction test (BARTM) biodetectors
(Cullimore, 1993).
3.4 Diffusion and sorption
The two clays used in these tests have been extensively studied. The Halton till
was taken from the site of the Halton landfill and is the same clay that is used to constnia
the compacted clay liner (see Rowe et al., 1996a). Rowe & Barone (1991) performed a
series of d a s i o n tests using a simple DCM solution and Halton till with negligible
degradation over the time penod of the test. Using the finite mass technique described by
Rowe et al. (1 995b) and cornputer program POLLUTE (Rowe & Booker, 1991; a bnef
outline of the procedure is given in the A3.2, Appendix 3), they deduced both difision
coefficient (D = 8 x IO''* ' 8.5 x 10'1° m21s) and pditioning coefficient Kd = (1.5 - 1.2)
cm3% for DCM directly fiom the difision test. Rowe et al., (1 9939, discuss the reasons
why both parameters can be obtained frorn a single test. Based on the measured organic
carbon content f, = (0.14 -0.45) %, a value of K, in the (1070 - 267) range was
deduced (K, = K d If=). As an independent check on the diffusion tests, a series of batch
tests was also perfomed and the results surnmarized in the Fig. 3.1. These tests provide
good confirmation of the range of Kd values deduced fiom the difision tests. Based on
these results, the values of diffusion coefficient and partitionhg coefficient considered to 3 be most appropnate for modeling were taken to be D = 8 x 10-'O m'h and K d = 1.5 cm tg.
Test of the Sarnia till (Rowe et al., 1995b) indicated that it has very similar
diffusion propenies to the Halton till and a slightly higher value off, (0.5 %). The
difision and partitioning coefficients considered to be most appropriate for modeling the
Sarnia till were taken to be D = 8 x 14" m2h and d d = 1.6 cm3@.
Batch sorption tests similar to those perfomed with DCM and Halton till (Rowe
& Barone, 199 1) were conducted using UASBR granules and a DCM-water solution. For
the range of DCM concentrations exarnined (3-10 mg/L), sorption was negligible,
resulting in Kd between 0.05 and 0.3 cm31g (cm3&), with a best fit of 0.1 cm3& The
UASBR granules are an aggregated form of various facultative bacteria that are protected
by layers of inorganic matenals. They become very active when a readily degradable
carbon source [such as the volatile fatty acids V A S ) in landfill leachate] is available.
Under these conditions, sorption of DCM is of secondary, if'any, importance compared to
DCM cometabolism and degradation. Modeling that included sorption in the granules
gave results that are not signiticantly dierent from those obtained ignoring sorption.
Consequently, sorption on the UASBR granules wili not be examined any funher in this
chapter.
3.5 DCM diffusion-degradation tests with synthetic leachate
3.5.1 Methodology
The tests were performed in single-piece giass diffusion-degradation cells similar
to that shown schematically in Fig 3.2 and with dimensions given in Table 3.3. (For more
details see Rowe et al., 1994). Each ceIl had an imer constriction around the penmeter
that was used to support the porous glass disk, which prevented sloughing of soil into the
receptor solution (see Fig. 3.2). These cells were designed to permit one-dimensional
diffisive transport f?om the source solution downward through the clay layer and into the
receptor solution with a ionic content seiected to simulate the pore water content of the
clay.
The preparation procedure for these tests involved the following aeps:
1. Cernent the bottom lid to the glass ce11 with two-part epoxy resin type 22 16B/A
(3M, St. Paul, MM.) with a 7-day cure.
2. Extrude the clay plug into the glass ce11 over the previously inserted porous glass
disk that served to prevent sloughing of the soil into the receptor solution. The
clay plug fit tightly into the glas cell.
3. Fil1 the receptor solution compartment below the clay sample with the synthetic
background pore water solution and then pour the same solution into the source
solution compartment above the clay sample.
4. Cernent the top lid to the glas ce11 and allow 7 days for hardening.
M e r the 7-day cure period, the simulated pore water was rernoved fiom the source
solution and the cells were c o ~ e c t e d to the feed network and the synthetic leachate
influent tank. The source fluid cornpartment was filied with shulated leachate. The flow
of leachate was maintained at a constant rate of 9 Uday. The tests were conducted in a
dark fume cupboard at a temperature 24 i 2OC.
To mimic leachate composition as weU as to maintain a constant level of DCM in
the source solution, the synthetic leachate as given in Table 2.1 was prepared by mixing
the inorganic constituents with distilled water, and then purging the fluid with a gas
mixture of C&:C02 at a 60:40 ratio prior to injection of volatile fats, acids and DCM at
the desired concentration and adjustment of the pH to the desired value of between 5.9
and 6.3 (with NaOH).
3.5.2 Test results and discussion
The results of the difision-degradation tests are presented in Figs. 3.3 and 3.4.
The average concentration of DCM in source solutions with synthetic leachate in both
cells 1 and 2 was maintained at about 1.9 mgX.'
Cornputer simulations were perfonned based on Rowe & Booker (1994) for the
diffision of DCM fkom the source through clay, and into the receptor as shown in Figs.
3.3 and 3.4. it was not practical to mesure diffision coefficient through the very
penneable glas disk; however, it is reasonable to expect that the difision coefficients
through the glass disk are greater than in clay and less than for diffiision in free solution.
Thus, the diffision coefficient Dgd foi the glas disk was taken to be between 8 x IO-'*
m2/s (the value in clay) and 12.6 x 1 0 " O m2/s (the pubtished value for the diffision in free
solution; Yaws, 1995). As shown in Fig. 3.3, the uncertainty regarding the diffusion
coefficient for the glass disk has a negligible effect on the results. A similar conclusion
was reached for al1 tests reported in this chapter and, hence, and intermediate value of D,d
= 1 1 x 10'1° m% was used to present many of the subsequent results to reduce the number
of curves presented.
Exarnining the results shown on Figs. 3.3 and 3.4, it can be seen that DCM closely
followed the predicted diffusion cunte for about fira 135 and 95 days, respectively, for
cells 1 and 2, with the concentration in the receptor approaching that in the source at the
end of this period. During this lag phase, DCM was diffushg into the receptor,
confirrning its stability within the solution at relatively low concentrations. It has been
suggested (Rittmann et al., 1980) that some of the recalcitrant micropollutants (present at
low concentrations) c m be consumed if a population is supported for growth and
maintenance from another, so-called primary, substrate that is present in high
concentrations. Thus, it was anticipated that the synthetic mixture of V F h would be the
preferred substrate over DCM, but once the microbial population becomes established,
DCM would eventually be cometabolized as a secondary substrate.
As DCM (and other species within the simulated leachate) diffised through the
clay, there was a gradual change in clay color h m du11 light gray to a dark black gray.
Also. it was observed that there was a gradual increase in dissolved gas in the receptor
solution afier 80 days for ce11 1 and d e r 90 days for ce11 2 (See Fig. A3.1, Appendix 3).
This appears to have been the result of biological activity developing in the receptor and
is likely to have been initiated by the arrival of nutrients in the receptor by difision from
the synthetic leachate.
Following the development of biological activity in the receptor (e.g. as evidenced
by gas production), there was an evident decrease in the concentration of DCM in the
receptor (Figs. 3.3 and 3.4), despite the faa that the concentration in the source remained
constant. This suggests that the biological activity in the receptor became quite enhanced
after the initial inductive (lag) period of 95 - 135 days, and resulted in a decrease in
receptor concentration of DCM corresponding to an apparent half-life of less than 55 days
and reducing, f i e r some time, to less than 10 days, conservatively assurning the same
half-life in the soi1 and receptor. Although DCM experienced a decrease in concentration
in the receptor, no chloromethane was deteaed. Thus, it would appear that DCM
biodegrades to non-hazardous compounds (e.g. COt and Cl&) through mechanism other
than reduaive dechlorination, similar to that reported by Freedman & Gossen (1 99 1).
The DCM measurements used to infer the degradation rate are based on samples
taken from the receptor, and hence, the inferred half-life based on these data represents
the net outcome of degradation in the system below the source solution (Le., in the clay
and receptor). The rate of DCM degradation in the clay alone cannot be direaly denved
fiom the data because the test methodology did not d o w for the clay sampling. However,
the clay sample was the primary source of microorganisms that could cause degradation,
and hence, it is to be expected that degradation could and did take place in ciay. Support
for this expectation is gained from the measured ATP concentration given in Fig. 3.5,
which indicates a higher level of activity in the clay (1 19.5 ng/g) than in the receptor (7
n g h L ) or source (9 ng/mL), suggesring that the clay did indeed play an important role in
degradation process. This higher ATP concentration in the clay at the end of this test was
more than 50 times the original background value (2.2 ng/g), indicating that the synthetic
leac hate had stimulated a significant increase in biological activity in the clay.
In addition to the 50-fold increase in ATP concentrations, B A R F tests indicated
a consistent 60-fold increase in heterotrophic aerobic bactena (HAB, 1.6 x 106 cPmL
compared to the initial 2.6 x 104 cfu/mL), and 250-fold increase in sulfate-reducing
bacteria (SRB, 6 x 10' cfu/mL compared to initial 2 x 10) cWinL). Given the substantially
increased ATP and bacterial concentration in the clay relative to the receptor, there is
some evidence to suggest that the bulk of the degradation occun in the soi1 and not the
receptor fluid. Thus, one can hypothesize that the half-life in the receptor should be
bounded by that for
1 . degradation, the same as in the porous media (lower bound); and
2. infinity (upper bound; implies no degradation in the receptor).
As previously noted, based on assumption 1, the half-life after the lag pet-iod is about 55
days, reducing to 10 days. This represents the maximum value for the half-life of DCM
under the conditions examined. Results based on assumption 2 (infinite half-life in the
receptor fluid) are also shown in Fig. 3.3 and 3 . 4 and it can be seen that for this limit a
reasonable fit to the experimental data can be obtained with haKlives in the clay plug of
12 - 20 days, reducing to I - 3 days. The fit lines produced with simulated hdf-lives
suggest that most (but not dl) of the degradation did occur in the porous media and not in
the fiee solution of the receptor. Thus, following the lag penod, the likely initial range of
degradation rate in the porous media is 12-55 days reducing to between 1 and 10 days.
Here, the precise rate of degradation is not as important as the fact that degradation does
occur in the porous media. The rate of degradation is relatively fast and even comparable
with what was observed in the independent serum bottle tests discussed in the Chapter 2.
Another impo~ant observation from these tests is that the intact clay sarnple taken
from the waxed core appears to have been the major source of rnicroorganisms capable of
metabolking DCM. It is very unlikely that either the source of the receptor solution
prepared for the test initiaiiy contained any signifiant microbial population that could
successfully carry on biological degradation. For example, there has been no
biodegradation of DCM evident from a number of serum bottle control tests.
3.6 DCM diffusion-degradation test with Kn leachate
3.6.1 Methodology
These tests were performed using the glass diffusion-degradation cells and the
same preparation procedure as previously described for the synthetic leachate tests. The
following provides a brief description of the tests conducted.
Cells 3 and 4 were essentially duplicates where the source leachate was in direct
contact with clay plug.
Cells 5 and 6 were essentially duplicates in which a layer of coarse sand was
placed over the clay plug to simulate a granular protection layer over a clay liner
(see Fig 3.2 and Table 3.3).
Ce11 7 was a repeat of the test conduaed with cells 5 and 6, with a minor
modification to the apparatus and a much longer clay plug.
Cells 8 and 9 were essentially duplicates where a layer of UASBR granules was
placed on the clay plug and covered by the layer of coarse sand (see Fig. 3.2). The
objective was to increase the potential biological activity above the clay.
M e r the cure penod, the cells were co~ec ted to the leachate feed network. The tests
were conducted at 27 i- 2°C.
3.6.2 Test results and discussion
Ce11 3 was terminated after 45 days due to problems resulting from a build up of
excessive gas pressure in the ceii. Cell 4 was terminated normaily afler 153 days. There
were no problems with gas in this test. The concentrations are shown in Fig. 3.6. The
increase in concentration in the receptor of ce11 4 was less than predicted assurning no
degradation (tln = infinite in Fig. 3.6) and remains consistently low, suggesting
degradation. Inspection of the specimen did not show any visual evidence of biological
activity at the bottom of the clay plug after 26 days; however, by 40 days, blackening was
observed at the base of the plug suggesting the presence of SRB. Examination of the clay
at the top, middle, and bottom of the clay plug d e r termination of the test indicated
different amounts of black iron sulfide, providing some evidence of focused bacterial
activity at the top and bottom of the clay plug. This visuai impression is confirmed by the
ATP concentration profile shown in Fig. 3.7. The top, middle, and the bottorn of the clay
have ATP concentrations of 17.7, 6.2 and 10.4 ng/g compared to the original background
value of 1.8 ng/g. This suggests that there has been a sigruficant increase in the viable
microbial activity in the clay that may be attributed to movement of nutrients from
leachate through the clay. At the boaom of the clay, there was a black "interface crust"
with an ATP concentration of 28.6 ng/g, wwhich was mbstantidly greater than in the clay
directly above (Le. 10.4 ng/g). Nutnents from the leachate that obviously did reach the
bottom of the clay and the pronmity of free solution (and water transrnissive glas disc)
could have contributed to the signifiant bacterial activity at the interface "cma" and a
14- 17 times higher count of SRB cfw'g relative to initiai conditions.
The leachate was a source of microbid action with an ATP concentration of 20.2
ng'mL. The receptor showed some evidence of activity with a concentration of 2.5 ng/mL,
but this represents a substantially lower concentration than in the leachate or clay.
Because the source DCM concentration was held relatively constant at about 2.8
mg2 over the 153 days of the test, any significant degradation of DCM that influenced
the concentrations in the receptor rnust have occurred either in the clay or the receptor.
The ATP data indicate that the bulk of the activity took place in the clay and at the
interface between the cIay and receptor.
To provide an indication of the order of magnitude of the hdf-life (tin) that would
be required to explain the low concentrations observed in the receptor, a number of
analyses were performed based on Rowe & Booker (1994). Again, two lirniting cases for
degradation in the receptor fluid were examined (i.e., a degradation rate similar to that in
the porous media as a lower bound and no degradation in the receptor solution as an
upper bound). Fig. 3.6 shows the predicted concentrations in the receptor adopting the
same diffusion coefficient for a DCM of D = 8 x IO-'' m2/s and a partitioning coefficient
Kd = 1.5 cm3&, assurning no degradation in the clay or receptor (i.e., il= = infinite). It is
apparent that this significantly overestimates the concentration, implying that there is
degradation. Assurning the sarne half-life in porous media and receptor, the low
concentrations of DCM in the receptor are consistent with degradation corresponding to
an induction (lag) period of order of 40 - 75 days, during which there is negligible
degradation, followed by a degradation stage with a half-life of about 20 days.
Considering the second limit where it is assumed that there is no degradation in
the receptor solution, the time lag is similar (40 - 75 days), with a subsequent degradation
rate of 8.5 days. Reasonable fits could be obtained for half-:ives of 5 - 10 days. Allowing
for uncertainty relating to the rate of degradation in the receptor, the likely inductiodiag
time is 40 - 75 days and a subsequent half-life for DCM degradation in the porous media
is between 5 and 20 days.
3.6.2.2 Sand over a clay plug: Cells 5,6 and 7
The results for ceils 5 and 6 are presented in Fig.3.8 for an average source
concentration of about 2.7 and 2.9 mg& respectively. The theoretical diffusion profiles
obtained, assuming no degradation, are shown for a DCM diffusion coefficient of D = 8 x
IO"* m?'s and a partition coefficient of Kd = 1.5 cm3/g for the clay as previously
discussed. No independent diffusion test was conduaed for the coarse sand however, it is
known that the tortuosity of the coane sand used in these experiments lies between that of
the clay and unity (Le. fiee solution difision; Rowe & Badv, 1996). Thus, the diffision
coefficient in the coane sand, D,, is bounded by 8 x IO-'' m% and 12.7 x IO-'' m%.
Analyses were performed for these values as well as for an intemediate value of 11 x 10- 10 m2 s and, as was a case for the glas disk, the effect of uncertainty was smdl as show
in Fig. 3.8. For convenience of presentation, only results obtained for D, = 1 1 x 10"' m 2 k
d l be shown in subsequent figures.
The tests were terminated at 45 and 47 days for cells 5 and 6, respectively,
because of gas build-up in the receptor. Examining the results shown in Fig. 3.8, it is
evident that there was diffision of DCM from the source through the sand and clay and
into the receptor. The data points for the two cells were similar. Unfominately, the tests
had to be terminated too early to draw any fim conclusion regarding the degradation of
DCM. However, biological aaivity in the receptor was evident fiom gas formation and
development of a black zone (indicative of SRB) at the bottom of clay plug. There did not
appear to have been signifiant biodegradation of DCM over the 42-day period during
which the results were obtained; this is consistent with the findings from cells 3 and 4.
Ce11 7 was similar to cells 5 and 6; however, the length of clay plug was increased
from 2.8 and 3 cm in cells 5 and 6 to 7.5 cm in ce11 7 (see Table 3.3), to examine the
effect of a longer residence tirne in the clay. The results fiom ce11 7 are shown in Fig. 3.9.
The concentration of DCM in the receptor remained very low despite ample time
for diffision through the clay, as indicated by the theoretical curve for no degradation
shown in Fig. 3.9. In this test, there was no visible blackening of the bottom of the clay
plug and no visible evidence of gas formation in the receptor. This might suggest that the
cornetabolizing of DCM is occurring primarily above the receptor. Because the source
concentration was maintained at a relatively constant level, the degradation rnust be
occurring in the sand and clay plug. The clay plug in this test was more than twice the
thickness of that in cells 5 and 6, and hence, there was a greater potential for degradation
in the clay (prior to reaching the receptor) in ce11 7 than in the other cells. The results
From cells 5 and 6 suggest that difision through the sand and clay plug followed that
predicted with no degradation for the fist 40 days. This suggests that there is an initial
lag period in excess of 40 days. Assuming an initial lag penod of 40 days, a subsequent
half-life of20 days or less is required to maintain the concentration in the receptor at the
low levels observed. If one assumes no degradation in the receptor fluid, a half-life of less
than 17 days is required in the porous media to explain the negligible concentrations
observed in the receptor (see Fig. 3.9.).
Although there is some uncertainty regarding the degradation rate in the clay, the
analysis of reasonable cases described earlier indicates that the half-life in the clay is
likely to be less than 17-20 days. While there is some variability, the important point is
that the half-life is small and of the order of 20 days or less.
3.6.2.3 Sand, g r m l e s and clay: Cells 8 and 9
The results fiom cells 8 and 9 are shown in Figs. 3.10 and 3.11. Ce11 9 had to be
terminated for the same reason as cells 5 and 6 after 52 days. Pnor to this, there had been
excellent agreement between the results from cells 8 and 9. Cell 8 proved to be quite
successful. It did eventually have to be terminated at 147 days due to excessive gas build-
up. The observed reservoù concentrations are show in Fig. 3.10 together with the
predicted concentrations assuming no degradation. It cm be seen that the observed and
predicted concentrations are in good agreement for the first 30 days. At 46 days, the
prediction only slightly exceeds that observed and this is generaüy consistent with the
findings from cells 5 and 6 just prior to their termination.
Afier 46 days, the concentration of DCM did not increase but, rather, remained at
relatively low values, suggesting that there was degradation of DCM. At 99 days, there
were no gas pockets evident in the receptor, however there was evidence of the
development of a black build-up (due to biological activity) at the bottom of the clay
PIW-
The test in ce11 8 was temhated after 147 days. Samples of the clay were taken
from the top, middle, and bottom of the clay plug. The top 8 mm of the clay was
gray/black and indicated the presence of active SRB. The middle of the clay layer was
largely brown, aithough there was a smdl amount of black material. The bonom of the
clay plug had a black iron sulfide crust between the clay and the receptor. When the black
cmst was scraped off, the bottom of clay was brown-gray, which indicated that some
black iron sulfide was present.
The source leachate, sand layer, UASBR granules, the top, middle, and bottom of
the clay, the black "interface cnist", and the receptor solution were dl analyzed for ATP,
and the results are s h o w in Fig. 3.12. The sand layer was found to have a higher
concentration than leachate, but as might be expected, lower than the UABSR granules.
The clay shows levels of ATP through its thickness, which indicates the presence of
significant bacterial adivity. The interface cma exhibited a notably higher level of
bactenal activity than the clay or the underlying receptor. As expected, the ATP values in
the top and bonom of the plug were higher than those in the middle of the clay plug. The
ATP levels in the clay were much higher than the original background value of 1.8 npg.
There is some uncertainty regarding the porosity of the granules, however, as
s h o w in Fig. 3.10, this has no significant effect on the interpretation of the results. A
value of 0.37 was adopted for subsequent analyses. As was the case for the glass plate and
coarse sand, the diffusion coefficient through the granules carmot be readily measured,
but for the reasons already discussed, it is likely to lie in the range 8 x 10-'O rn% and 12.6
x 10"' m'h. The level of uncertainty had no significant effea on the interpretation of the
results, and so an intermediate value of 11 x WC0 m2/s was adopted for the purpose of
cornpanson in this chapter, unless otherwise noted.
Analysis performed assurning a lag time of 40 days and a half-life of 20 days in al1
porous layers and receptors yielded good agreement with the observed data (see Fig.
3.1 1). An alternative andysis that assumed no degradation in the receptor fluid and a half-
life of 17 days, in al soi1 and granular layen, gave an over-prediction of the results as
shown in Fig. 3.11. A shorter haif-life of 8.5 days in the porous layers gave a very good
prediction as shown in Fig. 3.1 1. This is consistent with the eartier findings.
3.7 Summary and conclusions
Tests were conducted to examine the potential degradation of DCM in a synthetic
leachate solution as it difises through a clay plug and into a receptor at 24OC. The
concentration in the receptor followed the predicted behavior based on difision of DCM
through the clay plug in the absence of degradation for the fint 95 to 135 days. However,
&er this period, significant degradation was noted and the concentration in the receptor
solution decreased to negligible values over a period of 100 days following the initial lag
period. It was confined that degradation occurred in this environment with an apparent
half-life of' 12-55 days (&er the initial 95- 135 day lag penod) and dropping to 1- 10 days
at the end of the tests.
Examination of the profiles of the ATP concentration provides an indication of the
viable microbiai population. The profile of ATP showed more than a 50-fold increase in
ATP concentration relative to the initial background values, providing evidence of a
substantial increase in biological activity in the clay plug during the test. The Ievels of
ATP in the receptor were about an order of magnitude lower than in the clay.
Tests performed to examine the movement of DCM fkom KVL leachate through a
clay plug and into a receptor at 27OC indicated that DCM typically followed the predicted
diffusion curve for the first 40 - 70 days but &er this induction (lag) period, there was
evidence of significant degradation of DCM as it diffised through the clay into the
receptor. ATP and BARTTM analyses indicated bacterial action in the clay plug as well as
in the receptor. As in the case of the synthetic leachate tests, the ATP concentrations in
the clay were substantidy higher than in the receptor fluid. The inference of biological
degradation is also supported by the fact that DCM in water could readily d f i s e through
the clay plug, but when in leachate there was almost no breakthrough of DCM in the
longer clay columns. The half-iife with KVL leachate as a source of nutrients and bactena
appears to be about 20 days or less at 27°C.
An examination of ATP profile indicates. that the level of biological activity (as
inferred from ATP concentration) in the intact ~arnia tiil was substantially higher than in
the Halton till. It may be hypothesized that the air drying and compaction of the Halton
till caused severe distress to the anaerobic bacteria in the clay, which inhibited the
response because both clays had similar initial background bacterial populations. In the
test with Sarnia till and synthetic leachate, the highest concentrations of ATP were in the
clay plug (119 ng/g), and the source and >eceptor had much lower but similar
concentrations (9 and 7 ngn). This is attributed to migration (and subsequent growth of
the population) of some bacteria fiom the clay to both the source and receptor solutions.
In the tests with KVL leachate and compacted Halton till, the lower level of
biological activity in the Halton till appean to have also resulted in a lower level of
activity in the underlying receptor. This is considered reasonable because the bacteria in
the receptor would be expected to corne frorn the clay. For these tests, there was a distinct
"layering" of the AIT profile in the clay with the lowest ATP concentration in the center
of the plug. In these tests, the KVL leachate appears to have been both a source of
nutrients and bacteria, with the highest ATP concentration in the clay being at the top of
the clay plug.
In the tests where UASBR granules were used, there was enhanced ATP
concentration in the Ieachate, sand, clay and receptor solution. Furthemore, there was a
more uniform distribution of ATP through the Haiton till plug than with the KVL leachate
alone. However, the difision tests indicate that the presence of granules did not
significantly affect the lag time or the rate of degradation of DCM and that the clay and
KVL leachate alone were sufficient to provide the bacteriai population required to give
rise to degradation of DCM.
These tests provide strong evidence for the potential breakdown of DCM as it
migrates through both intact and compacted clay.
3.8 References
Braus-Stromeyer SA, R Hermann, AM Cook, and T Leisinger, 1993,
Dichloromethane as the sole carbon source for an acetogenic mixed culture and
isolation of a fermentative, dichioromethane-degrading baaerium, Applied and
Environmental Microbiology, 59 (1 1) p 3790-3797
Cullimore, DR, 1993, Practical manuai of groundwater microbiology, Lewis Publishers
Feenstra, S, 1994, Groundwater contamination by chlorinated organic solvents,
Geotechnical News, 11(4), p 41-43
Freedman, DL and JM Gossett, 1991, Biodegradation of dichloromethane and its
utilization as a growth substrate under methanogenic conditions, Applied and
Environmental Microbiology, 57 (10) 2847-2857
Gibbons, RD, D Dollan, H Keough, K O'Leary and R O'Eara, 1992, A cornparison of
chernical constituents in leachate from industrial hazardous waste and municipal
solid waste landfills, Proceedings 5& Annual Madison Waste Conference, p 251-
276
Gossett, SM, 1985, Anaerobic degradation of Ci and C2 chlorinated cornpounds, Report
No AD-AI65 005, Engineering and Service Laboratory, Air Force Engineering
and SeMce Center, Tyndall AFB, na. Kosaric, N, 1988, UWO anaerobic sludge granulation process, Final report, Department
of Chernical and Biochernical Engineering, University of Western Ontario
Rittmann, BE, PL McCarty and PV Roberts, 1980, Trace-organics biodegradation in
aquifer recharge, Ground Water, 18 (3) p 236-243
Rowe, RK, L Hrapovic, N Kosaric and DR Cullimore, 1997, Anaerobic degradation of
DCM diffising through clay, Journal of Geotechnical and Geoenvironmental
Engineering, ASCE, 1997, 123 (12) 1085- 10%
Rowe, RK and K Badv, 1996, Advective-diffisive contaminant migration in unsaturated
sand and gravel, Journal of Geotechnical Engineering, ASCE, 122 (12), p 965-975
Rowe, RK, CJ Caen and C Chan, 1996a, The design and operation of a state-of-art
landfill facility, Proceedings 4" Canadian Society for Civil Engineers, Special
Conference, Edmonton, Nt., May 1996,4, p 179-190
Rowe, RK, L Hrapovic and MD Armstrong, 1996b, Diffusion of organic pollutants
through HDPE geomembrane and composite liners and its influence on
groundwater quaiity, Proceedings, 1' European Geosynthetics Conference,
Maastricht, the Netherlands, p 737-742
Rowe, RK, 1995, Leachate characterization for MSW landfills, Proceedings, sLh International Landfill Symposium, CISA Environmental Sanitary Engineering
Center, Cagliari, Sardinia, Italy, 2 p 327-344
Rowe, RK, L Hrapovic and N Kosaric, 1995a, Diffision of chloride and
dichloromethane through an HDPE geomembrane, Geosynthetics International,
2 0 ) p 507-536
Rowe, RK, RM Quigley and JR Booker, 1995b, Clayey Barrier Systems for Waste
Disposal Facilities, E&FN SPON (Chapman & Hall) London, üK, 390 pages
Rowe, RK and SR Booker, 1994, Prograrn POLLUTE v6, Geotechnical Research
Centre, University of Western Ontario, London, Ont., Canada
Rowe, RK, L Hrapovic, N Kosaric, DR Cullimore and RM Quigley. 1994, A
laboratory investigation into the degradation of dichloromethane, Report prepared
for Intenm Waste Authority Ltd., Oa., 1994, 146 pages
Rowe, RK and FS Barone, 1991, Diffusion tests for chloide and dichloromethane in
Halton till: Halton waste management site, Report prepared for Gartner Lee Ltd.,
Markham, Ont., Canada
Rowe, RK, CJ Caen and FS Barone, 1988, Laboratory determination of diffusion and
distribution coefficients of contaminants using undisturbed clayey soil, Canadian
Geotechnical Journal, 25 (l), p 108- L 18
Rowe, RK, CJ Ceam, SR Booker and VE Crooks, 1985, Pollutant migration through
clayey soils, Proceedings, 1 1' International Conference on Clay Mechanics and
Foundation Engineering, San Francisco, CA., 3, p 1293- 1298
Shakelford, CD and DE Daniel, 1991, Difision in saturated clay II: Results for
compacted clay, Journal of Geotechnical Engineering, ASCE, 1 17 (3), p 485-506
Yaws, CL, 1995, Handbook of transport property data, Gulf Publishing Co., Book
Division, Houston, TX
Table 3. 1 Clay cbaracteristics
Sample depth (rn)
Minerdogy (%)' Quartz Total carbonate FeIdspars Illite Vcrmiculite/smectite Chlorite
Organic c&n content &) (%) ' ~ a ~ r n w n dty density (Mg/rn3) Optimum water content (Yo) Relative density (-) Water content (Yo) Plastic limit (*A)'
Sarnia tiU Haitoa till
1 data from Quigley, (1991)
Table 3. 2 Summary of bacteriologicd background data
Possible population (mg)
Sulfate reducing bacteria 1 2.9 10'
HaIton till
Slime forming bactena Iron related bacteria Total aerobic bacteria
Compaaed sample
1.1 IO^ 1.3 los 1.3 x 106
1.5 x 10' - 1.5 xloS (av - 7.5 x 10')
2.8 x lo4 2.2 x 105
1.5 x 1# - 1.8 x 10' (av - 9.0 x 10')
KVL leachate
Fresh leachate
Dimensions
dl values in (cm)
( s a Fi& 3.2)
Hcight of source solution (H,)
Hciglit of sand (b) Heigkt of granules (s) Height of clay (H,)
Hcight of glas disk (Hd)
Height of receptor solution (H,)
Diamctcr of cc11 (D)
Table 3. 3 Dimensions of diffusion - degradation ceJls
Synthetic lewchwte
I
Synthetic lewchwte
2
KVL lewchwte
3
Lewchwte type and cell number
KVL leuchute
4
KVL lewchwfe
S
KVL fcuchate
6
KVL kachute
7
KVL Leacbate
8
KVL leachate
9
na - not applicable
Averaged and range
concentration of d\a
/ /i .O
J'
/ ' /' / / ,,y / ,
Soil : Solution 1.48g: 29mL
Dissolved Concentration [mg/L]
Figure 3.1 DCM batch sorption test results for Halton Till ( modied fkom Rowe & Barone, 199 1)
g l a s tid cemented with epoxy resin to the ce11
single piece g las ce11
source solution pling ports plugged with reservoir coated rubber septa
<- source solution influent connected to the feed network
coarse sand layer (some tests)
UASBR granules (some tests)
gas accumulation tube plugged with Teflon coated rubber septum
1 , inner constriction dong the ce11 perimeter for disc support
receptor sotution g la s base plate ccrncnted with epoxy resin to the ceil
Fig. 3.2 SCHEMATIC of the LEACHATE DETUSION-DEGRADATION TEST CELL
t,, = infinite
1. CELL I m a s u d d a t a
P - - CELL 1 theory D,= 8~10''~rn~/s
// . - . K,=1.6 cm1lg ~~=12.6xl0"~m'h
- CELL 1 thoory ~ , ~ 8 x 1 0 - ' ~ r n %
Kd=l -6 d l g ~ ~ ~ 8 x 1 ~ " O m h
O 30 60 90 120 150 180 210
Time [days]
Fig. 3.3. DIFFUSION-DEGRADATION TESTS with SYNTHETIC LEACHATE: Cell 1 - Receptor solution; (t,, = infinite implies no degradation)
- - - - - - - - - - - - - F t,=17d
/ I S d - - /
/ 12d - - - /'
/
1% . t,=s d - - / v . 8.5d . - - a .
/ ,.'.'-.. - - *
/ tm= 18 d\a / r CELL 2 mcasurcd data
/ - - CELL 2 theary ~,=8xl0- '~m~ts v - 1.
/ v t,=5 d \. i &= L .6 m'tg D i 1 ~ x 1 0 ' ' ~ m ' ~ ~ b m.
- - . - . . . t, a in pomus media ody -- - t, = h porous media & raxptor
Time [days]
Fig 3.4 DIFFUSION-DEGRADATION TESTS with SYNTHETLC LEACHATE : Cell2 - Receptor solution; ( t,,= intlliite implies no degradation)
SOURCE SOLrnON 92 ng/mL
Fig 3.5 ATP CONCENTRATION PROFILE for Cell2: DBbsion-degradation test with synthetic leachate : t = 230 d
1
RECEPTOR SOLüïION 7.02 ng/rnL
WROUS GLASS DISC
15 -
. . . . . tm = 20 dnys in pomus CELL 3 mtasurcd data laycr & rrceptor
r CELL 4 meusund data -.+- tvr = 8-5 daYs
- ( l e g 4 O , 6 ~ & ï S d ) . * CELL 3 thaxy no dccay in reccpdr
..... CELL 4 thcoq -..-
D, = 8~10-'~mYs Kd=l .5 cm31g
DOd- 1 1 x l 0"4n2~s ? . f, = infinite v
A
! ! I t 0.0 4 1
v
O 30 60 90 120 150
Time [days]
Fig. 3 .6 DIFFUSION - DEGRADATION TESTS with KVL LEACHATE : Cells 3 & 4 - Receptor solution; (t,,= infinite implies no degradation)
LEACHATE 202 nghL (SOURCE SOLUnON)
TOP SOIL LAYER 17.7 nglg
MIDDLE SOiL LAYER 63 ndg - iNTERFACE CRUST BOTTOM SOL LAYER 10.4 ng/g 28.6 ng/g
ROUS G M S Df SC
RECEPTOR SOLUllON 2.5 ng/mL
Fig 3.7 ATP CONCENTRATION PRORLE for Ce11 4 : Dfision-degradation test with KVL leachate : t = 153 days
CELL 5 measund data w CELL 6 rneuured dain
THEORY Dm = 8x1 b%211 K, = 1.5 cmslg
CELL 5 - D, = 12.6x10-tamZ/s Dpd = l lxlO-lOmYs
CELL 6 -- D, = 12.6~1û-~~rn~/s DD = l l ~ l ( r ~ ~ r n ~ k ?
Tirne [days]
Fig 3.8 DEFUSION - DEGRADATION TESTS with KVL LEACHATE : Cells 5 & 6 - Receptor solution; (t,, = infinite implies no degradation)
0.20 n CELL 7 measund data s CELL 7 thtory E u D, = 8x1 0-'0rn2/s K,= 1.5 cmslg
: L q = 40 dir 0 L . . . . . t, = 20 days in por. media & mcptor ?
E t, = 17 d in pot. media no decay in a t, = variable
Time [days]
Fig. 3.9 DIFFUSION - DEGRADATION TEST with KVL LEACHATE : Ce11 7 - Receptor solution; (t,, = infinite implies no degradation; t,, = variable, see legend for different cases)
0 CELL 8 mcasud data v CELL 9 measund data
Time [days]
Fig. 3.10 DrFFüSION - DEGRADATION TEST with KVL LEACHATE : Cells 8 & 9 - Effea of porosity; (t,, = intinite implies no degradation)
CELL 8 rncasurtd data CELL 8 ttieary : Da= 8 ~ 1 0 - ~ ~ r n ~ l s Kd = 1.5 crn31g
D, - D, = D, = 1 1x1 o - ' * ~ ' I s
Time [days]
Fig. 3.1 1 DIFFUSION - DEGRADATION TESTS with KVL LEACHATE : Cells 8 - Receptor solution; (t,, = infinite implies no degradation)
LEACHATE 63.7 ng/mL (SOURCE SOLüiiON)
SAND MKED WTH GRANULES 102.4 UAS B R GRANULES 1 57 nglg
TOP SOL LAYER 18.7 ng/g 1 MiD.SO[L LAYER 17.6 nglg BOTTOM SOiL F
[NTERFACE CRUST 24.3 nglg
1 'POROUS GLASS DISC 1 RECEPTOR
Fig. 3.12. ATP CONCENTRATION PROFILE for Ce1 8 : Diffusion-degradation test with KVL leachate : t = 147 d
CEAPTER 4 INTRINSIC DEGRADATION OF VOLATEE FA'ITY ACIDS IN
LABORATORY COMPACTED CLAYEY SOL'
4.1 Introduction
The work presented in this chapter is focused on the intrinsic degradation of
organic contaminants in compacted clayey soil. Biodegradation of chernicals selected as
representative of organic contaminants released in MSWL leachates (three volatile fatty
acids and eight volatile organic chemicals, see Table 1.1) is examined under the
laboratory simulation of dominantly diaisive transport through compacted Halton tiii.
This laboratory study extends beyond the prelirninary tests presented in Chapter 3 (Rowe
et al., 1997) in an attempt to elucidate some aspects of degradation reacîions in the soil
and, most importantly their influence on irreversible removal of contaminants. Its main
objective was to assess whether there would be degradation under adverse, difision-
limited conditions where only the indigenous bacteriai population in the soi1 would be
available to initiate and carry out the contaminant breakdown.
Long-term diffusion experiments (see 5 3.2 and Fig. 3.2, Chapter 3) were
performed to assess whether changes in soi1 and contaminants in the soil pore water could
be observed. Recognizing the complexity of the potential interactions that could occur in
such experimentation, an attempt was made to delineate sorption and difision, as
processes recognized to be dominant and influentid in overail migration. Separate batch
sorption and short-tenn difision tests were performed to provide an independent
assessment of sorption and difision rates (coefficients). These independently infened
coefficients are then used as a basis for estimation of degradation rates in the long-terrn
intnnsic degradation experiment, and ultimately an assessment of degradation rates for
the selected chernical in the teaed system is proposed as weii.
1 Thrs manuscript is in preparation for publishing
This chapter focuses on the results obtained for volatile fatty acids (VFAs). The
results obtained for the volatile organic chemicals (VOCs) from the same experiments are
discussed in Chapter 5.
4.2 Sorption o f VFAs on clayey soils
As acids, volatile fatty acids (VFAs) have the ability to deprotonate and assume
anionic form in the solution. Since the anionic (dissociated) form of acids depends on pH,
the sorption of the organic acids (as much as any charged species) fiom the solution ont0
clay will also be highly pH-dependent. Soils containhg large arnounts of high sunace-
area hydrous Fe and Al oxides and positively charged (Le. acidic, low pH) sites are found
to retain anions (Parf'it, 1978).
ûenerally, if the pH of the system is above pK, (-log dissociation constant for an
acid), the acids are in their anionic (dissociated) form and as such, they stay dissolved in
the solution. Conversely, if the pH of the system is below the pKa, the acids are in their
fiee (nonionic) form (Le. protonated) and tend to be sorbed to much greater extent by soil
or other components of the system. The selected VFAs are weak acids, having pKas of
4.75, 4.78 and 4.8 1 for acetic, propionic and butyric respectively (Schwarzenbach et al.,
1995), therefore for pH greater than these pK,s, each of these acids will be in a
dissociated (anionic) fom in the solution and consequently be less prone to sorption on
the soil. The study of Bingham et al. (1965) revealed that retention maxima of acetate on
H-montmorillonite occurs in the pH range 2 to 6, also indicating that the sorption was an
exchange mechanism with possible chernical bonding. Harter & Ahlrichs (1967)
demonstrated that the acidity of the clay surface could affect the sorption of benzoic acid.
Ernploying IR spectroscopy, it was confirmed based on intensity ratio of caboxylate and
carbonyl IR bands that the clay surface had a pH lower than the pH of the clay bulk
solution. As a consequence of the pH variation as well as soi1 dehydration, the sorption of
both carbonyl (undissociated, COOH) fiee form of the acid as weli as carboxylate
(dissociated, COU) anionic fom was confirmed. Sorption of weak acids and their anions
ont0 such acidic (protonated) sites is, even more enhanced with dry soils since the surface
acidity increases as the water content decreases (Mortland, 1970). The presence of the
cations with increasing polarking power (i.e. those which hydrate strongly and have hi&
charge density) is also known to increase the adsorption of polar organic compounds,
including organic (aliphatic) acids. Apart fiom the few rare studies conducted with pure
and treated clay rninerals and concerned solely with the sorption mechanisms (reviewed
by Mortland, 1970; Bohn et al., 1979 and Mord1 et al. 1982), vimially no information
was found on the rate of sorption of VFAs ont0 natural clayey soil. The only quantitative
evidence on sorption of a VFA on natural soils is provided by von Oepen et al. (1991),
who found that none of the three different types of soil tested [one acidic @H 2.8,
Podzol), one agricultural @H 6.7, Aifisol) and one sublimnic @H 7.1 lake sediment)]
exhibited detectable sorption of acetic acid.
In case of water saturated, clay rich soils such as natural deposits of Halton till
composed of predominately negatively charged inorganic Fraaions, and low in organic
matter, with neutral to alkaline bulk pore solution, it is expected that anionic forms of
VFAs will be prevalent in the pore solution. It is hypothesized that as anions VFAs are
unlikely to sorb ont0 soil, but will rather stay in solution and consequently be subject to
unretarded transpon under the difisive gradient.
In the following paragraphs the potential for sorption of acetic, propionic and
butyric acid is exarnined. Batch equilibnum tests are performed with the objective of
infemng the sorption coefficient for each of these acids. The independent estimates of the
sorption obtained fiom this batch testing are used as parameters when assessing the
difision rates as well as in predictive transport modeling simulations.
4.2.1 Materials and test method
Halton till (from the Halton Waste Management Site, Milton, Ont.), with
characteristics surnmarized in the Table 2.2 was prepared in the manner described in
Chapter 2, 2.2.2.1 [Le. the soil was air-dried, passed through the US. No. 4 sieve (4.75
mm) and pulverized]. This is the same soil and method of preparation that was used for
the laboratory compacted plugs examined in Chapter 3 ( 5 3.2 and 3.4) and is examined
here with the objective of assessing sorption of VFAs. It is noted that the soil was not
sterilized for any of the tests. A p t from being ineffective, chernical sterilization has
been found to change the sorptive properties of the mineral surfaces (von Oepen, 1989).
Previous work with Halton till and with VFAs and DCM (as discussed in 5 2.3.4) has
indicated that there is long degradation lag. Additional triais were repeated with
suspensions made with soil dried under N2 and 3% Na2S x 9H20 - reduced solutions of
VFAs. The results of these trials confïrmed that potential removal of tested VFAs by
carbon aarved and disturbed rnicroorganisms from Halton till is not likely within at least
one month of incubation at high (g/L) concentrations (Fig. A4.1, Appendix 4). Thus, these
trials indicated that sterilization might not be necessary with Halton till for shon term
tests or as long as oligotrophic conditions prevail. Anticipating that biodegradation would
interfere with sorption (or other abiotic reactions), many researchers have opted for some
type of soil sterilization as a measure of precaution (von Oepen et al., 1989; Maraqa et ai.,
1998;). Since the main objective of the overall study was to investigate the potentiai of
intrinsic degradation in this soil, it was decided to avoid creating a fdse environment and
not to alter any of soils properties as might have occurred had it been sterilized. The
approach adopted for the entire project was to register and delineate the processes of
interest as they happen under the simulated "natural" conditions through fiequent
monitoring and a sufficient number of replicate tests.
Working solutions containing a mk of VFAs (Le. acetate, propionate and
butyrate) at concentrations representative of a MSWL leachate were prepared in distilled
de-ionized water and each adjusted to pH 6 with 7M NaOH. It is noted that ail solutions
containing these acids were pH adjusted, thus tuming the acids into their dissociated
forms. The convenient abbreviation "VFAs" is used in the text to refer to this pmicular
group of chemicals, but when dealing with each of them separately, the rems acetate,
propionate and butyrate are used.
The stock and the solution of the highest VFAs concentration was made with - 8
mL acetic: 6 nzL propionic: 1 mL butyric per 1L and was fiinher diluted to - 75%, 50%,
25%, 10% and 5% of its fûii arength. The final solutions were also reduced with 3%
Na2S x 9H20 to bnng Eh to - (- 80 mV). Thus, the objective was to examine the sorption
of VFAs when present in the fermenthg leachate in a state conducive to sulfate reduction
and met hanogenesis with appropriate pH and redox potential.
The batch sorption tests were perforrned relying on the procedure outlined in
standard test method ASTM E 1195-87 (1988).
A mass of - 10.25 g of dried Halton till at 2.4 % moisture content belds - 10 g
oven-dned mass) was placed in a 35 mL heavy duty glass centrifuge tube (Kimble Glass
Inc.). The set of three replicates was then cornpletely filled to the top with a working
solution having particular concentration of selected VFAs. The solution to soil (solids)
ratio (glg) was - 3.6, which is considered representative for the testing of equilibrium
sorption on soi1 like Haiton till and the selected class of chemicals. Each set of triplkate
tubes with soil included two control "bIank" tubes, which did not contain soil, but were
filled with the same VFAs solutions as the tnplicates with the soil. The exact amount of
soil and solutions added to each centrifuge tube was determined gravimetncally using an
analytical balance. The tubes were closed with hole-caps lined with 0.005 "PTWO. 12"
silicone septa (Kimble Glass Inc.), placed on a wrist action shaker (Burrell Corp.
Pittsburgh, PA) for 48 hours at lab temperature (24 * 2 O C ) and then centrifbged at 2000
rpm for 20 minutes. M e r centnfbgation, 0.1 to 0.5 mL aliquots were withdrawn tiom the
supernatant and prepared for aqueous concentration analyses. Amount of WAs removed
from solution after 48-hour equilibration was considered to be sorbed onto Halton till.
Calculation of the sorbed arnounts was done according to the procedure specified in the
ASTM standard E 1995-87, from the difference in initial and final aqueous concentrations
per specified volume of solution.
4.2.2 Theoretical considerations and data analysis
The relationship between the amount of species sorbed ont0 the soil and the
equilibrium concentration of the species in the sod solution at a given temperature is
defined by sorption isotherms. Freundlich and Langmuir equations have been used
extensively in soil science to describe sorption equilibrium rates and plot isotherms.
Originally empiricai and mathematicaily simple, the Freundlich equation hplies that
sorption energy decreases IogarithmicalIy as the fraction of surface covered increases
(Bohn, et al., 1979) and has the form:
defined with Cf = sorbed mass of chernical per unit soil mass, C, = equilibrium
concentration of chernical in soil solution, and empirical constants Kj- and n (ofien
expressed as lln). An important case of the Freundlich isotherm is the linear sorption
isotherm, applicable for n 4, i.e. Cs = KdxCw, with Kd p3m, known as distribution or
linear sorption coefficient.
The Langmuir equation was originally derived for adsorption of gases on solids
under the assumptions that free energy of sorption remains constant and the sorbate
molecules do not interact (Bohn et al., 1979). It has a fom of rectangular hyperbola,
allowing for sorption maximum:
where: S., is maximum arnount of sorbate species that can be (ad)sorbed [units same as
Cs], b is constant related to binding strength [units (cJ'] and C, and C, are as defined
previously. Freundlich, linear and Langmuir sorption isotherms are listed as fundamental
in describing equilibrium (non-kinetic) sorption rates (Bohn et ai., 1979, Momll et ai.,
1982), and as such, they have been used in estimating the extent of sorption of both VFAs
and VOCs on the Halton tu, as wiil be presented in Chapter S. Collected data have been
processed with commercial software GraphPad PrizmTM V. 2.0 using Iinea. and nonlinear
regression analyses as discussed in Chapter 2 (5 2.2.4).
Since each of the selected isotherm equations gives at best an empirical account of
sorption, the preference is given to the linear sorption as one with the least number of
parameters to fit, provided that it gives a satisfactory match to the data.
4.2.3 Analytical measurements
Analysis and quatification of the VFAs (acetic, propionic and butyric) were done
by gas chromatography. A detailed description of the procedure with analytical conditions
is given in tj 43.2.1.
4.2.4 Results and discussion: Batch sorption tests
Results of the batch equilibrium sorption tests for the VFAs and Halton till are
presented in the Fig. 4.1 for la) acetate, (a) propionate and (c) butyrate, respectively.
Despite the evident data scatter, linear sorption coefficients, Kd, (Le. the dopes of the
liner isothems) for each of the acids are very low, suggesting little or no sorption ont0
the Halton till. This observation is in general agreement with no detectable sorption of
acetate found onto the lake sediments having comparable arnount of - 36 % clay and
neutral(7.1) pH, as reported by von Oepen et al. (1 99 1).
Regression analyses indicated that the values of the slopes are valid as such, (i.e.
the data follow straight h e s with the slopes significantly different fkom the dope of a
horizontal line). However, the coefficients of detemination (R) are low (between 0.5 and
0.75) (except for the trivial case where Freundlich sorption degenerates to linear sorption)
and standard error and 95% confidence intenml of the estirnated parameters are higher
than 50%. Fit to the Freundlich and Langmuir equations, (curves not show) produced
worse statistic, with very low d (O. 16 to 0.41) and standard enors occasionally exceeding
100%. This points to the fact that the sorption is very low and as a result the scatter of the
data became significant.
Generally, there was little or no dEerence in the solution concentrations before
and afler shaking, and control tubes indicated very stable levels of VFAs without losses.
The fact that sorbed amounts (values on the ordinate) could be calculated suggests that
some sorption could have occurred on Halton till, however given the very low values it
can be stated that sorption of the VFAs is practicaliy negfigible. For example, the
retardation coefficient R (defined as 1 + pK&, in Eq. 1.1, Chapter 1) of 1.6 for
propionate (Kd = 0.1 1 cm3ig Fig. 4.1b) implies that sorption would hardly attenuate
migration for the Halton till and pH of interest. For the purpose of predicting their impact
on ground water quality, the value of Kd for acetate, propionate and butyrate cm be taken
zero. To the extent that there is a small amount of sorption, this assumption would result
in slight over-prediction of VFA concentrations in the solution.
4.3 Diffusion of volatile fatty acids (VFAs) througb compacted clayey soil
Difision is generally known as mass transport of any species induced by the
presence of chernical potential (gradient) in a single or multi-cornpartment system. It
bears enormous significance in many practicai aspects of reactor design and testing of
materials and its physical and mathematical concepts are discussed in many textbooks and
research publications (e.g. Crank, 1956; Danelewski, 2000; Laskar et al., 1990).
In the field of environmental engineering, diffusion has received particular
attention in the last two decades due to the problems of subsurface contamination of clean
soil or ground-water onginating fiom various point and non-point sources.
Difision is well recognlled as a dominant contaminant (mass) transport
mechanism in low hydraulic conductivity layers such as natural aquitards (clayey
confining beds, preferably, k < 10-~ d s ) or engineered compacted clay liners constructed
in modem landfills (Rowe et al., 1995). Numerous textbooks, handbooks, review-adcles
and research papers deal with various aspects of diffusion of the contarninants in the soil
and offer testing protocols, which could be used in order to estimate the rate of diffusion
(only few are cited for an interested reader: Rowe et al., 1995; Gratwohl 1998; Page
1980; Shackelford 199 1; Barone et al., 1992).
Although found as contarninants in many geologicd settings, diffusion of volatile
fatty acids has not received much attention. The concentration of VFAs in MSWL
leachates rnay be as high as 22 gd (as C, Robinson, 1995), while values of up to 10 g L
(total M A S as C, Blakey & Towler, 1988) have been reported in chalk aquifer pore
water. Other examples include 625 pM (total aliphatic, - 37.5 mg& as acetate) in fuel-
contaminated aquifers (Couareli et al., 1994) and 0.1 - 6pA4 (6 - 360 pg/L) to several
hundred (acetate, or x 60 b @ L ] ) in marine sediments (Nedwell, 1984). McMahon &
Chapeile (1991) gave the values of Fick's law "difhsivities" in sediments for acetate as
D, = 8.7 x 105 m'/a (-2.8 x 1 ~ " m2/s) and formate &= 1.4 x IO-' m'/a ( 4 . 4 x 10'1°
m' s), however no information could be found on their origin or scope of applicability.
Because of conducive conditions for anaerobic degradation in the presence of
volatile fatty acids and their eventual impact on cometabolic removal of other
contarninants under diffisive transport in natural sediments, one of the objective of this
work was to infer an independent estimate of difision coefficients for the few most
frequently detected volatile fatty acids in MSW landfill leachates. In the following section
the short-tenn difision tests and their results, based on Eq. 1.1, are presented for the
difision of acetate, propionate and butyrate through laboratory compacted Halton till.
These data are used to estimate the diffusion coefficients for these acids in an attempt to
delineate bulk difision from biodegradation and to predict impact of contamination fiom
these acids on a large scale.
4.3.1 Materials and methods
Diffusion tests were conducted in a single-piece glass diffusion cell, similar to one
shown in the Fig. 3 -2 (Chapter 3). Muent/effluent inlets and gas accumulation tube were
removed as unnecessary, but the preparation procedure was the same as described in
Chapter 3, 5 3.2 and 3.5.1. The testing scenario involved placing a fuite amount of the
solution containhg a known concentration of one dissociated acid together with a m k of
selected VOCs in the source reservoir solution, while allowing sufficient time for
difision of these chernicals through the compacted Halton tili plug into the receptor
solution. The source solution, (having -7 g 4 acetate, or - 5 g/L propionate, or - 1 g/L
butyrate) was selected to mirnic leachate at the transition between acetogenic to
methanogenic stage, thus its pH and Eh were adjusted to 6 and to -(80 to 100) mV,
respectively. As stressed earlier, this test was intended to give an account of diffusion rate
for the VFAs (tested together with VOCs) under the conditions close to those existing in
natural or in-situ compacted attenuation layers, therefore soi1 sterilization was
intentionally avoided. It was expected, however that a class of chemicals susceptible to
fermentation such as VFAs could start degrading within a retention tirne of a month,
usually required for testing of difision. Fermenting butyrate and propionate could break
down into acetate (Dolfing, 1988; Widdel, 1988), thus only one of the acids was tested at
the time in order to have better control over the eventually generated amounts of acetate.
Anticipating dl of this, most of the difision testing cells available were designed to
create a sharp concentration decline in the source and measurable concentration
breakthrough into the receptor solution within relatively short period of testing. Thus, a
large volume (Le. height H, in the Fig. 3.2, Chapter 3) of the source solution relative to
the volume (i.e. height H,) of the receptor solution and a short soi1 plug, were utilized in
addition to creating a high initial concentration gradient. Based on the previous testing
with Halton till suspension and VFAs mix (see Chapter 2, 9 2.3.2 and Fig. A4.1,
Appendix 4), it was considered likely that the indigenous carbon starved microorganisms
in the compacted plug, would not be able to acclimate fast to high concentration of VFAs
(grL), therefore sufficient time would be lefi for distinct diffusion profiles to develop in
the t esting compartments.
The teaing methodology included fiequent monitoring of the VFA concentrations
in the source and receptor solutions. Particular Gare was taken to observe changes of soi1
plug color and occurrence of odor in the solutions, since these signal the commencement
of VFAs fermentation and consequently prompt test termination. The plug thickness of 2
and 3 cm was chosen as suitable for pore-water extraction (to get a concentration profile
through the sample) while also allowing for fast diffusion. At the time of test termination
the soi1 plug was cut in 2 to 3 slices (layers), each of which was squeezed to provide pore-
water for the chernical analyses.
4.3.2 Analytical measurements
4.3.2.1 Gas chromatography
Analysis of VFAs was routinely done using gas chromatography with Shimadm
GC-9A and with Varian 3400 GC equipped with 8200x SPME Autosarnpler. On both of
the instruments, the chromatographic separations were performed on 15 rn x 0.53 mm (or
0.25 mm) ID, 0.5 p n film, NUKOLW (modified PEG with Ntroterephthalic acid)
capillary columns (Supelco, Bellfonte, PA) and Fm detectors. Column ovens were
programmed for an initial hold of 1 min at 100°C. then, they were rarnped at 1 S0C/min to
180°C and kept at final hold for 2 min. Helium was used as camer gas, at - 20 mumin
with 0.35 mm ID and 5 mumin with 0.25 mm ID capillaries respectively, while nitrogen
served as a make-up gas for the FIDs at - 30 mlimin. Injector/detector block was set to
260°C.
Before the chrornatographic analyses, al1 aqueous samples containing M A S were
acidified (protonated) in 5 mL glass vials with 1% H3PQ4 to yield - pH 2. Al1 quantities
were determined gravimetrically using an analytical balance. VFAs standards were
prepared in the sarne manner, by diluting a stock solution with known VFAs
concentrations in known amount of 1% &PO4. Iso-valeric acid was used as intemal
standard and thus, was added in designated amount to any of the prepared 5 mL acidified
solutions with standards and unknowns. Acidified samples which were analyzed
automatically, were transferred in amounts of 1.2 mL from the 5 mL preparation via1 to 2
mL autosarnpler vials.
VFAs samples eom the soil plugs were obtained directly fiom the soil pore water.
Soi1 cut from a particular layer of the soil plug was squeezed in a stainless steel rnould
under the pressure of 24 W a in the pore squeezing apparatus. Owing to the fact that the
selected VFAs are strongly hydrophilic (polar) and dissociated at systems pH (27) and are
not as volatile as they are implied by their popular narne (see vapor pressures and Henry's
constants listed in Table 1. l), these acids could be physicdy extracted within the
squeezed fluid from the wet soil samples, without signifiant concem regarding losses.
Once collected from the pore squeezing apparatus, the squeezed pore water solution was
treated as any aqueous sarnple containing VFAs.
The chromatographie analyses of VFAs with the Shirnadzu GC-9A were done by
direct manual injections of 0.5 or 1 & of acidified samples or standards. The peaks were
processed with HP 3396A integrator. This instrument was used for the samples with low
concentrations where higher precision was required. The analyses with Varian 3400 GC
suit able for high throughput were automated and employed solid phase micro extraction
(SPME) of VFAs fiom the sarnple ont0 a 75 p n CarboxenTM/PolydimethylsiIoxane
(Supelco, Bellfonte, PA) fiber. The syringe needle with sheathed fiber pierced a 2 mL
autosample via1 septum, immersing the fiber into an acidified VFAs sample for 10
minutes. M e r this time, determined to be sufficient for the extraction of the VFAs ont0
the fiber, the syringe (with the fiber in retracted position) was moved fiom the via1 into
the injection pon of the GC where the fiber was desorbed for 2 min. Although not
designed for this andytical application by the manufacturer, the fiber performed
satisfactorily for the extraction of VFAs at reiatively high concentrations and its use was
continued throughout the project. Peak integration, calibration and quantification were
done by supplementary Varian Star Chromatography software. Generally, both of the
instmments gave very good reproducibility, high precision and recovery of the standards
and unknowns. The detection limit for the acetic acid, recognized as the most difficult to
separate, was 3- 10 @.
4.3.2.2 Bacterial p o p Iation sise und A P content memrements
Possible bacterial population expressed as colony forming units (cfu) was
estimated using the Biological Activity Reaction Test ( B A R F ) developed by Cullimore
(1993). BART" employs a set of distinctive reactions customized to detemiine the
aggressivity and composition of a particular consortium of rnicroorganisms potentially
present in a system. A B A R F polypropylene tube reactor (container) with a selective
crystallized growth medium at the bottom and a floating bal1 that restncts the entry of
oxygen into the reactor, is filled with 15 mL aqueous sample, capped and observed for 1
to 1 5 days. Dunng the incubation period, a series of nutrient and oxygen diffusion fronts
is established in the B A R V tube allowing different microorganisms (if) present in
sample to gradually become active. Usually the occurrence of charactenstic turbidity
accompanied with distinct change of color, odor, gassing and slime fonation signals the
reaction, presence and growth of particular bactena. Two major observations are made
dunng the test: time lag to the reaction and reactions signature pattems. The time lag (Le.
"days of delay to sighting the reaction") is directly linked to the possible log population
(PLP) in the sarnple, (provided as log of colony fonning units i.e. number of cells, cfidmL
of cftu'g in the User Manual), as correlated by Cullimore (1993) based on comparative
studies with other determination methods. The reaction signature pattems gathered during
incubation and checked against coded Reaction Comparator Chart @roycon Bioconcepts
Inc.) are devised to describe the chronological sequence of the observed reactions and
charactenstic features of the dominant group in the rnicrobial consortium (BART, User
Manual, 1999).
Three customized B A R F tests were employed to characterize microbiai
population in Halton till and testing solutions: HAB-BART for heterotrophic aerobic
bacteria (HAB), SRB-BART for sulfate reducing bacteria (SRB) and BIOGAS-BART for
methanogenic bacteria w). HAB-BART with a selective medium for a range of
comrnon waste-water aerobic (and facultative aerobic) heterotrophs contains methylene
blue dye at the beginning of the test. The dye fades away (becornes reduced) as a
consequence of bacterial respiration, which at the same tirne serves as evidence of HAB
activity and growth. SRB-BART works on a sirnilar p~cip le ; it contains ferrous iron and
sulfate in the SRB
mediated reduction,
growth medium, which becomes visibly black under microbially
thus formation of black iron suifide in the reactor tube signals the
presence and growth of SRB. BIOGAS-BART with selective medium for methanogens,
utilizes 60 rnL of sample and is additionally equipped with a light polypropyiene thimble.
Initially the thimble is su* but upon methane generation, it rises to the top of the tube
reactor, signaling active methanogenesis. BART analyses were performed upon test
termination when sufficient amount of sample allowing for adequate number of
replicates. was available for testing.
Liquid samples were initially tested undiluted (Le. in batches of 15 mL) as
suggested in the analysis protocol. The reaaion became very fast at the late stage of
experimentation, making it difficult to monitor signature patterns* thus the samples had to
be progressively diluted with sterile water. The liquid dilution ratio (total volume/sample
volume) ranged from 1.5 or 10 and this was used as a multiplier to assess final count. Soi1
was tested in amount of O. 1 - 0.3 g for HAB- and SRB-BARTs and 1 - 3 g for BIOGAS
BART, by piacing the soi1 directly into the BART tube and adding 15 ml. sterilized
distilled water (Cullimore, personal communication). Both H M - and SRB-BARTs have
been used intensively and the "PLPs" are weil established based on observed time lags
( B A R T User Manual, 1999), thus the population size of the two bactenal genera
targeted could be estimated. BIOGAS-BART is used only to confirm the presence of
methanogenesis, since the quantification procedure is not yet well defined.
Total level of microbial activity in the soi1 and solutions was also checked by
measuring adenosine-5-triphosphate (ATP), a principal energy- stonng nucleotide
(molecule), present only in living cells. A bioluminescence (light emission) essay,
catalyzed by firefly luciferase was perfonned with Optocomp 1 luminometer (MGM
Instniments). The rnethod relies on reaction of ATP with luciferin/luciferase, which
generates light registered by photo-detector and recorded as relative light units F U ) .
The arnount of light generated is proportional to ATP released from bactenal cells. The
procedure for waste-water and sludge samples originally developed by the reagent
supplier Celsis (former Lurnac, Landgraaf, the Netherlands) was employed in this snidy.
A 100 pL aliquot of the liquid sample was pipetted into a aenle disposable cuvette and
placed into the luminometer. The assay was started by automatically injecting 100 4 of
nucleotide releasing agent (mumire of ionic surfactants) to the sample in the cuvette.
M e r 10 sec 100 pL of reconstituted enzyme (luciferine/luciferase) was injected by
another micropump to sample treated in the cuvette. RLU count was recorded and assay
repeated with the same sample, but in another cuvette quenched with 20 jd of appropriate
ATP standard dilution. Amount of ATP was calculated based on difference in RLU
counts and known concentration of quenching standard. Soi1 suspensions were prepared
by homogenizing 1 g of wet soil and 49 mL of TRIS-EDTA buffer in the blender for 1
min and consequently treated as liquid sarnples. Organic carbon content of the soil using
the Modified Walkley-Black Method (Allison, 1965) was aiso checked in parallel to
B A R T and ATP tests
4.3.3 Results and discussion: Diffision tests
The results of the laboratory diffision tests for the VFAs through compacted
Halton till are show in the Figs. 4.2 - 4.4. Diffusion coefficients are obtained through
iteration using program POLLUTE v.6.5 (Rowe & Booker, 1999), which solves mass
transport (i.e. diffision) equation pq. 1.1: (n*X/ût) = n - ~ * ( d ~ / h ' ) , as defined in
Chapter 11. The best estimates of the difision coefficients used in modeling of the test
data are sumrnarited in the Table 4.1.
Despite some data scatter, particuiarly in the source solutions, distinct diffision
profiles for concentration vs. time in the solutions and concentration vs. depth in the soil
pore water were obtained for acetate (Fig. 4.2). propionate (Fig. 4.3) and butyrate (Fig.
4.4). The datasets are in very good agreement with the theoreticai prediction, producing
curves with high & (X.9) for al1 of the inferreci d a s i o n coefficients.
Diffision of acetate and propionate was modeled without sorption (Le. with
Kd = O, as suggested in 8 4.2.4 and in Table 4.1) and without degradation. For the short-
term tests, data collected for al1 of the three profiles are in excellent agreement with the
theoreticai fit-Iines generated with diffision coefficients of 2.5 x 10-'O m2/s ( R ~ = 0.987)
for acetate (Test 1, Fig. 4.2) and 2.0 x 10*1° m2/s (l?? = 0.983) for propionate (Test 1, Fig.
4.3), respectively. Mass balance calculations performed for acetate after 23 days (Test 1)
and propionate d e r 26 days (Test 1) indicate 106% and 97 % mas recovery, as can be
seen in the Figs. A4.2.1 and A4.2.2 in Appendix 4. Furthemore, during the testing
periods of 23 and 26 days, the soil plugs exposed to acetate and propionate retained their
msty oxidized color and clear and odorless solutions. Upon the termination of these tests,
B A R T biodeteaors did not indicate any confirmable change in count of HAB and SRB
relative to untreated soil.
Based on presented evidence it could be concluded that neither acetate not
propionate started degrading to the extent that could be measured by available analytical
techniques, therefore the estirnates of the diffision coefficients inferred £tom these short
term tests could be taken as valid and representative of the pure difision. These tests
were repeated (results for the VOCs are presented in the Chapter 5, that follows) and left
nmning for 71 days. Designated as "Tea 2" in the Figs 4.2 (acetate) and 4.3 (propionate),
duplicate diffision tests were successfuily modeled without sorption and without
degradation using the same difision coefficient of 2.5 x 10.'~ m2/s for acetate and
slightly lower coefficient of 1.5 x 10-1° m2/s for propionate. The resulting fit-lines were
again closely rnatched to the data (lt2s > 0.97 for the two acids) and accompanied with
full mass recovery for both acetate and propionate. However, after 48 days of testing,
some characteristic changes were noted suggesting that degradation of acetate and
propionate might have started. Very tiny grayish dots (initially 5 Imm) appeared first on
the contact wall surface in the ce11 with acetate. At the time of termination @y 71 days),
these dots were more visibly black and bigger (2-3 mm) and unevenly distributed along
the 2 cm plug. In the ce11 with propionate, tiny grayish spots were localized at the
boundary between the glas disc and soil plug. However, there was neither recorded
increase in soi1 porosity nor visible soi1 expansion and gas generation. Source solutions
were both clear and odorless, while receptor solutions, although clear, had traces of
decaying, but unlike-H2S, odor. BART" tubes confirmed possible 100-fold increase in
HAB, (6 x 10' cfwg) and IO-fold increase in SRE (6 x 105 cju/g), suggesting growth of
these bactena in the soil at the expense of acetate and propionate. Thus, it appears that
degradation of acetate and propionate had begun in the short 2 cm compacted plugs, but
the rate was so low that it is Wtually impossible to measure the loss of mass from the
system. The data for the "Tests 2" upon termination after 71 days indicate distinct
diffusion profiles for acetate and propionate in both source and receptor solutions as well
as in the soi1 pore water. These profiles can be satisfactorily defined with the same
diffusion rates (coefficients) as for the early stage of difisive transport discussed for the
case of the "Tests 1"when degradation could not be deteaed.
Results fiom the difision of butyrate through compacted Halton till are shown in
the Fig. 4.4. Butyrate measurements collected in time indicate somewhat slower decrease
in concentration in the source solution and consequent slow breakthrough in the receptor
solution, relative to acetate (Fig. 4.2). The depth profile after 28 days of testing revealed
the presence of acetate in the pore fluid of the squeezed soil samples as seen in Fig. 4.4.1.
The soil plug had visible blackening at the bottom contact surface with the glas disc and
receptor solution, with tendency of upwards spreading? however the top surface of the
plug retained its original rusty color without any black metallic sulfide formations. The
receptor solution had a characteristic H2S smell, while the source retained the initial
slightly rancid butyrate odor. Butyrate was not labeled, however it was the only acid
tested together with (low mg/L levels of) VOCs spiked in the source solution. It is
speculated that acetate was generated as a result of butyrate fermentation. It is unlikely
that acetate could have been formed at detected levels fiom VOCs fermentation under the
tested conditions. A 14% loss of butyrate was recorded d e r 28 days (Test 1, Fig. 4.4),
however, when recovered butyrate and acetate were expressed as (total) dissolved oganic
carbon, the overall mass recovery exceeded 100 %. For details of mass balance
calculations see Fig. A4.2.3, Appendix 4. The recorded mass surplus is -10% of initial
mass, which is considered tolerable given the possibility of expenmental error and
dilution required by analytical method.] Butyrate concentration profiles in source,
receptor and pore water solutions were modeled using slightly lower difision
coefficients (of 1.5 x 10-'O and 9 x 10-") m2/s than in case of the other two VFAs. This
produced good fit lines, however, the overall fit was improved when degradation of
butyrate was taken into account, as can be see in the Figs. 4.4, with details given in Fig.
4.4.1. It is noted that the dflerences resulting fiom simulating diffision with degradation
and diffision with sorption, as illustrated in the Fig. 4.4.1 are marginal at the tested scaie.
Given the properties of the acids tested and evidence in favor of degradation, sorption is
mled out as likely mechanism that could influence (Le. retard) the m a s transport, even on
the large scale in the environment such as low permeabiiity oligotrophic clayey deposit.
In order to descnbe the pattern of the data points, butyrate degradation is modeled
in the bottom layers only (Le. below top 0.6 cm) with the rationale that the low arnounts
of poor fermentable substrate such as butyrate could be conducive for the carbon starved
and disturbed indigenous soil microorganisms. This suggestion is in agreement with
findings of Hoeks & Borst (1982), who observed longer delay (lag) of CHJ fermentation
associated with higher VFAs concentrations in column experiments with sandy-loarn. For
the tests presented herein, the degradation induced by low and thus more conducive
substrate concentration is also suggested by the fact that acetate, as the likely
intermediate, was detected in the receptor and in the bottom soil layers. In addition, the
smell of HzS was present in the receptor, pointing to the possible terminal consumption of
butyrate by sulfate reducen. A - 50 fold increase in both HAB and SRB (- 7.5 x 10') in
the soil after 28 days of testing, confirmed that some microbial growth occurred as well.
Nevertheless, it seems that manifest explanation for this "early" butyrate disappearance,
indeed is low concentration (290, 160 and 20 mg/L in Layer 2, 3 and 4 respectively, as
opposed to 740 mg/L in the Layer 1 pore-water.). Butyrate was initially introduced at low
(and representative) levels (Co .e I g/L), and became available to the microorganisrns
through difision (Le. dowly) and without inhibition imposed by shock organic loading.
Butyrate concentrations higher than - 1 g/L were not tested, thus it is not known
whether diffision under such conditions would be affected. It is believed that the high
concentration (g/Z in the pore water) of acetate and propionate, both otherwise readily
degradable, is a factor in delaying the degradation, as observed in the short term diffision
"Tests 1", discussed first (shown for "Tests 1 and Test 2" in Figs 4.2 and 4.3). This view
is reinforced with additional tests, designated "Test 3" in the Figs 4.2 and 4.3 for acetate
and propionate and "Test 2" for butyrate in the Fig. 4.4. Test 3 containing only acetate or
propionate in the source solutions, each left running for 120 days (ody 85 days s h o w in
Fig. 4.2 and 4.3), had 89.6 % and 59% mass recovery for acetate and propionate
respectively. Test 2, shown for butyrate in the Fig 4.4, is the only one employing source
solution with al1 three of the acids, which nevertheless, resulted in mass recovery of 107
% for acetate, 93 % for propionate and 86% for butyrate or, 101% expressed as TOC
(total organic carbon) d e r 172 days of testing (only 6 1 days show). Interestingly, for al1
three of the mentioned tests, nice fit lines (see data and lines for the "Tests 3" in the Figs.
4.2, 4.3 and for the test 2 in Fig. 4.4 and Figs. A4.2.1, A4.2.2 and A4.2.3 in the Appendk
4) were matched to the data in the source and receptor solutions with the same diffision
coefficients (given in the Table 4.1 and) simulated for short-term diffision tests discussed
earlier. Variation of the diffusion coefficients within the bounds proposed in the Table 4.1
resulted in negligible difference in the VFAs impact for the tested scale. (Details of the
sensitivity analysis for acetate are enclosed in the FigA4.2.1.1, Appendix 4.)
It appears that even at 60, 120 or 172 days, diffision of the M& practically
proceeds at the rates inferred 4 t h -30 days of testing, although mass balance indicates
some loss and B A R T tubes confirm commencement of degradation. The extent of loss
is for most of the presented tests within marginal limits even, with occasional tolerable
surplus, which could be attributed to experimental error. It is also noted that control tests
performed in the serum bottles with the original source solution stock as prepared for the
Tests 1 & 2 for acetate and propionate and the Test 1 for butyrate confirmed that each of
these acids was stable in the solution without measurable loss during testing period of 60
days (See Fig. A4. lb, Appendix 4). NI of this points to the fact that difision could be
successfully tested with non-sterile soi1 (and solutions) and as such could be differentiated
from degradation at the early stage of monitoring. Recommended values ranging fiom D
= (2.5 - 5.0) x 10*'~ nr2h for acetate, D = (2.0 - 5.0) x 1 ~ ' ~ m% for propionate and D =
(1.5 - 3 .O) x 10." m2/s for butyrate are slightly lower than D = (5 - 7) x IO-'' m2/s which
was obtained as diffusion coefficient for chloride (Cl3 for the laboratory compacted
Halton till under sirnilar testing conditions (Rowe & Barone, 1991). It is possible that on
large scale with heterogeneities, diffision coefficients for the VFAs tested rnight be
slightly higher than those used to generate the best fit-lines, which is not obvious from the
short soi1 plugs and simulations with only limited number of datapoints. It remains yet to
be seen whether the range recornrnended for use within the factor of 2, (as given in the
Table 4.1 and in Figs. A4.2, Appendk 4), would hold for a large scale and in state-of-the-
art landfilis, but no data fiom the field were available for examination. It is recognized
that for the field profile with history of contamination, it might be difficult to reconnrua
difision profile because of degradation. Nevertheless, VFAs, as polar and anionic at
neutral (and higher than neutral) pH, could be generally regarded as the most mobile class
of organic chemicals in the pore and ground water. It is considered that the difision rates
for this class of chernicals could not be much higher than inferred from the presented
tests. For the conditions examined, VFAs are not expected to diffuse faster than chloride,
an extensively tested, inorganic and, for al1 practical purposes, non-reactive anion. These
acids will, however, be subject to biodegradation even in the compacted clay deposits
where mass transfer limitations could hinder the rates of any microbially dnven reactions.
Had the diffision rates been higher than deduced, only higher mass loss (Le. degradation)
would have explained the data in the soi1 pore-water and source and receptor solutions
seen in the Figs. 42-44 . Such high mass losses were not recorded, thus diffusion and
degradation faster than inferreci, could be ruled out. The results fiom short term testing
compiled herein indicate that degradation of VFAs under dominant difisive transport in
compacted soil and at neutral pH is detectable, yet indeed very slow and, not quantifiable.
I t is considered that the observed results would reflect the situation that might occur in the
field since the tested concentrations are environmentally representative for the relatively
early stage of landfill operation at the onset of methanogenesis when total (and dissolved)
organic carbon pollution is still signifiant and as such will be subject to diffision
through uncontarninated soil, regardless of prospective removal due to degradation.
Bearing this in mind, it is considered that deduced diffusion rates for VFAs given in the
Table 4.1 are within reasonable and expected lirnits and as such, they could be used for
prediction of fate of these chemicals in the low permeability soils.
4.4 Intrinsic degradation of volatile fatty acids (VFAs) under diffusive transport in
compacted clnyey soil
The fundamental microbiological and chernical processes taking place in sanitary
landfills have been extensively investigated and numerous research documents have been
published during last twenty five years (Farquhar & Rovers, 1973; Baedecker & Back,
1979; Rees, 1980; Aragno, 1988; Christensen & Kjeldsen, 1989; Christensen et al., 1994;
Watson-Craik & Jones, 1995). ui bief, complex organic waste polymers (e.g. cellulose,
proteins, peains, food, paper) are hydrolyzed to bio-monomers [amino acids, sugars,
long-chah fatty acids and alcohols] and fermented further to volatile fatty acids (Wh)
such as, acetate, propionate and butyrate, commonly found in anaerobic digesters. At this
stage, waste is anaerobic and acidic producing leachate containing high concentrations of
these acids as well as dissolved metals and inorganic salts with gaseous fermentation
products such as Ht and CO2. Graduaily the terminal mineralization processes establish
resulting in generation of Iandfill gas ( C h CO2), with less Hz, and neutral to alkaiine
leachate low in volatile fatty acids and inorganics. The duration of various degradation
stages is quite variable and unpredictable because of the uncertainties associated with
waste organic load and landfill operation regime. Acidogenic and acetogenic stages
characterized with massive organic load (high COD) are not expected to exceed several
years, while methanogenic stage could last few decades or longer (Robinson, 1995).
Many modem landfills employ well functioning leachate collection systems,
which do remove the peak organic load during initial years and could be engineered to
remove less contarninated leachate generated at later stage of confinement. However due
to concentration gradient imposed by the waste disposal into unpolluted surroundings,
diffision of the contaminants inevitably proceeds, ofien enhanced by leachate-mound
driven advection thus posing potential threat to groundwater resources. Although modem
landfills also employ various engineering components designed exclusively as diffision
and advection barriers, the impact of contamination could still be grave and sometimes,
difficult to tackie given the very stringent regulations and declining water resources.
Because of this, it becomes important to explore possibilities of contaminant attenuation
in addition to those provided by confinement.
The focus of the work presented in this thesis is Uitrinsic degradation of volatile
fatty acids in compacted clayey soil, which is used for the construction of engineered
diffision bamers in a modem landfill. These acids have been recognized as major organic
readily degradable contaminants in landfill leachates (Harmsen, 1983; Robinson, 1995),
and in the aquifers undemeath landfills (Blakey & Towler, 1988; not implicitly in
Christensen et al., 1994); their treatment in waste water plants has become a routine
(Harper & Pohland, 1986; Rittmann et al., 1988; Pavlostathis & Giraldo-Gomes, 1991;
Nedwell & Reynolds, 1996). Biodegradation of these acids has been studied and
rnetabolic pathways are weli known (Fuchs, 1986; Doifing, 1988; Oremland, 1988;
Widdel, 1988; Chapelie, 1993). As products of hydrolysis and fermentation, propionate,
butyrate (and other short chah C3-Cs) and, most importantly acetate and hydrogen,
released as reduced medium and low-energy intermediates of oxidation/reduction steps in
the breakdown of complex organic matter are utilized as growth substrates by mixed
cultures fonned of the most simple microorganisms (prokaryotae and archaea) from the
very bottom of the phylogenetic tree of life. Acetate and hydrogen are well recognized as
methane precursors, and competitive substrates for both methanogenic and sulfate
reducing bactena (Dolfing, 1988; Widdel 1988; Cord-Ruwisch et al., 1988; Hoehler et al.,
1998). Both of these groups are instmmental in the terminal carbon and hydrogen flux in
natural environments.
Al1 known phylogeneticaily and rnorphologically diverse species of methanogens
are archaebactena and aria anaerobes that have unique ability to produce methane in
their respiration. They are found in freshwater and manne sediments, marshes, even in
interioi of soi1 particles generally considered oxidized, and are active in the environments
where pool of HCOi and/or CO2 exists while more energetic electron acceptors (other
than 02, such as NO3-, ~e~~ and ~01'3 are depleted (Ehrlich, 1996). Most of the species
( e . g . Methano bacterium, Methuno brevibacter, Me thtococcus, Me thanogenium,
Methanomicrobium, MethanospirilZum) are obligate or facultative autotrop hs t hat obtain
energy from reduction of CO2 with H2 (or its equivalents, formate or CO, sometimes
called Hz-oxidizing methanogens, Harper & Pohland, 1986). Heterotrophs (e.g.
Methanosueta, formerly Methanothnk, and Methmowrcina), utilizing acetate as energy
and carbon source are also known and engaged in other major CH4 release pathway:
aceticlastic (or acetotrophic) methanogenesis (Ehrlich, 1996). Only a few of very simple
organic compounds such as Ht, CO, formate, methanol, methylamines and acetate [with
just a couple of exceptions with Ca alcohols reported recently, (Ehrlich, 1996)l are used
as energy source, thus in order to thrive, methanogens fonn syntrophic associations with
other heterotrophic fermenters and /or anaerobic respirers. The syntrophic partners,
characterized as obligate proton-reducing (or hydrogen-producing) acetogens, provide the
simple substrates for methanogens (i.e. mainly hydrogen and acetate) while methanogens,
by consuming hydrogen, keep its concentration low and regulate the metabolism of
partners thus enabling the final metabolic steps to proceed without inhibition. This
process, known as interspecies hydrogen transfer, is comrnon in anaerobic digesters
(Young 1984; Harper & PohIand, 1986), and its importance is well recognized in natural
anaerobic habitats, such as anoxic sediments (NedweU, 1984; Hoehler et al., 1998) and
landfills (Rees, 1980; Aragno, 1988). In their 1967 landmark paper, Bryant et al. (1967)
speculated that non-methanogenic bacteria play crucial role in catabolism of fatty acids
other than formate and acetate. It took more than a decade before the first CO-cultures of
proton-reducers and Hz-consumers, Syntrophobacter wolinii (Boone & Bryant, 1980) and
Syntrophomonas woljei (McInemey et. al, 1981) capable of anaerobic oxidation of
propionate and butyrate, respectively, were isolated. In each case, acetate and hydrogen
were the intennediates of the oxidation and were fiirther mineralized either to methane or
to hydrogen sulfide.
Fementative microorganisms obtain their energy through substrate-level-
phosphorylation in which regeneration of their coenzymes proceeds through release of
more reduced intennediates such as propionate, butyrate, lactate, ethanol or succinate (i.e.
various electron acceptors are used for electron disposal from reduced NADH to yield
oxidized NAD'). At very low H2 concentrations (< 1 o4 atm or < 10 Pa), obligate proton-
reducing bacteria are able however, to re-oxidize their coenzymes (and regain energy) by
the direct release of H2 (Le. proton-reduction, since oniy protons serve as electron
acceptors). This metabolic "shortcut" results is slight increase in ATP relative to
fermentation, and decrease in the arnount of reduced fermentation intermediates with
stoichiometnc increase in acetate (Nedwell, 1984; Harper & Pohland, 1986; Aragno
1988; Dolfing 1988; Schink, 1988). In their reviews, Nedwell (1984) and Shink (1988)
pointed out that many bactena known as hydrolytic and fermentative have ability to shift
their carbon metabolism away from fermentation products such as propionate, butyrate
and acetate directly toward H2, COÎ and acetate (as observed in many reported lab studies
on anaerobic degradation of cellulose, benzoate and glucose) if the Ht partial pressure in
the environment is maintained low.
Another geomicrobiaily and geochemically important group of bacteria other than
methanogens that can utilize H2 and regulate its partial pressure in reduced environrnents
and carry out anaerobic mineralization, is sulfate reducing bacteria (SRB). Most of the
known species are eubacteria with only two belonging to archaebacteria (Ehrlich, 1996).
Sulfate reducers ubiquitous in many naturai habitats rich in inorganic sulfates, particularly
in soil and marine sediments are strict anaerobes nutritionally more versatile than
methanogens. Several species are autotrophs that grow on H2 as energy source, but
majonty are heterotrophs, able to attack and very ofken, to mineralize various aliphatic,
aromatic and heterocyclic organics (Ehrlich, 1996), also recognired as contaminants. This
remarkable group of baaeria is generally capable of both fermentation and obligate
proton reduction, depending on the environmental conditions (Widdel, 1988).
Desulfobacter species (e.g. postgatei, and hyùiogenophilus) can easily grow on acetate,
oxidizing it completely to hydrogen sulfide and CO2, as well as somewhat slower but
more versatile fresh water isolate Desulfotomaculum acetwxidans. A propionate
degrader, Deszrlfobulbus propionims oxidizes propionate incompletely to acetate.
Desulfovibno species ("sqvovorans group") are capable of incomplete oxidation of C-
even fatty acids to acetate and C-odd fatty acid to acetate plus propionate, while some
other more versatile species such as Desulfosurcina vanobilis, Deszilfococcus,
Desirlfonemu and Desulfobacterium may completely oxidize C io-C i6 rnonocarboxylic
acids to HCOs-/CO2 and hydrogen sulfide (Widell, 1988).
Activity of methanogenic, sulfate reducing as well as vanous facultative
fermentative bactena has been confirmed in sanitary landfills. Viable counts of anaerobes
beneath the domestic landfill ranged from 3.3 x 10' to less than 1 x 102 c fups (wet soil)
in unsaturated and - 2 x 10' cf./gws in saturated zones, and 6.6 x 106 to non-detectable in
uncontarninated chalk aquifer, each along 18 m or more of depth of borehole profiles
examined by Towler et al. (1985) and Blakey & Towler (1988). Sleat et al. (1989)
reponed variation in count for fenentative bacteria with depth in the Aveley Landfill,
UK. The highest count of 106 - IO* cwgwr ((wet refuse) was observed in unsaturated
refuse at the highest concentration of total VFAs (> 10 g/L) and p H = 6, while count
dropped to - 104 c f u p r at pH o 8 and 8 m below the surface, in saturated zone where
VFAs were not deteaed. h contrast, methanogens were the highest at - 10 rn below the
surface, aceticlastic at - 106 and Hz-oxidizing at 10' cfu/gr, respectively. SRB exhibited
the least variation dong the depth, with max count of - 5x10~ cfu/gr at depth of 4 m
which corresponded with the highest level of sulfates (- 3 g 4 ) in the refuse. These counts
by Sleat et al. (1989) reported for the real landfili, were higher than 2.5~10' for
fermentative, but lower than 5x 1 o8 (and 4x 10') for methanogens (and acetogens
respectively), as reported by Barlaz et al. (1989) for the laboratory landfill simulators.
Numerous studies of rnicrobial ecology of various natural subsurface
environments also cofinned that rnany indigenous microorganisms are not stktly
autotrophic. On the contrary, they are active and capable of utilizing organic substrates as
well as carrying out terminal oxidation/reduction reactions such as methanogenesis and
sulfate reduction (Ghiorse & Wilson, 1988; Kolbel-Belke et al, 1988; Martino et al.,
1998). Jones et al. (1988) reported that native microorganisrns from deep Coastal Plain,
SC, sediments sampled throughout the depth fiom 20-300 m. although not generally
anaerobic (aerobes were 10'- 10'-times higher) were capable of methanogenesis and
sulfate reduction. The count of SRB varied with depth, ranging from non-detectable in
consolidated clay layers, or from c 1 tu > 10' c . g in penneable water bearing strata.
Methanogens could not be numerated, however, methanogenesis was detected in
metabolic activity tests in aquifer slumes from various depths. Coliforms, chosen as
representative of facultative anaerobes, were easily detected in the shallow layers (< 50
m) ranging up to 104 cfw'g. Lactate and formate were consumed from most of siumed-
layer samples, even in those of the clay deposits from 125 m depth, but after - 92 to 240
days. More interestingly, Jones et al. (1988) reported that acetate metabolism in the
slumes exhibited four distinct patterns: (1) no anaerobic degradation, (2) slow
disappearance after 3 - 4 months without methanogenesis, (3) acetogenesis (acetate
generation) at initial 120 days of incubation with neither subsequent removd nor
methanogenesis (observed in samples fom deep sand and clay layen), and (4)
acetogenesis dunng initial 60 days followed by methanogensis (observed in samples From
surface and saturated sand layers). Acetogenesis was attributed to the reserves of
fermentable carbon naturally present in these sediments (possibly from lignin decay) and
supply of the inorganic nutnents (N, P) from the cultivation media. Methanogenesis was
also detected after 2-5 months in the slumes fiom the shallow sandy layers, and was
positively correlated with the presence of SRI3 and their activity. Furthemore, anaerobic
benzoate degradation was localized to the layers with active methanogenesis. As stressed
by Jones et ai. (1988), it was only in the transmissive (sandy) saturated layers where
acetate and benzoate degradation as well as methane generation proceeded unhindered.
Although the tests described in the Jones et al., (1988) remarkable study were not
performed with intact consolidated soil samples, its findings point to the potential for
intrinsic degradation that indigenous soil population might have when in contact with
readily degradable organic compounds, such as VFAs and inorganic nutnents. Very few
studies were found dealing with bio-chernical reactivity of VFAs and their impact on
environment geochernistry as a consequence of sanitary landfill emissions into natural
low hydraulic conductivity deposits or engineered diffusion barriers (Baedecker & Back,
1979; Back, 1986). Early work of Hoeks & Borst (1981) indicated potential for "methane
fermentation*' in pH (6.5 - 7.0)-bufEered leachates percolated through sandy loam. The
column tests were performed with leachates containing various VFAs at different but hi&
concentrations (and high COD) and mildly acidic pH (5.6 - 5.8), however the type of soil
used by Hoeks & Borst (1981) although not specified was not quite representative of (low
hydraulic conductivity) diffusion barrier since it allowed percolation. Nevertheless, VFAs
degradation was fast, reporiedly with visible methane bubbling out of the columns. It was
stated, although not s h o w that at long retention times common to field conditions, VFAs
would be cornpleteiy rnineralized from the leachate in the top first meter of soil beneath
the waste pile. In the other document on VFAs degradation, McMahon & Chapelle (1991)
reported a build-up of acetate and formate in organics-rich sulfate reducing aquitard at the
Lake City site, SC. These acids were dnven by difision gradient into the underlying
deep anaerobic aquifer, where the conducive conditions developed for their fast
mineralization. The sarne observation was gathered in the laboratory with glucose
anaerobic assay and field inocuia: d e r 25 days, accumulation of acetate, formate and
CO2 occurred in the aquitard but not in the aquifer sediments.
Considerate body of research has been compiled on the fate of volatile fatty acids
in marine sediments. In these natural habitats, anaerobic mineralization of volatile fatty
acids (originating from organic detritus) takes place predominantly under sulfate
reduction (Bardona, 1982; Nedwell, 1984). High concentrations of sulfates (- 1 o-~M)
usually found in these environments (Young, 1984), sustain this terminal electron
accepting process, thennodynamically more favorable than methanogenesis (Thauer et
al., 1977; Harper & Pohland, 1986) and as a result methane generation becomes inhibited
(Shaw et al., 1984). Sulfate reducers generally have higher affinity for hydrogen than
methanogens, [e.g. half-saturation constant for hydrogenase from Desulfovibrio vulgarii
was K, = 1 pbf, which is lower than K, = 6 pM, found for Methanobrevibacterium
arborphilus, Kristjansson et al. (1982)l and are able to maintain its levels below those
required for optimum growth of methanogens. Research with naturai sediments as well as
sediment slurries and wastewater showed, that in spite of this disadvantage, arising From
competition for hydrogen (Hoehler et al., 1998), the two terminal processes are not
antagonistic, but rather commensal. Methanogenesis does proceed simultaneously with
the sulfate reduction. It is very slow in sulfate-nch marine sediments but in fresh-water
sediments nch in bicarbonates it prevails (Senior et al., 1982; Harper & Pohland, 1986;
Widell, 1988). Rates of hydrogen and acetate (and consequently, other waste fermentation
intermediates) uptake are faster under sulfate reduction than under rnethanogenesis but no
such benefits, as energy recovery from methane generation are available. However it is
believed that neither of the two processes would be effective in the compacted sediments
having (10" - 10"') mis hydraulic conductivity: the rates of any reaction would be
severely dampened, thus the eventual end produas either in form of corrosive HzS or
calorific C& would have neither much harm nor much benefit, respectively. Given that
both sulfate reduction and methanogenesis are terminal processes of anaerobic waste
breakdown, that is, each means mineralization, it is most extraordinary that they can take
place in the naturd environments and without deliberate technological intervention. In the
recent compilation of intrinsic degradation case shidies, Wiedemeier et al. (1999) pointed
that almost 90% of the BTEX biodegradation capacity, observed at the 38 petroleurn
contaminated sites could be attributed to anaerobic mineralization processes, particularly
those driven by sulfate reducing and methanogenic bacteria.
Compacted clay liners or natural confining beds are constituting part of a modem
disposal facility designed for d e confinement and their inherent limitation to mass
transport, is taken or engineered as an asset. However given that contaminants can difise
through those liners, it becomes important to account for mechanisms of contaminant
degradation in these liners that can irreversibly destroy a part of the pollution. Because of
the slow or "no-" flow combined with slow difision through these clayey barriers,
retention tirnes of the contaminants are very long. Generalîy, environmental rather than
thennodynarnic constraints have been recognized as major determinants of VFAs
metabolism. In ail reported cases of fast turnover and/or VFAs consumption in marine
sediments, the reaction was either drastically irnpeded or completely ceased below the
few top centimeters of a bio-reactive layer (Nedwell, 1984). Phelps et al. (1988) stressed
that unsaturated conditions and soil texture with 20 % clay impaired the size and activity
of the indigenous populations. In addition to low hydraulic conduaivity and suboptimal
pH, Chapelle & Lovley (1990) attributed small pore size (< 0.05 p) and very low
effective porosity (0.078) to low (<0.025 yi', i.e -28 year-haIf-[fi) rates of acetate
turnover observed in clayey confining beds. Diffusion which govems the bio-availability
of organic substrates is considered a factor that severely lirnits rnicrobial respiration and
consequent biodegradation (McMahon & Chapelle, 1 99 1 ; Verstraete & Top, 1 999)
Nevertheless, the decades (if not centuries) of retention associated with modem
landfills might be very conducive to intrinsic degradation of organic contarninants, since
the indigenous soil population have sufficient time during just a few years of
incubation lag to acclimate to contaminant-substrates and remove some of them fiom the
environment under the rates which would, in any other man-made industrial system be
considered "impractically" slow.
The main objective of this chapter is to examine whether few representative
organic pollutants, found in municipal solid waste landfill (MSWL) leachate, could
degrade in the compacted clay. In this chapter the emphasis is put on the C2-C volatile
fatty acids, acetate, propionate and butyrate recognized as predorninant products of
organic waste hydrolysis and fermentation. The conditions that would evennially lead to
their breakdown (Le. mineralization in the soil) are exarnined. Experiments are proposed
to sirnulate the described adverse conditions for reaction imposed by mass transfer
limitations and difhsion. Finally, the rates of degradation for the volatile fatty acids
tested are inferred in an attempt to facilitate the prediction of their impact on environment
and groundwater resources.
4.4.1 Materials and method
The intrinsic degradation tests were performed in two glass-ceIl assemblies shown
in Fig. 4.5. One assembly employed eight (8) giass ceUs comected to the continuous feed
distribution network (see also schematic A4.3 Appendk 4), intended to simulate
degradation under the 1-D d a s i o n dominated mass transport for a long penod of testing.
The schernatic of the glass diffusion ce11 is given in the Fig. 3.2 (Chapter 3) and the
details of the preparation procedure followed in its entirety, are elaborated in 5 3.5.1
Chapter 3. Halton till compacted plugs prepared as described above (5 4.2.1 and 5 3.2
Chapter 3), were used in the intnnsic degradation experiments that follow. Expecting the
measurable degradation to take place at slow rate and &er a considerable lag penod, the
test was started with 8 replicate cells, which were intended to record the occurrence of
degradation in tirne. At a number of times, when the information might be
chronologically usehl or when some of the changes became obvious, one ce11 or a pair of
cells was terminated (sacrificed) and soil and solutions were analyzed for designated
parameters. The terminations continued until no cells were left ninning. Our previous
tests and trials indicated that 3 cm soil thickness, although convenient for short-term
testing might not always provide desired insight and sufficient number of datapoints,
therefore another assembly with 5 cm soi1 thick plugs was ernployed as a replicate (see
Fig. 4.5, upper photograph) in order to collect more information particularly on soil and
pore water concentration profiles.
The synthetic medium used in intnnsic degradation tests was made up to resemble
both organic as well as inorganic composition of the reai Keele Valley Landfill (KVL)
leachate used in the preliminary diffisionldegradation tests performed earlier (Chapter 3,
also see Rowe et al., 1997). Based on the data collected from numerous chernical analyses
of this leachate (see Table 2.1, Chapter 2 and Rowe, 1994), a synthetic KVL leachate
solution was created with the composition given in the Table 4.2. Volatile (short chain)
fatty acids (VFAs) abundant in the leachates generated fiom municipal solid waste
(MSW) landfills such as acetic, propionic and butyric represented a major source of
dissolved organic carbon (DM) and chemicai (and biochemical) oxygen demand (COD,
BOD), while volatile organic chemicals (VOCs) spiked to this synthetic leachate were
representative of several micro-pollutants fkequently found in domestic landfills. This
synthetic mix was designed to simulate aabiiized land£ül leachate at the beginning of
methanogenic stage, except that, although not sterilized, it did not have an active
microbiai population characteristic for a MSW leachate.
Intrinsic degradation tests were intended to simulate continuous feed of organic
and inorganic nutrients from synthetic leachate solution to the soil rnicroorganisms under
dominant diffisive mass transport through compacted clay plug. Apart from system
limitation imposed to mass transfer itself, this testing scenario also includes the most
adverse conditions of sirnulated contamination because it does not employ any physicai
removal of the contarninants From the point source. A properly designed and maintained
primary leachate collection system is a key component of the state-oGthe-art landfils,
however, certain simplifications had to be made in order to simulate Iaboratocy scale
degradation in the soil attenuation layer below such leachate collection system. Driven by
need to make the experimentation manageable and yet to methodically assess the impact
of a pariicular cornponent without much interference, the leachate was represented by
solution of organic and inorganic contarninants, as elaborated elsewhere (Rowe et al,
1995). It is recognized that such testing at constant and relatively high organic
contamination rnight generally result in over-prediction of impact. However given the
very uncertainties associated with initial and actuai leveis of field contamination as well
and limited time fiame available for any experimentation it is considered that testing
scenano and organic load employed in this methodology can realistically represent, or
rather simulate processes in a municipal solid waste landfiil and field low permeability
attenuation layer at their initiai and somewhat adverse phase.
The concentration of the eleven organic chemicals selected (three fatty acids and
eight volatile organic compounds) was monitored in time in the ce11 solutions and soil
pore water in order to assess the prospects of their intrinsic degradation in the laboratory
compacted clay liner. The results for the fate of the three volatile fatty acids follow in the
next paragraphs, while those for the volatile organic chemicals are given in Chapter 5.
4.4.2 Analytical measurements
Detailed description of al1 analytical methods and protocols used for
quantification of the parameters tested is given in 5 4.3.2 above.
4.4.3 Results and discussion: Laboratory intrinsic degradation tests
Fig. 4.6 shows visually the results of 162 days of intrinsic degradation. Comparing
this with the picture at the beginning of the test in the Fig. 4.5, it is evident that the soi1
and solutions significantly changed with time due to biodegradation drïven by the
indigenous soil bacteria. The results coiiected from monitoring the change of selected
parameters due to the intrinsic degradation corroborate the image in Fig. 4.6. The change
of porosity in the soil plugs is given in Figs. 4.7.1 and 4.7.2. The growth of selected
genera and activity of soil indigenous microorganisms induced by degradation of WAs is
displayed in Figs. 4.8.1 - 4.8.2 and 4.9.1 - 4.9.2. Finally, the results of intnnsic
degradation of the selected volatile fatty acids are given in Figs. 4.10.1 - 4.10.2 to 4.13.1 - 4.13.2, for the two teaing assernblies respectively. The top graph (or 2 graphs, for some
of the parameters) on each figure show(s) the change of tested parameter in tirne, while
the bottom graphs show the change ofthe corresponding parameter vs. depth (i.e. in pore
water or per unit dry mass of the soil) at the times of test termination. The best estimates
of the diffusion coefficients and half-lives for VFA decay were obtained using cornputer
program POLLUTE v.6.5 (Rowe and Booker, 1999), which solves mass transport
equation coupled with first order decay in the sahirated pore space (defined as Eq. 1.1 in
Chapter 1; also see A3.2, Appendix 3). The sumrnary of the monitored and estimated
parameters and their change in time is given in Table 4.3.
As expected based on observations from the tests perfomed earlier (see Chapter 3
and Rowe et al., 1994; 1997) the soi1 plugs graduaily started to change in appearance
upon the exposure to the synthetic KVL leachate. Rusty (tan, earth) color clearly seen in
the Fig. 4.5, characteristic of oxidized soil (i.e. one containing some dissolved oqgen in
pore water as well as natural pool of other electron acceptors such as femc iron, sulfates
and carbonates) started subtly to fade. Dark gray or black tiny dots (< 1 mm diameter)
first appeared randomiy dong the contact surface with the ce11 walls and soil within less
than a mont h of testing.
The first ceIl was terminated f ie r 23 days. Breakthrough of al1 of the VFAs as
well as some of the more mobile VOCs (VOC details follow in Chapter 5) had already
been detected (albeit at low concentration) in the receptor. Celi 1-1, with a 3 cm thick soil
plug looked intact and unaf5ected by discoloration. However there was a hint of HIS odor
in the receptor solution. The receptor solution was clear, while the source solution was
only slightly cloudy (although ail1 transparent). The second termination involving
another 3 cm thick soil plug in the ceil 1-2 followed after 29 days of testing. The soil in
ceIl 1-2 looked sirnilar in appearance to soi1 in the cell 1-1. It was hypothesized that
diffision of VFAs and VOCs in the leachate would not be significantly affected by
biodegradation at this early stage of testing. With early tennination of these cells, it was
also intended to reassess the rates of diffision under the continuous supply of VFAs and
compare the results with those inf'erred from independent short-term tests discussed
earlier in 8 4.3.3.
The results gathered on the examined parameters indicate that cells 1-1 and 1-2
could be treated as duplicates. Porosity, calculated from the measured moisture content
for each layer cut form the soi1 plug(s), remained unaffeaed relative to the initial 0.34
throughout the entire thickness, as can be seen in Fig. 4.7.llb)(l) and Table 4.3.
B A R T ' S confirmed a uniform increase in number of heterotrophic aerobic bacteria
(HAB) as well as sulfate reducing bacteria (SM) in the soil layers relative to the
background levels, as shown in Figs. 4.8.1 (c)(l) and 4.8.l(d)(l), respectively. (For layer
designation, see Table 4.3, Figs. 4.7.1 and 4.7.2.) It seems that on average (i.e. relative to
mid range count reported in Table 2.2, Chapter 2) HAB grew -160-fold in 23 days in cell
1-1 plug and -1400-fold in 29 days in the cell 1-2 plug, as shown in Fig. 4.8.1 (c)(l),
respectively. Given the high degree of variation associated with heterogeneities inherent
to the soil properties, including distribution of soil indigenous rnicrobial population in
panicular, as well as limited precision of the determination method allowing standard
deviation of 1.5 log change in numbers, the IO-fold difference in count seen in Fig.
the expected range. Growth of
over background (on average)
4.8.1(c)(l) between the two presumed duplicates is within
SRB appears less vigorous with a 3 1 0-20)-fold increase
L
and was uniform for the two 3 cm plugs in the Fig. 4.8.l(d)(l). There is a high count of
HAB in source and receptor solutions in contact with the two tested soil plugs, all
pointing to the possible growth of this ubiquitous group of bacteria. Measurements of
ATP, shown in the Fig. 4.9.l(c)(l) confirrned - (10-30)-fold increased activity of the
viable soil biomass in both soil plugs and detectable although low, activity in source and
receptor solutions. Measurements of the three VFAs however, indicated very distinct
difision profiles in the 3 cm soi1 plugs confodng to the rates expected based on the
diffusion coefficient deduced from shon-tenn tests and given in Table 4.1. The long-dash
fit-line in the Fig. 4.10.1 (b)(l), generated with diffision coefficient of D = 3.5 x 1 0 " ~
m2,s for acetate is in very good agreement with the data points collected for cell 1-2. Data-
points collected for ce11 1-1, although fonning diffision profile appear to deviate £Yom the 10 2 D = 3.5 x 10' rn /s theoreticai (dotted) fit-line. Generdly, the analogous patterns are
recreated with D = 2.5 x 10''~ rn'h and propionate data-points for the two terminated tests,
but with somewhat better fit for the ce11 1-1 data points, than the one generated for acetate,
as can be appreciated from the Fig. 4.11.1(5)(1) and compared with lines in the Fig.
4.10.1 (b)(l). Pore water data and the D = 1.5 x 10"~ m2/s - generated fit lines for butyrate
in the Fig. 4.12.1@)(1) however, are in excellent agreement for both 1-1 and 1-2 cells.
The sarne results were obtained with a set of duplicates employing 5 cm thick soi1
plugs. Cells 1-2 and 1-1, tenninated d e r 33 and 49 days of continuous exposure to the
synthetic KVL leachate confirmed no change in soil porosity relative to initial average of
0.34 as seen in the Fig. 4.7.2(8)(1). A photo of the ce11 1-2 terminated at early stage of
testing after 33 days is given in the Fig. A4.4, Appendi 4. The possible population of
EiAB plotted in Fig. 4.8.2(~)(1) has uniformly increased fiom average background 7.5 x
10' cfwg to - 2 x 10' cfu/g dry soil throughout entire thickness of both 5 cm plugs and is
comparable to the HAB size of - 1.5 x 10' cfug recovered fiom E 1 ce11 3 cm plug shown
in the Fig. 4.8.l(c)(l). SRB in the Fig. 4.8.20(1) indicate the overall higher count for the
plug from the ce11 1-1 terminated later (Le. after 49 days) than for the plug in the ce11 1-2.
Furthemore, the SRB count of - 2 x 106 c - g at the top 1 cm intedace with WAs
containing source solution flayer 1, as shown in the Fig. 4.7.2) is slightly (2.5-times)
higher than the count for the rest of the cell 1-1 (5 cm plug). Generally, the analogous
S R B trend was captured with the 5 cm plug inside the cell 1-2, terminated after 33 days.
The top interface (layer 1) as well as the layer 2 [see Fig. 4.8.2(d)(1)] had 10-fold
increased count of - (3 and 6) x 105 cfug, wMe the count in the rest of plug below was
-3.5 x 104 cwg, and nonot much ditferent from the background 9.5 x 10' cwg.
Measurement of ATP content confirrned increase in an activity of the viable bacterial
population, showing -average 61 ng/g readings throughout ce11 1-1 plug in Fig.
4.8.2(c)(l) çimilar to the - average 65 ng/g readings recovered from the 1-1 & 1-2
duplicates terminated first and discussed earlier. [Compare with Fig.4.8.l(c)(l). It is
noted that only ATP data fiom ce11 1-1 are available in the Fig. 4.8.2(c)(1) because the
cell 1-2 sarnple was used up for repeated extractions in chromatography.] It is noteworthy
that methanogenic aaivity was not detected in any of these "early terminated" tests. Once
again, the data for pore-water concentrations of the three VFAs, fonned distinct diffision
profiles for the 5 cm duplicates. These results plotted in Fig. 4.10.2@)(1) for acetate, in
Fig. 4.1 1.2@)(1) for propionate and in Fig. 4.12.2@)(1) for butyrate are in excellent
agreement with fit-lines simuiated with respective difision coefficients: 3.5 x 10"~. 2.5 x
1 0 " ~ and 1.5 x 10*1° (al1 in m2/s), inferred h m earlier tests (Table 4. l ) , and already
successfully applied for the set of 3 cm duplicates. This finding is reinforced by the Figs.
4.1 3.1 and 4.13.2, which show dissolved organic carbon cdculated From the three acids
(VFAs-DOC = L(12 x acidi[g/l] / acidi mol-weight}). As can be appreciated From the
bottom left graphs (b-1) showing the early termination data and fit-lines for the four
respective cells Pig.4.13. l (b)(l) and 4.13.2(b)(l)], there is little if any loss of VFAs-
DOC potentially attributable to biodegradation profile calculated for a diffision -10 2 coefficient in the recomrnended range Dm = (2.5 - 3.5) x 10 m 1s range. (See Figs.
A4.5.1 and A 4 5 2 for details.) Over-estimated values for the source solutions seen in
these graphs, are attributed to the variability in concentration caused by initial problems
with feed lines and peristaltic pumps.
As anticipated based on the observations from eariier testing possible population
and activity of indigenous soi1 bacteria noticeably increased within 23 to 49 days of
continuous exposure to preferential substrates (Le. VFAs), yet the consumption of the
substrates was not obvious and could not be measured within that short period of 23 to 49
days. Consequently, the rate of diffusive transport of the VFAs - substrates remained
unaffected by biodegradation. The evidence collected from the four ceils confirmed that
the diffusion coefficients inferred fiom the independent short term (batch) tests discussed
in 5 4.3.3 and given in the Table 4.1 are valid for the small scale even when conditions
become more favorable for degradation. Thus careful monitoring and sampling, during
the continuous supply of "VFAs-contaminant-substrates" coupled with short term
ancillary testing ($ 4.3) could be successfully used to delineate ditfusion and intrinsic
degradation in compaaed Haiton till at early stage of the long term testing.
As testing progressed the changes of the soil plugs and solutions became more
visible. Tiny sparse dots on soil ceIl contact surface grew in size to becorne irregular
shaped spots (2-5 mm diameter), either dark gray or markedly black, with tendency to
localize at the bottom i n t e h e with glass porous disc and receptor solutions. This thin
black ring ("black crust", noticed in the prelirninary tests described in Chapter 3) was
already fomed &er -40 days of testing around both 3 cm and 5 cm soil plugs. The top
surface of the plugs slowly started fading (becorning du11 gray), however the soil core in
the middle (i.e. between the top and the bottom) remained msty. The solutions were also
changing: the source appeared increasingly cloudy with traces of unpleasant mix of
"fishy" and "septic" odor, while the receptor remained transparent. but somewhat
darkened and with unmistaken HIS odor. B A R T testing was not done during regular
concentration vs. time monitoring, thus the estimates of cfi counts could not be made
accordingly. Nevertheless, not only did the odor point to bulk progressive activity of soil
bacteria but, combined with its locaiization, the odors signaled the dominant adivity of
particular genera and their metabolism. The source solutions (above the soil plugs)
developed a "fishy" odor charactenstic of various aerobic heterotrophs (such as
Pselïdomonas sp.; B A R F , User Manual, 1999). The source was constantly replenished
with new synthetic KVL leachate in order to keep constant concentration of the VFAs and
VOCs. The feed was chemically reduced in the preparation tank, however the compacted
soil plugs were initially oxidized. This could be conducive for heterotrophic ("eating
various food") aerobic and facultative bacteria (HAB), which are more aggressive and
more tolerant to subtle variations in oxygen and pH. The H2S (rotten-eggs) odor is
characteristic to the activity of sulfate reducing bacteria (SRB). Its generation in the
receptor solutions was a qualitative evidence of gradua1 development of reducing
conditions in the receptor solutions, which started up as oxidized (Eh = + 80 mV) and
were initially in contact with oxidized soil.
During routine sarnpling at 36 days, a 3 cm soil plug inside the ceil III-2 was
unexpectedly ruined due to development of gas pressure in the receptor solution.
(Problems with gas accumulation occurred earlier with preliminary tests, see Chapter 3)
The plug was lified from the disc and its top layer disintegrated into the solution, without
perrnitting the recovery of the data. The same incident almost reoccurred with ce11 III4 at
- 63 days, when another 3 cm plug was only displaced -2 mm upward, nevertheless
creating a gap (discontinuity to aqueous diffusion) and prompting test termination. The
top soi1 layer, -1 cm thick, at the contact with source solution appeared as having
increased moisture content and easily broke into chunks disturbing slightly the left-over
soil core. The results plotted in the Fig. 4.7.1@)(2) indicate a smail although noticeable
increase in porosity for al1 three layers cut fiom this sample. HAB counts recovered from
ce11 111-1 in the Fig. 4.8.1(~)(2) indicate - 4- and (marginal) 1.5-fold respective increase
in the source and receptor soiutions relative to the earlier times of 23 - 29 days [see graph
(c)(l) in the same Fig.] and unifonn and unchanged count of -108 cfu in the soil plug.
SRB in the Fig. 4.8.1 (d)(2) exhibited noticeable -30-fold increase in the top layer (layer 1
in the contact with source solution) and only siight (-4-fold) increase in the rest of the
plug, relative to earlier times of termination. At this stage of monitoring described trends
in bacterial count can be better appreciated fiom the graphs (a) and (5) for HAB and SRB
respectively, also show in the Fig. 4.8.1. It appears that within 63 days the two genera
have been growing at the expense of the available VFAs. This is particularly obvious with
SRB data plotted for receptor solutions in the Fig. 4.8.1(5), which seem to foliow
exponential (concave-up) pattern, while the SRB and HAB from other positions dready
exhibit saturation (concave-down) pattern. More importantly, BIOGAS-BART% have
reacted positive for methanogenesis after 20 days of delay in al1 three soil layers of the
cell III4 plug. ATP data in the Fig. 4.9.1(~)(2) mirrored the diffusion profile, with a
significantly increased (100-foId) ATP content in the top layer.
The changes noted with count and activity of bacteria in Halton till plugs were
reflected on the concentration of the VFAs as weIi. At 63 days the breakthrough of the
acids into the receptor solutions characteristic of diffusion dominated transport is evident
f?om the graphs (a) in Figs. 4.10.1, 4.11.1 and 4.12.1. The diffusion depth profile for
acetate in the Fig. 4.10.1@)(2) was still holding, but it indicated a surplus with a
possibility of acetate generation, when compared with the dotted fit-line simulated for
diffision (at D = 3.5 x IO-" rn2/s) without degradation. Propionate distinct diffision
profile in the Fig. 4.1 1.1@)(2), was in very good agreement with theoretical fit-line for D
= 2.5 x 10'" m'k, thus ruling out degradation. Butyrate profiles for receptor concentration
vs. time on graph (a) and for concentration vs. depth on graph (b)(2), and matching fit-
lines for diffision only, at previousiy inferred D = 1.5 x 10"~ m2/s in Fig 4.12.1, clearly
indicate losses attributable to degradation.
The observations gathered after 100 - 1 15 days of testing provided more evidence
on increasingly active indigenous bactena and consequent degradation of the VFAs,
accompanied with transformation of soil appearance. The soil plugs were becoming
visibly cracked particularly close to the top interfaces with the source solutions. Also a
distinctive change of soil consistency of those top layers characterired as formation of
"thinning" ancilor "caviar" like fluidized dark gray structure in place of initially
compacted soil core surface was taking place as time elapsed. M e r 117 and 118 days,
cells IV4 and 11-2 with a 3 cm plug were terminated. In both cells, the top fluidized layer
was -5 - 7 mm thick, blackened, swollen and cracked with tiny gas bubbles sporadicdly
popping up into the source solution. It seemed that this top loosened layer would have
easily dispersed had the cell been taped or accidentally moved. The middle of the soil
plug was msty and had visible although tiny (1-2 mm) cracks filled with (what appeared
like) liquid. The bonom of the plug had already recognizable 3-5 mm "black cmst",
indicative of microbially dnven sulfate reduction and consequent formation of metallic
sulfides. Both solutions had strong odor, the source mixed but dominantly "septic" and
receptor more distinctly "rotten eggs" @-Ils). Results plotted in the Fig. 4.7.1 (a) & (bl
indicate an increase in porosity in the 5 mm thick layer 1 (induced by increase in rnoisture
content and bacterial activity at the top interface). Porosity of the soil plugs below the
expanded top interface remained unaffected relative to the initial reading
[(Fig.4.7.l (b)(3)]. Although the soil was Iiterdly "fermenting" the active HAB and SRB,
seemed to be entenng the stagnant phase, generally without conclusive increase in cfu
the Fig. 4.8.1. Bacterial counts in source and receptor solutions seemed to level off as
weil. Moderate increase in soil ATP content ploned in the Fig. 4.9.1 (c)(3), generally
higher for the ceil 11-2 than for the ceil TV-1, seems to indicate that activity of viable
bacteria is virtually the same at the top and at the bottom and only slightly lower in the
mid 1 cm of the particular plug. Slight increase (- 2-5 - fold) is noticed in the source
solutions at this time (of -1 17.5 days). General trend for ATP seen in the Fig. 4.9.1 (a) is,
however sirnilar to the trend developed by KAB and SRB in the Fig. 4.8.1 (a) and 0). It is
also noted that methanogenesis was detected again within - 20 days of BIOGAS-
B A R T incubations. For the fist time d e r - 117.5 days of testing both source and
receptor solutions (in addition to al1 soil layers, first registered with ce11 III4 &er 63
days) tested positive, showing that conditions for the syntrophic association of
methanogens with more aggressive SRB were gradually being established.
Data shown in Fig. 4.10.1(%)(3) recovered from the pore water indicate a surplus
of acetate, implying that it was being generated in the system. These extra amounts of
acetate, [showing even slightly above the receptor breakthrough long-dash curve on the
graph (a)], is considered to have onginated from the breakdown of propionate andfor
butyrate (Zinder, 1993), but some extra amounts might be even corning kom
"hornoacetogenesis" [autotrophic reduction of CO2 with hydrogen (Lovley & Klug, 1983;
Phelps & Zeikus, 1984)). This "surplus" of acetate in Halton till plugs suggests that rate
of fermentation (breakdown of propionate and butyrate) noticed at 117.5 days of testing
might be exceeding the rate of anaerobic respiration (consumption of H2 and other
fermentation produas by terminal trophic bacteria), as observed by McMahon &
Chapelle (1991). Propionate data for both IV4 and II-2 plugs in the Fig. 4.11.1@)(3),
however form consistent diffusion profiles which are stili, (Le. &er 117 and 118 days) in
excellent agreement with D = 2.5 x 1UL0 m2/s-simulated fit-lines. Thus for propionate,
degradation cannot yet be discemed fiorn diffusion. Butyrate data for both 3 cm soil plugs
s h o w in Fig. 4.12.1@)(3) are unmistakably aligned into distinct diftision profiles with
readings only skghtly lower than predicted by the D = 1.5 x IO-'' n2/s theoretical (dotted)
lines with no degradation. However, the receptor data for butyrate vs. time in the graph
(a) clearly show loss, which given the cirnimaances, can only be associated with
biodegradation. When the three acids are re-plotted as VFAs-DOC in Fig. 4.13.1 (a) &
(b)(3), the level of intrinsic degradation remains as inconclusive as it was at the earlier
time of 63 days, with cell III-1. Data in the receptor solution seem to be encompassed in
the range of & = (2.5 -3.5) x 10'1° m2/s which seems reasonable and acceptable given
the scatter, scde of VFAs concentrations and capabilities of analytical methods used. (For
differences in numbers and curve statistic, see details in the Fig. A4.5.1, Appendix 4.)
Data for the pore-water, expressed as DOC recovered from total measured VFAs
concentration in the Fig. 4.13.l@ (3) indicate excellent agreement with D = 3.5 x 10'1°
m?s diffusion dotted fit-line at -1 18 days. Thus satisfactory account of total mass could
be obtained, with observed "surplus" of acetate included, without taking into account
degradation at -1 17.5 days. Nevertheless, the steady activity of SRI3 as wel1 as
methanogenesis, combined with receptor profiles for butyrate provides qualitative
evidence of fermentation and mineralization. Bearing in mind that the amounts of acids
(and consequently VFAs-DOC) applied to the indigenous microorganisms are substantiai
(glL), this lack of measurable decline could be viewed as a load-imposed lag to
degradation. In addition, it is known (Chapelle & Lovley, 1990) that the rates of organic
carbon metabolism (usudly originating frorn p M or low mM amounts of M A S ) in deep
aquifers and clayey confining beds are extremely slow (likely c than IO-' - 104 mM C
lLlannum), yet these sediments remain metabolically active generating CO2 a d o r
methane. It appears that, in the intrinsic degradation experiment simulated herein, the
removai of the VFAs tested could not be measured and discemed fiom difision even at
such late stage, while the generation of fine gas bubbles which could only be released
fiom the fermentation of these acids, could be seen by naked eye.
The remaining two cells from the 3 cm soi1 plug-assernbly were terminated d e r
162 and 163 days respectively. Both II4 and N-2 cells are shown in the Fig. 4.6. Source
solution had pungent odor, "septic" prevailing with nrong "rotten eggs" (H2S) smell,
pointing to graduai development of anaerobic conditions and dominance of various
enteric (e.g. colifomis; capable of fermentation and gas generating, B A R T user's
manual) bacteria, as weU as hcreased activity of sulfate reducers (SRB). Receptor
solutions had similar, fou1 decaying odor with "rotten eggs" smell (and possibly SRB)
consistently prevailing. The rusty remains of the plug in the ce11 II-1 appeared ''cornpacf'
and less cracked than the "almost broken" plug in the cell N-2. The top soil layers (iayer
l), particularly the one in the ce11 IV-2 expanded (and appeared thicker) in the course of
elapsed 45 days (relative to previous termination of ceUs IV-1 and JI-2 after 118 days),
however when the source solution was drained, this fluidized layer settled to - 5 mm
oniy. Estimates of the porosity for these two plugs in the Fig. 4 . 7 . 1 0 and @)(4) indicate
drastic increase in layer 1 (from 4 . 3 4 to 0.62) for ce1 TV-2, while the porosity increase
in the ce11 11-1 was only moderate (also see Table 4.3). Generally, there is noted increase
in porosity in both plugs (i the Fig. 4 - 7 4 which is considered to be a direct
consequence of VFAs fermentation and generation of gases (Hz, CO2 and C%) associated
with terminal degradation processes (Chapelle, 1993). [It is noted that the dominant pore
size in the untreated compacted Halton till was -0.1 p. Generally, such small pore
"throat" could restrkt growth and transport of the soil bacteria, typically having sirnilar
(0.2-0.4 p n ) size in an oligotrophic environment (Fredrickson et al., 1997), such as
natural aquitard.
It is believed that initial adverse conditions for microbial metabolism changed (Le.
the reported increase in bulk porosity was accompanied with a development of a larger
dominant pore size) as a consequence of gas generation. Mercury intrusion porosimetry,
however was not performed with the treated soil samples, and quantitative account of the
resulted pore size could not be given for a cornparison.] With increase in porosity. there is
a slight increase in HAB and particularly SRB count (IO-fold) in the top Iayer(s)
indicating maxima for both groups, as can be seen in the Fig. 4.8.1 on graphs (a). fi).
(c)(4) and ( i ( 4 ) . Nevertheless, the saturation pattern observed with the count of
indigenous microorganisms in al1 positions throughout the soil cores still holds. ATP
content for the these two plugs in the Fig 4.9.1(~)(4) shows significant increase (-10-fold)
in the top layers, while the rest of the plug remained unafTected in time. Furthemore, the
BIOGAS-BARVs with the soil sampled from layen 1 from cells TI-1 and IV-2 reacted
afier ody two days of incubation, indicating that the Halton till methanogens became
quite aggressive at this (final) stage of testing. Also such fast response clearly points to
the faa that top interface become highiy reduced, with hydrogen levels conducive for
methanogens [Eh < -0.3 V, H2 z (10 - 40 ngL or 7 - 20 NM)]. However the environment at
163 days had been (and remained as such) conducive for sulfate reducers, thus becoming
the locus of very intense syntrophic activity of the two hydrogen utilking bacterial
groups.
The co-existence (syntrophism) and thriving of sulfate reducers and methanogens
has been observed in natural anoxic sediments (Senior et al., 1982; Phelps & Zeikus,
1984) where the slow and steady decay of cornplex organic matter takes place. Since
these "terminal" bacteria consume H2 pelieved to be produced by heterotrophic
fennenters andor acetogenic obligate-proton reducers (Le. propionate and butyrate
degraders, as well as fermenters of more cornplex organic substrates found in landfills
waste or organic detritus)] as well as acetate (i.e. aceticlastic methanogens and sulfate
reducers)] to produce H2S, CO2, C h it was speculated that such (terminal) gas generation
would eventually reveal the decreasing amounts of available VFAs in the tested solutions
and Halton till pore water.
Concentrations in pore water plotted for the last terminations of the two 3 cm soi1
plugs indicate loss of acetate in Fig. 4.10.l(b)(4), loss of propionate in Fig. 4.1 1. IO(4)
as well as unmistakable loss of butyrate in Fig. 4.12.1@)(4). Acetate degradation was
manifest oniy in the pore water [as seen h m the colleaed profiles in Fig.4.10.l(b)(4),
while the receptor solution profiles in Fig. 4.10.l(a)] still have the "generation" (Le.
increasing) trend with oniy the final two points that suggest decline. The propionate pore
water profiles in Fig.4.11.1@)(4), also indicate significantly lower readings than predicted
by "difision-only" dotted lines. The propionate readings in Fig.4. I l . l(a) for receptor
solutions however, form declining patterns, more obvious than those collected for acetate.
This is to be expected because both facultative heterotrophs and sulfate reducers,
(obviously thriving in the fermenting Halton till, see Figs. 4.8.1) could cany out
anaerobic oxidation of propionate and butyrate to acetate (Dolfing 1988; Widdel, 1988).
The same observation holds for butyrate concentrations for these final terminations at
-1 18 days: both pore-water and receptor solution readings are significantly lower than
those predicted by dotted difision lines. Butyrate is degrading and some acetate could be
generated as a breakdown product as well. It is noted that the pore water data for the three
acids look somewhat erratic with noticeable scatter and patterns which do not form
difision profiles as clearly seen on o(4) graphs in the Figs.4.10.1, 4.1 1.1 and 4.12.1. It
is considered that very Little of the recorded scatter could be attributed to analytical error,
since the most of the scatter appeared upon intenssed degradation [Compare the (b)(4)
graphs with the (b)(3) and (I>)(2).] The observed patterns are to be expected given the
uneven rates of butyrate, and propionate breakdown to acetate, and the rates of acetate
breakdown to gaseous products, in addition to the possibiiity of homoacetogenesis, aii of
which could take place in such a complex mini-ecosystem (e-g. see Lovely & Klug,
1983).
When the measured acid concentrations are plotted as DOC, in Fig. 4.13.1, some
of the scatter diminishes and the degradation trend becomes more obvious. Depth profile
in Fig. 4.13.1 (b)(4) plotted fiom averaged readings for both 11-1 and IV-2 3 cm plugs
indicates somewhat uniform DOC distribution dong the plug thickness with tendency to
equilibration with the DOC concentration in receptor solution. Unlike the depth profiles
collected for earlier times in graphs (b)(l), (2) and (3) when degradation was not
effective, this last [(a)(4)] profile at -163 days does not have distinctively diffision
pattern. The profile collected for the receptor solution concentrations in time in the graph
4.13.l(a) is however quite characteristic for difision dominated transport and the data
points remained within the proposed DOC difision bounds of (2.5 - 3.5) x 10"' m'A
almost until the very last sarnpling at the times of termination. Thus, the results of the
final stage of the experimentation with the 3 cm plug ce11 assembly revealed that intrinsic
degradation of the M A S tested proceeds very slowly becoming obvious (and probably in
effect) only in the soil, or rather in the pore water afler - 163 days of testing, while the
receptor solutions, with exception of butyrate, remained only marginally affected wit h
degradation. This finding is somewhat different fiom one reported by McMahon &
Chapelle (1991), who showed that organic acids transported by diffision from clayey
(organic rich, thus not oligotrophic) aquitard into the aquifer, mainly degrade in the
aquifer. Although the prospects of degradation in the receptor solution (hypotheticd
aquifer) could not be ruled out, it is certain fkorn the experiments that VFAs can degrade
in the compacted Halton till, that is in the (initially oligotrophic) aquitard, as well.
The same findingç were generally reproduced with another testing assembly
which ernployed 5 cm thick Halton till plugs (see Fig. 4.5) and served as a replicate
intrinsic degradation experirnent. Generally the soil plugs have gone though the same
transformation described earlier, however it seemed that the slight increase in plug
thickness and diaision dominated (slow) spreading of VFAs resulted in sornewhat
delayed reactions. [It is noted that the two testing assemblies had been running almost
simultaneously (the one with 3 cm plugs, was started just 12 days later than the assembly
with 5 cm plugs).].
The routine testing continued with termination of the ce11 3-1 at 128 days.
Although significant time had elapsed, the soi1 plug remained compact and dominantly
oxidized with very little visible activity at the top interface (compared with quite vigorous
layer 1 expansion observed with 3 cm plugs in cells II-2 and IV-1 terminated only 10 days
later). Only the charaderistic blackening ("black cnist") appeared around the perimeter at
the bottom interface with the soil plug and disdreceptor solution. There was virtually no
increase in porosity for the entire 5 cm plug as show in the Fig. 4.7.2(3)(2). However the
increase in bacterial cfu count is evident: significant at -3680-times for HAB (fiom
background 9.5 x 104 to 3.5 x 10') in the Fig. 4.8.2@)(2) but only marginal at - 2.4-times
for SRB (from 7.5 x 104 to 1.8 x 10') in the Fig.4.8.2(#)(2). BIOGASTM-BARTs with
samples from two top soil layers and source solution reacted positive after - 25 days, thus
confirming that the conducive conditions for methanogens were developed as well. A
slight increase in ATP content is also recorded for this 5 cm soil plu& forming diffusion
depth profile in the Fig. 4.9.2(~)(2) in contrast to uniform distribution for both HAB and
SRB. It is speculated that at this stage of incubation (- 2 months with 3 cm plugs and
seemingly 3 months with 5 cm plugs) soil microorganisrns have been growing at the
expense of VFAs, and available carbon has been used for biomass synthesis rather than
conversion and mineralization (gas production) of VFAs.
Concentration profiles collected for acetate in Fig. 4.10.2(a) and o ( 2 ) and
propionate in Fig. 4.11.2(a) and @)(2) are characteristic for difision and are in excellent
agreement with D = 3.5 x IO-'' rn% and D = 2.5 x IO-" m2k, as iinferred earlier for each
acid respectively. Profiles for butyrate at 128 days, however do not exhibit clear
breakthrough in the receptor solution indicating degradation loss, which is even more
evident with pore-water data pattern. Thus, again with the 5 cm replicate, it is confirmed
that butyrate starts degrading the first (before the other two acids) and with careful
monitoring it becomes possible to discem degradation frorn diffusion. It is also noted that
acetate "surplus" in pore water profile did not appear with this 5 cm plug [Compare with
recorded acetate generation in the Fig. 4.10.1 (b)(3) with 3 cm soil plugs upon termination
of cells IV4 and 11-2 after - 117.5 days.] Data for ali acids plotted as DOC in Fig.
4.13.l(a) and /b)(2) f o m distinct diffusion profiles and conform to the D = (2.5 - 3.5) x
10-l0 mL/s bounds previously simulated with 3 cm plugs.
The next termination was afier - 202 days with cells 4-2, 3-2 and 2-2. By this tirne
the top soil layers at the interface with receptor solution were aiready expanded and
reduced with gas bubbles popping up. For details regarding plug transformation in the ce11
4-2, see Fig. 4.14, and compare it with Fig. 4.5.
As a consequence of the enhanced rnicrobid aaivity at later times, the porosity of
the top layers for the three temiinated plugs increased markedly from initial 0.34 to 0.55
(average), as seen in Fig. 4.7.2(u). The porosity in the remaining core of the 5 cm plugs,
was unaffected. The HAB count increased (-10-fold) but only in the top layer of the ce11
4-2, as shown in Fig. 4.8.2(a) and (c)(3). The increase in HAB count was not recorded in
the rest of either "4-2" or "3-2" plug relative to the earlier time (128 days), while ce11 2-2
had even lower count (4.7 x 10j cfu) than the plugs in the cells 1-2 and 1-1 (2.2 x 107cfu)
terminated at the very beginning of this intrinsic degradation expenment. The KAB
growth in these plugs seems to be reaching saturation, as noticed before with the soi1 fiom
3 cm plugs, however the reason for the HAB decline in ce11 2-2 is not obvious. The SRI3
count for cells 4-2 and 3-2 in Fig.4.8.2(4(3) increased noticeably (-100-fold in Layer 1,
- 36- fold in Layer 2) relative to the earlier time, unlike the count for cell 2-2.
Methanogens exhibited the parallel activity trend, top layers reacted positive within 3 to 4
days and source solutions followed d e r 10-12 days for the cells 4-2 and 3-2 respectively.
The rest of the rnid (rusty) layers reacted partially d e r 30-45 days, while the soil samples
fi-om ce11 2-2 had not reacted within 60 days. ATP content distribution in Fig. 4.9.2(~)(3)
was consistent with HAB and SRB trends, top layers in cells 4-2 and 3-2 and even in ceIl
2-2 had higher (-10-fold) activity than the layers in the rnid core. Generaily, the data for
HAB, SRB and ATP were showing saturation trend for the 202 days [see graphs (a) and
fi)], as already noticed with 3 cm soil plugs discussed earlier. Analysis of organic carbon
content performed on the rernains of the treated soil indicated general increase in f, [%]
fraction in the dry Halton till with time, which is supposed to be parallel to the changes in
carbon content induced by microbial growth, as illustrated in the (d) graphs in both Fig.
4.9.1 and 4.9.2. The increase in f, could potentially be attributed to increase in rnicrobial
organic carbon as weii as to formation of new soil organic matter (and C) upon biomass
decay .
Results on concentration of the VFAs collected on the 5-cm plug cell assembly,
however show some features different Eom those observed with the 3-cm ce11 assembly.
Data on acetate plotted in Fig. 4.10.2@)(3) for pore water, although forming characteristic
diffusion profile clearly indicate loss, (Le. systematic depamire fiom D = 3.5 x 10'~~rn'/s
"diffision-only" fit-lines). Because of unpredicted andyticd problems, the VFAs pore
water data for ce11 2-2 could not be recovered and are not available for display. This trend
is becorning evident even in the receptor solutions in the graph (a). Thus, there is no
visible acetate generation and accumulation in the soil plugs after - 202 days of testing in
cells 4-2 and 3-2 plugs, as opposed to trend observed with 3 cm plugs. Fermentation of
propionate and butyrate to acetate as well as eventual homo-acetogenesis seerns not to be
degradation rate limiting (McMahon & Chapelle, 1991; Chapelle, 1993), and since there
is a decrease in acetate it could be foreseen that SRI3 and methanogens driven terminal
anaerobic respiration would proceed as well. Propionate in Fig. 4.11.1 and butyrate
concentrations in Fig. 4.12.1, however fom distinct diffusion profiles both for receptor
solutions in graphs (a), as well as for plugs pore water in graphs (b)(3). Data satisfactorily
conform to dotted theoretical fit-lines, thus indicating that neither propionate nor butyrate
degradation could be differentiated From difision after - 202 days of testing with 5 cm
Halton till plugs. DOC data calculated fiom total acid concentrations indicate only slight
departure of the pore water data (averaged for clkty) from the fit-line in the Fig.
4.13.2@)(3) as well as equaily slight, yet still inconclusive likelihood of degradation in
the receptor solution on the graph (a) in Fig.4.13.2.
The remaining 1st two ceiis with 5 cm plugs were terminated after - 271 days. As
expected, the porosity in the top fluidized layen increased further due to gassing, as seen
in the Fig.4.7.2. HAB were unifody distributed in the NO plugs in Fig. 4.8.2(~)(4), but
increased - 10-fold oniy in the ce11 4-1 plug. SRI3 in the Fig. 4.8.2(4(4) seem to be
unchanged relative to the earlier time of -202 days. Tt could be stated that soi1 indigenous
population does not visibly grow and it has likely reached steady state under the constant
and (unlimited) supply of organic (and inorganic) nutnents. Methanogenesis was
detected, however the reaction was not more "aggressive" than at the earlier time of 202
days. The ATP content in Fig. 4.9.2 is only slightly higher than before and at its maxima
for the 5 cm soil plugs. However it seems that the overd activity of the viable biornass in
the - 1.5- 1.8 cm top expanded layers is significantly lower (- 5 times) in the 5 cm plugs
afler 202 and 271 days (- 450 nglg) than ATP recordeci in 0.65 cm (oniy) thick expanded
top layers formed in 3 cm soi1 plugs d e r 163 days (- 2360 nglg), as cm be appreciated
f o n Fig. 4.9.1(c)(4) and Fig. 4.9.2(c)(3) and (c)(4). It is not known as to what could have
caused such unexpected decrease in ATP content in the top layers with 5 cm soil plugs,
particularly in view of the increased thickness of the expanded layer dunng longer penod
of testing under continuous supply of M A S (27 1 relative to 163 days).
It is considered that the graduai increase in porosity (coupled with increased
thickness) of the expanded top layer is parailel to increased activity of indigenous soil
bactena and consequently to rate of VFAs @OC) degradation. The small pore size
(initially for Halton till 0.5 pz range) and low effective porosity associated with the
natural clayey confining layers are often listed as imposing constraints to rnicrobial
activity (Chapelle & Lovely, 1990; Fredrckson et al., 1997), thus the beginning of VFAs
degradation observed both with 3 and 5 cm Halton till plugs, could be attributed to the
relief of this "physicai" constraint rather than to the hi& count of ferrnenting and HZ-
utilizing bacteria. Indeed, the most prominent increase in count of HAB and SRB and
ATP content occurred at early stage of incubation d e r only - 30 - 50 days [graphs (a),
(5). fc)(l) and (dj(1) in Figs. 4.9.1 and 4.9.2 and 4.10.1 and 4.10.2, respectively], whiie
no increase in porosity in the top layers was deteaed. At the final times, the readings for
HAB in the "layers 1"seem to be stable at - 8 x 108 cfu, at 163 days (top 0.65 cm in 3 cm
plugs fiom cells II-1 and N-2), and -7 x 108 cfu at -202 days (top 1.5 cm in 5 cm plugs
from cells 4-2, 3-2 and 2-2), dropping only marginally (- 4-times is negligible given the
analytical method) to - 2 x 10' cfu at 271 days (top 1.8 cm in 5 cm plugs from cells 2-1
and 4-1). SRI3 count however do seem to drop rnarkedly (- 10-times) earlier, Le. from
-1.1 x 10' cfi at 163 days to - 1.2 x 10'cfi at 202 days and remain unchanged (-1.3 x 10'
c f i ) till the end of test at 271 days. This trend with SRB count matches the observed drop
in ATP content and biornass activity, thus is could be speculated that the two mi& be
related. This drop in SRB number could be related to local depletion of sulfate pool in the
soil pore water and increase of hydrogen partial pressure (concentrations), thus shifling
the regulation of terminal carbon and hydrogen flow to methanogens. Not enough
evidence could be gathered on every aspect relevant to the intrinsic degradation processes
during the course of this quite complex experimentation: the srnail scale and necessity to
keep the system as air-tight as possible precluded the monitoring of gases and electron
acceptors pool in time. Nevertheless, it is possible that the large sulfate pool (-5 g/L in
pore water) becomes exhausted under initially dominating and thermodynamicaily more
favorable sulfate reduction. Since the level (and constant supply) of bicarbonates (- 5
glL) in the synthetic leachate is higher than sulfates (- 0.15 glL, designed to mimic the
contents and metabolism of these electron acceptors in the field), it is expected that the
methanogens will eventually prevail over sulfate reducers, at least on the tested scale with
- 220-500 g soi1 plugs.
The pore water concentration profiles for ail three of the acids indicate
unmistakable loss due to degradation with characteristic erratic pattems along 5 cm plugs,
as seen in Fig. 4.10.2 for acetate, in Fig. 4.1 1.2 for propionate, in Fig. 4.12.2 for butyrate
as well as for DOC data in the Fig. 4.13.2. Receptor solution patterns in these figures still
have characteristic difision breakthrough trend, with only the last points at final times
reading lower than predicted by diffusion only. This however, gives again unequivocal
evidence that intrinsic degradation dominantly takes place in Halton till.
More importantly at this late stage of testing, it seems that quite high organic
contamination @OC = 2.4 g/L, or COD = 17.4 g/L), although considered readily
degradable, is finally being rernoved by indigenous soi1 microorganisrns without culture
enrichment and under adverse mass transfer conditions. This now active population
appears stable at - (7.5 x 10'- 8 xlo8) and at (-1.5 x 10' - 1 x 10') cfiugfor HAB and
SRB respectively, capabie of M A S fermentation and methanogenesis. Thus intrinsic
degradation of VFAs is finally taking place in, what began as rusty cornpacted clay plug
interface and turned into dark gray "fluidized caviar-like bioreactof' (see Figs. 4.5 and
4.6). In theory, it was possible to envision such an outcorne, based on terminal
degradation processes taking place in eutrophic lake and marine sedirnents nch in sulfates
and bi-carbonates as described in Nedwell, (1984) and Senior et al. (1982), where natural
detntus from plants and anirnals slowly breaks down to VFAs and is finally mineralized.
In the case simulated here, (and in landfills), the organic substrates came from waste
which starts fermenting and releasing VFAs into the naturai soil with quite high sulfate - 5 g L in pore water, which is comparable to some marine sediments) and bicarbonate (- 1
g/L in pore water, even higher in leachate) content. The same rusty oxidized and clay rich
soil was the source of active bacterial groups and its indigenous rnicroorganisms had
developed the obviously advantageous abiiity to ferment VFAs and respire on electron
acceptors other than oxygen (Chapelle, 1993, Kolbel-Boelke, 1988).
An attempt was made to mode1 the impact of VFAs and infer the likely
degradation rates under the dominant diffisive transport as simulated in the presented
intrinsic degradation experiments. Degradation of acetate as such, is not considered
because of the complexity associated with its simultaneous generation fiom propionate
and butyrate. Separate tests with propionate and butyrate as dominant readily degradable
carbon source were not performed, thus the rates of their conversion into acetate
potentially applicable to the tested scenario were not available. Butyrate degradation is
attempted because it is considered that under the test conditions it is very likely that no
measurable quantities of butyrate could have been generated in the system, thus any
interferences fiom such process would not bias the estimates of butyrate bulk anaerobic
removal in clayey soil. [The possibility of formation of higher fatty acids from Ci and C2
compounds is acknowledged (Dolfing, 1988)l.
Based on observation of increased rnicrobial activity with the tested soil plugs, it
was assumed that degradation would also be localized and dominant at the two interfaces
where the nutnent gradients and mass transfer limitations exist. A simple approach
assuming first order decay reaction at the top source solution/soil interface (Le. "layen
1") with gradua1 change of porosity in time as given in the Table 4.3 and diffusion
coefficients fiom Table 4.1, was initially considered. This simulation performed with
compute program POLLUTE v6 (Rowe & Booker, 1999), with half-life of 1 day, resulted
in excellent fit for the pore water profile only at the early time of 63 days [solid line Fig.
4.12.1 (b)(2)], while the rest of prediction was not successful, as seen for the later times of
1 18 and 163 days simulated in Figs. 4.12.1 @)(3) and 4.12.1/0)(4) for the cells 11-2 and
IV- 1 respectively, and receptor solution profile in Fig. 4.12.l(u), dl ploaed as solid lines.
The fit for the receptor solution and h a 1 pore water profile was improved when
degradation in the bottom soil layer with a 5 day half-life (bottom interface with plug and
receptor solution) was considered, as plotted in dash-dot iine in the sarne figures. Equally
improved fit was, however generated with the degradation simulated in the receptor
solution (at 30 day half life in the receptor, after 50 day lag, as given in the legend in Fig.
4.12.1). instead of degradation in the bottom layer. Regardless, of the good fit iines it is
considered that dominant degradation in these tests took place in the soil. It is noted that
no consistent prediction for butyrate degradation could be achieved for the 3 cm plug at
1 18 days. The same stands for the overall butyrate degradation modeling attempt with 5
cm plugs, given in the Fig. 4.12.2. It seems that there is virtually no measurable
degradation (or certainly less then observed with 3 cm plugs) until the late stage of 270
days, at which point the rates of removal are different, and highly variable for the two
terminated plugs.
Modeling of degradation of VFAs, expressed as DOC (dissolved organic carbon),
was more successful. The same approach was adopted regarding the locus of the
dominant degradation activity and gradua1 change in porosity that was recorded in time.
For the 3 cm soil plugs it is assumed that after 140 day lag degradation prevails in the top
(expanded) layer at a fast rate, (Le. half-life of 0.75 to 1 day). This resulted in reasonably
good fit to the collected data, as show with solid lines plotted in Fig.4.13.l(a) for
receptor solution and in 4.13.1(b)(4) for the latest (and last ) pore water profile recovered
at 162.5 days. (Data for the close terminations were averaged for clarity.) Modeling of the
DOC degradation was successfully reproduced for 5 cm soil plugs. Somewhat longer
half-lives of 5 days in the top soil layer and lag of 180 days resulted in very good match
in the Fig. 4.13.2 for the collected data points and fint order degradation rate under
dominant difisive transport of DOC in compacted clay plugs for pore water profiles at
202 and 270 days of testing on graphs (b)(3) and (4) and receptor solution profile on
graph (a). The details pertaining the changes of the parameters are given in the legend in
the Figs. 4.13.1 and 4.13.2 and in Table 4.1 and 4.3.
It is recognized that the approach used to predict an impact of VFAs-DOC
(concentrations) in the tested system, relies on certain simplifications, which rnight not
hold in natural large-scale system. Bio-degradation of VFAs-DOC is modeled as first
order reaction, implying CO-metabolisrn but without ident&ng the underlying
mechanism of biodegradation. At the sarne the , it was evident f?om the tests that
indigenous rnicroorganisms grew at the expense of the degraded VFAs substrates. The
rate of the microbial growth in these compacted plugs could not be assessed with
certainty due to its complexity and dependence on the system itself It seems that growth
follows logistic-like saturation pattern, with maximum at early stage of incubation when
there is no virtual (and measurable) removal of the growth substrate at all.
Notwithstanding the significance of various rnetabolic mechanisms which govem the
intnnsic degradation and necessity to account for their nurnerous parameters in pursuit for
mathematical models that descnbe the processes better, it is only possible and it suffices
at this initial stage of laboratory research to put an ernphasis solely on rates of removal of
substrates-contaminants and predict contaminant impact accordingly. No attempt was
made to mode1 the degradation using a program, which could account for growth of
indigenous rnicroorganisms on a particular substrate (e.g. Monod) and couple the growth
to the substrate rernoval, because such simulation requires more parameters to be known
in advance. In view of small scale tested and only a few data points collected even in the
laborious expenments presented herein, it is considered that such complex simulation
would not provide an additional utility to proposed modeiing approach.
4.5 Summary and coaclusions
Long terni diffision dominated tests were perfomed for sixteen (16) laboratory
compacted clayey soii plugs exposed to continuous supply of synthetic leachate
containing an effectively unlirnited amount of volatile fatty acids (VFAs). The results
indicate that significant microbial activity develops upon exposure of the soi1 indigenous
microorganisms to these readily degradable substrate-contaminants. The growth of
selected groups of rnicroorganisms, such as HAB (heterotrophic aerobic bacteria), SEü3
(sulfate reducing bacteria) and methanogenic bacteria carrying out fermentation and
terminal mineraiization process of the VFAs became evident after 30 to 50 days of testing
reaching maximum with (2 - 8) x 10' cfulg and (1.3 - 11) x 10' cfu/g for HAB and SRI3 in
the soil layer at the interface with the source of organic and inorganic nutrients.
Regardless of this rapid and vigorous growth, the consumption of VFAs was small and
the measurable degradation of the VFAs did not occur until d e r a somewhat long lag of
140-1 80 days. It is considered that this lag of othenvise readily degradable organic
compounds (such as VFAs) persisted due to very high initial concentration of these acids
(2.4 g.2 as DOC, or - 17.4 g/L as COD) applied to carbon starved soi1 rnicroorgankms.
This biodegradation lag is enhanced even more with the mass transfer limitations imposed
by compaction (designed to provide the lowest practical hydraulic conductivity) and
srnall pore size of the clay sediments. Once the significant amounts of gas were generated
fiom fermentation, conditions developed for improved mass transport and exchange (Le.
better contact and rnixing) of the nutrients and bactena and the outcome of the intrinsic
degradation was more apparent in the upper cm of the soil. The manifest breakdown of
VFAs that followed after the lag, was localized at the top soil interface and source of
nutrients, and is characterired by very fast rate and consequently short half-Iife of 0.75 to
5 days, simuiated for DOC (total VFAs as dissolved organic carbon).
Based on these expenments it became evident that compacted clay soi1 is a source
of microorganisms capable of ineversible and teminal degradation of organic
contaminants. M A S tested are representative of major organic pollution found in
municipal solid waste (MW) landfills, and compacted clay liners designed to act as
diffusion bamers to the transpon of this pollution are placed at the very contact with
pollution that escapes fiom leachate collection systems. Thus is appears that degradation
and further reduction in the overail organic pollution by the indigenous microorganisms
present in soil could potentially occur in these liners and natural confining layers alike,
even under the most adverse mass transfer conditions and without any man-aided
intervention. It is also recopked that this ability of the indigenous microorganisms to
carry out mineralization of fermentation intermediates, particuiarly acetate might be
crucial for the anaerobic degradation of more complex compounds found in contaminated
soils and aquifers.
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Table 4.1 Summary o f diffusion and linear sorption coeiricients for the VFAs and Halton Till used for modeling
Acctic (etharioic ) acid
Propionic (propanoic) acid
Butyric (butanoic ) acid
Dissolved orgnnic cûrbon (DOC)'
D, Diffusion coetriciciii in soi! pore waier. [in%) (lab simulations)
Likcly' B, lid1lhC
x 10-
2.5 - 5.0
2.0 - 5.0
1.5 - 3.0
2.5 - 5.0
D,,, Diffusion coellicicnt i: gl,ms
disk, 1 m'/s] '
' Likcly range based on conditions similar 10 thor examincd
' sanie as diffusion coçflicienî in the frce salution, values taken from Yaws (199 1);
pK, R = 1 + - , caiculaied for dry dciisity p = 1.80 glcni2 and avcraged porosity 11 = 0.34 for Halton till cornpacted in laboratory; Il
4 »OC clilculiiicd üs total orgmic ciirboii: WC' = ( (coriceiitriition of a coiiipurid, g/L1 x 12.00 1 ) / iiiolccular wcighi of a coiiipound (Hunslow, 1995)
Table 4.2 Composition of Synthetic KVL Ieacbate (KVLL)
Component
kneral Indicators 10D 'SSNSS )RI' [mV1 'emperature, TC]
I-1 (-log H3
Jolatile Fatty Acids (VFAs)
[norganic Nutrients
Trace Metal Solution NaOH (pH adjustment) 3% w h NazS x 9H4,0 (Eh adjustment)
Volatiie organic chernicils (VOCs): Dichloromethane @CM) 1.2-DicNoroethane (1.2-DCA) Trichloroe thene (Ta) Benzene Toluene E thy 1-benzene Xy lene isomers
Table 4.3 Change of soü porosity iaduced by intriosic degradation of VFAs
I I "3 cmw thick compacted clay plugs I 1 I 1 1 I
Terminated 1 1 1 1 1 after [days]
Loyer 1 Laytr 2 ïaycr 3
C h u y e of panmeten 1 1
II-1
162
modcled over the pcr~od Porosfty o f
I I "5 cm" thick compacted clay plugs I
IV-2
163
0.47 0.38 0.35
23
Laytr 1 Difi ion cocfRdcnt
n h n h
0.62 0.43 0.43
0.80 1.40 0.60
ni
0.32 0.33 0.32
O - 55 days 0.34 (initiai) or as caîculltcd
0.50 1.20 1.00
h'
0.90 0.80 0.80
29
m2:s
I n . [-1, total porosity assuming 100 % saturation, dculated fiom measured gravimetnc moisture content. W. and speciflc gravity, G, (2.75 for Halton till), as: n = wG((1 + wG,)
n
0.34 0.31 0.30
55 - 140 days 0.34 or as crlculared
îhe pciiod Pomity of
'h. [cm], measured thickness of the layer cut Born the soi1 plug at the time of termination
1.00 0.90 0.6
63
L40 - 163 days 0.55 or as calculatrd
m'fa
Ce11
TcnnLi1Ptcd aRtr [dani
Layerf Layer 2 Layer 3 Laycr 4 Layer S
I
0.40 0.40 0.31
m'f s
1 0.34 (Uiitial) or u calculalcd
0.30 1.00 1.00
117
Chnntt 1 p a n m e t e n Or 1 0 - 90 diys modehi ovtr
1 -2
3 3
t
1- 1
27 1
3-2
202
118 h n h n h n h
0.44 0.34 0.31
165 - 230 da13 90 - 165 &y
n h
0.34 0.32 0.30 0.31 0.32
1-1
49
0.59 0.37 O 0.29 0.32
4-2
20 1
0.56 0.33 0.31 0.30 030
0.55 or u calculatcd
0.44 0.34 0.33
0.50 0 1.00
230 - 271 Aap I
1.0 1.0 1.0 1.0 1.0
0.34 0.31 0.30 0.29 0.30
2-2
204
1-83 0.83 0.8 1.85 0.85
0.58 037 0 3 2 0.31 0.32
n h n h n h n r h n h n h n h
1.5 1.2 1.3 1.0 1.0
3- 1
128
0.63 or as cdculatcd
0.50 1.20 1.00
2- 1
270
1.1 1.0 1.0 1.2 1.2
0.51 0.36 0.30 O31 0.32
1.75 0.9 1-05 1.5 1.2
0.35 0.30 0.29 031 0.32
0.68 0.39 0.32 0.31 033
1.2 0.9 0.9 2.0 1.0
1.1 1 1 1.13 1.0 1.15
1.7 1.35 1.0 0.9 1.0
Acetate in the solution [mg&]
Propionate in the solution [m@]
O 500 Io00
Butyrate in the solution [mg/L]
Fig. 4.1 SORPTION of VFAs onto the HALTON TILL: linear isothems with 95 % confidence interval of K, :
(a) acetate @) propionate (c) butyrate
Fig.4.2 DIFFUSION of ACETATE THROUGH HALTON TILL: a) source and receptor solutions b) depth profile
Source n
---............. u
Fig. 4.3 DIFRISION of PROPIONATE THROUGH HALTON TILL: a) source and receptor solutions b) depth profile
Fig.4.4 DIFF'USION of BUTYRATE THROUGH HALTON TILL: a) source and receptor solutions b) depth profle
Fig. 4.5 Ceil assemblies used for testing of intrinsic degradation: (upper) glas ce& with 3 and 5 cm plugs connected to the feed network (Note msty oxidcidized goil color and clear solutions at the beginning of the test.); (lower) close-up showing 8-ce1 assembly with 3 cm soil plugs
Fig. 4.6 lntrinsic degradation of organic chernicals fiom synthetic KVL leachate in compacted Halton till: (lefi) CC1 11-1 after 162 days; (right) Ce11 IV-2 after 163 days; (Note distinct blackening at the upper and lower interface with source and receptor solutions, indicative of action by sulfate reducing bacteria. Plug color remains msty in the middle. Evident in the photographs are fermentation induced cracks in the soi1 core with entrapped gas bubbles. Compare with Fig. 4.5. Also, note the formation of reduced and expanded layer at the top interface with source solution.)
Time [&YS]
(b) Porosity [n = w G,/ (1+ w GJ], calculated for each layer at designated times :
Fig.4.7.1 VARIATION of POROSITY in 3 cm THICK S O L PLUGS (Halton Till, cornpacted); (a) time profiles @) depth profiles
3i : SRB [dLImkcfL/gl
SRB [&KG mg] SRB [CtirimL;
Fig. 4.8.1 DISTRIBUTION OF HAB and SRI3 in HALTON TILL : 3 cm compacted soil plugs: (a)@) t h e profiles (c)(d) depth profiles (values for soii [&dg] are expressed with respect to the dry mass of soil)
SOüRCE RECEPTOR U Y E R 1 U Y E R 2 WYER 3
30 60 90 120
The [days]
I O " 100 10' 10' 10' IO" 100 10' 102 10'
CeIl 1-1 9 23 &y3
O CeIl 1-2 4jI 29 days
O 2 4 6 0 - .
(d) j f, [?/a of dry mil)
Fig. 4.9.1 DISTRIBUTION OF ATP and fJn HALTON TILL : 3 cm compacted soi1 plugs: (a)@) time profiles (c)(d) depth profiles (values for soil [ng/g] are expressed with respect to the dry mas of soil)
Source
Time [days]
-10 2 (b) Acetate [g5], in pore water for D = 3.5 x 10 m /s
a Cell IV-l @ 1 17 day
A CcIlII-2@llH&y O Ccll I-l @! 23 days
a Cell 1-2 fa 29 days
. - . . . without degradation 1-1 -- without &gradaiion 1-2
Fig. 4.10.1 INTRWSIC DEGRADATiON OF ACETATE in HALTON TILL: 3 cm compacted plugs; (a) source & receptor solution (b) depth profiles
Time (days]
-10 2 (b) Propionate [g/L], in pore water for D = 2.5 x 10 m Is
Fig. 4.1 1.1 INTRINSIC DEGRADATION OF PROPIONATE in HALTON TILL: 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
(a) - D = 1.5 x 1 ~ ' ~ r n ~ / s n variable. degradaticri in Layn 1 only
v celis m-1 & ln-2 --- - - D = 1.5 x 10-'~rn~/s n variable ~ c r i i r l L a y c r I a u l d
O CelL N-1 & N-2 Laya 3, IV-2
D = 1.5 x 10''~rn'/s n variable degradarion in Laycr 1 and mxpm solution. IV-2
O 30 60 90 120 150
Tirne [day s]
-10 2 (b) Butyrate [@], in pore water for D = 1.5 x 10 m 1s
Fig. 4.12.1 INTRINSIC DEGRADATION OF BUTYRATE in HALTON TILL: 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
.... D = (23- 3 5 ) x 10"~rn'ls
-11s 11-1 & IV-2 n = % W i t h o U t ~ o I l
- D = (2.3 -3.5) x 10"~rn~ls ocIt n-1 & IV-2 n = variable with degradation in Layer 1
O 30 60 90 120
Time [days]
10 2 (b) VFAs calculated as DOC [a], in pore water for D = ( 2.5 - 3.5 ) x 10' m Is
VOCs-DOC &gadation a m s i c i d in tht hyer 1 only and modtllcd as: t,, = 0.S &ymtbc Lnycr l(0.65 tq~ cm) afta 140 &y-la& pomsity: initia@ 0.34 d as calculalcd at daigrilttd tinrs (ni Table 43): n = 034(Oto 5S)d .r ; n=0.44 (55 to 140) biys; n = O S @ 140 days (note: ha wïlh Ibc srmc gmbol rrprrsait the range of diffusion d e i - vhcrr 2.5 x 10.'~m~ls givs lmro WC in soluüoo or porc w i r p t h . 3.5 x 10*'~m~ls)
Fig. 4.13.1 INTRINSIC DEGRADAnON OF VFAs (as DOC) in HALTON TEL: 3 cm compacted plugs; (a) source & receptor solution @) depth profiles
Time [days]
(b) Porosity [n = w GJI + w G,)], calculated for each layer at designated times :
..... Cell 1-2 @ 33 days
- CeII 1-1 @49 d v
[y (2) 7
- cea 3-1 @ 128 drys
Fig. 4.7.2 VARIATION of POROSITY in 5 cm THICK SOIL PLUGS (Halton Till, compacted); (a) tirne profles (b) depth profiles
8 Ccll 1-2 33 da. V Cc11 3-1 128 &y3 Ce11 4-2 @ 20 1 days A CcIl2-1 @ 270 da43 O Ccll 1-1 49 d a y V Ce113-2@202day O Ccll4-1 f@ 271 &y
A Ce11 2-2 @ 204 days
O b
8 8 O 0
O t 10' - O
O O O
O 0 I 1 a 1 r I -
Fig. 4.8.2 DISTRIBUTION OF HAB and SRI3 in HALTON TILL : 5 cm soil compacted plugs: (a)@) time profiles (c)(d) depth profles (values for soil [cfù/g] are expressed with respect to the dry mass of soil)
O SOURCE 0 RECEPTOR
1 - A U E R 1 O 30 60 90 120 150 180 210 240 270 O UYERZ
V WYER3 O MYER4 O MYER5
IO9 10'
(b)
10' 1 9 g B Q
O
B 10' r u O
O 9 8
1049 O 10' 1
O 1oZ6 8
1 0 ' ; . - 1 - I I 1 I 1 1 1 I J
SOURCE RECEmOR LAYER l LAYER 2 LAYER3
IO-' IO-' 10"01 IO2 10' 10" 10'~ loO 1 0 ' 102 ld 1 0 ' ~ 10-l 100 10' 102 10' 10-1 10-1 100 101 102 10J
Fig. 4.9.2 DISTRIBUTION OF ATP and f, on HALTON TILL : 5 cm compacted soil plugs: (a)@) time profiles (c)(d) depth profiles (values for soil [ngg] are expressed with respect to the dry mass of soil)
10 2 (b) Acetate [mu], in pore water for D = 3.5 x 10- m 1s
O Cell 1-2 9 33 dap O CeII L-l(ijl49&ys
Fig. 4.10.2 INTRINSIC DEGRADATION OF ACETATE in KALTON TILL: 5 cm compacted plugs; (a) source & receptor solution (b) depth profiles
I O Celb 1-1 lé 1-2 / A CcL2-1 LZ-2
O 30 60 90 120 150 180 210 240 270
Time [days]
10 2 (b) Propionate [a], in pore water for D = 2.5 x 10- m 1s
Fig. 4.1 1.2 INTRINSIC DEGRADATION OF PROPIONATE in HALTON TILL: 5 cm compacted plugs; (a) source & receptor solution @) depth profiles
O 30 60 90 120 150 180 210 240 270
Time [daysl
-10 2 (b) Butyrate [fi], in pore water for D = 1.5 x 10 m 1s
= 1.5 x IO-" m'h n wiable degradation in Layci 1 only
Fig. 4.12.2 LNTRINSIC DEGRADATION OF BUTYRATE in HALTON TILL: 5 cm compacted plugs; (a) source & receptor solution (b) depth profiles
..... D = ( 2.5 -3.5 ) x 10-'~m'/s O Celb 1-1 & 1-2 cells2-1& 4-1 n = h A Ceils 2-1 & 2-2
1 - V Cdls 3-1 & 3-2 withouldtgradation
O Cclls 4- t & 4-2 - D = (2.3 - 3.5) x 10'"'mZ/s =Ils 2- 1 & 4-1 n variable w i t h ~ a n i n L j y e r 1 dcgrrdation lag 180 da)%
O
Time [daysl
-10 2 (b) VFAs calculated as DOC [g/L], in pore water for D = ( 2.5 - 3.5 ) x 10 m Is
I Cell4-2 @ 20 1 days Ceil 3-2 @ 202 dan rvaagd for 202 days
pomnty: mitiaiiy 064 and u calculalcd at Qipfcd h m (m Table 4.3): n = O 3 4 (O to 165) diys, n = 0.55 (165 - 230) diys, n =0.63 @ î30 days
A C~ll2-1 @ 270 dap Ccll S-1 1ZJ 271 day avcraged for 270.5 da'
Fig. 4.13.2 INTRINSIC DEGRADATION OF VFAs as DOC in HALTON TILL: 5 cm compacted plugs; (a) source & receptor solution (b) depth profiles
Fig . Intrinsic degradation of organic chernicals from synthetic KM. leachate in compacted Halton till: Ceil 4-2 (with 5 cm thick plug) &er 201 days (Note the formation of the fluidiad soi. layer at the top interface with the source solution and gas bubbles entrapped. This layer is visibly reduced and has dark gray color as opposed to the rusty and stU compacted soil in the middle of the plug. Blackenhg indicative of thriving of sulfate reduMg bactena is also developed at the bottom interface with the receptor solution. Evident in the photograph are lateral cracks in the soil plug formed due to generation of H2, C a , H2S and C a upon VFAs fermentation. Source solution has turbid "cloudy" appearance, wMe receptor remains ciear, but murQ grayish. Both solutions have pungent foui odor. Compare with Fig. 4.5.)
CHAPTER 5 INTRINSIC DEGRADATION OF VOLATILE ORGANIC
COMPOUNDS TaROUGH COMPACTED CLAYEY SOL'
5.1 Introduction
The laboratory study presented in this chapter is a continuation of a large body of
work on difision dominated mass transport of various groups of chemicals recognized as
pollutants in municipal solid waste landfill leachates (e-g. see Rowe et ai., 1995). The
following work deals with intrinsic degradation of selected volatile organic compounds
(VOCs) as they difise through laboratory compaded clay bamier made of Halton till.
One of the objectives was, as elaborated previously in the Chapter 4, to perform both
batch sorption and difision tests with this soil and a selected class of organic chemicals
in order to obtain an independent estimate of sorption and diffision rates. The major
objective was to perfonn a long terni difision test with sarne soil and with continuous
supply of the synthetic growth medium resembling municipal solid waste landfill
(MSWL) leachate contaminated with VOCs. With such a test it is intended to examine the
potential for intrinsic degradation of these chemicals under "no flow" and adverse
diffision dominated, mass transport conditions. Thus the indigenous soil microbial
population is lefi un-arnended to initiate the breakdown of VOCs without any
technologicd intervention that would Eicilitate removal of these chemicals. The final
objective of this expenmental study was to estimate first order decay rates, (Le. estimates
of the half lives for selected VOCs) that could be used for assessing of potential
contaminant impact in the environment.
1 This manuscript is in preparation for pubiishing
5.2 Sorption of volatile organic compounds (VOC) on clayey soi1
Sorption, generally denoted in this study as an uptake of dissolved pollutant(s) by
soil or by any of its constituents, is regarded as one of the dominant mechanisms
controlling the transport and fate of nonpolar (nonionic) organic pollutants in the
subsurface environment. For water-saturated soils and aqueous soil systems, this
mechanism is characterized as partitioning (solubilization) of nonpolar organic chernicals
into nonpolar (hydrophobic) soil organic matter (Chiou, 1989; Schwarzenbach et al.,
1995). The soil organic phase fundons as an organic solvent, which extracts the
hydrophobic organic chernicals from soil aqueous solution or pore water, while the
uptake by soil mineral phase remains marginal, mostly due to strong dipole attraction
between charged surfaces of the soil and water.
In order to estimate the extent of hydrophobic sorption on the soil, and under the
assumption that this process is driven by the excess free energy of aqueous solution
(Schwarzenbach et al., 1995), environmentai chemists have established correlations
between soil organic matter (or carbon) and chemical-physical properties and molecular
descriptors of organic compounds based on well-known linear fiee energy relationships
(LFERs). Many of such empirical expressions (Chiou, 1989; Kariclchoff, 198 1 ; Lyman,
1982; Schwarzenbach et al., 1995) correlate the soil organic carbon (or matter) /water
partitioning coefficient Km (or Kom) to water solubility, C,,?, or octanoVwater partitioning
coefficient, K,, of a chernical, as does the fkequently cited equation of Karickhoff et al.
(1 979), used herein as well:
log K, = log Km - 0.2 1
Although, comrnonly used, this equation (and equations based on LFERs), may
give inaccurate estimates of sorption if extrapolated to incompatible solute-soi1 system
(Brown & Flagg, 1981), particularly if the soil has an organic carbon content lower than
0.1% (Schwarzenbach et al., 1995), thus their use should be focused on checking for the
"reasonableness of the values" of the tested parameters (Mackay et al., 1992).
The organic carbon.water partitionhg coefficient, K,, can be also calculated fiom
the percentage of soi1 organic carbon, f, [%], and the empiricdy determined sorption (or
distribution or partitioning) coefficient, Kd, as given by Lyman, 1982:
Independent estimate of K . is made from sorption isotherms by measuring the mass of
chemical sorbed ont0 soi1 [mg or ,u@g of dry soil] vs. chemical aqueous concentrations
[mg or pg/L] over a representative range at constant temperature under phase equilibrium.
The Freundlich (Eq. 4.1, C, = K/xCw") linear (Cs =KdxCw) and Langmuir isotherms (Eq.
4.2, Cs =S,,xbxCJ[l + bC,,,]), as defined in Chapter 4, 4.2.2 are used herein to infer
sorption rates for the tested VOCs.
If the Freundlich exponent, n = 1 then Kr corresponds to Kd, and the sorption is
linearly related to concentration and independent of temperature at reaction equilibrium
with srnaIl enthalpy changes. This case is consistent with hydrophobic partitioning
concept (Hassett & Banwart, 1989). In addition, it is implied that sorption is instant with
fast equilibnum (Kd is not dependent on time) and reversible, exhibiting superimposable
desorption. However, there is growing number of recent studies, that reponed isotherm
nonlinearity (Huang et al., 1997; Xing & Pignatello, 1997) as well as distinctly slow
(tirne-dependent) and hysteretic sorption (Pignatello & Xing 1996). Under these
conditions, more complex concepts and equations may be required for defining sorption.
In this study, baich equilibrium sorption tests are performed with objective of
identifjmg whether sorption of a selected group of nonpolar organic pollutants follow the
simple linear sorption model. Given the chernicals high vapor pressures and somewhat
variable polarities, (variable and relatively high solubilities, for a class of chernicals
considered nonpolar, see Table 1. 1), it was of interest to see whether this variation was
reflected in the sorption characteristics. It is expected that sorption on the soi1 with low
organic carbon content such as Haiton till will not be the dominant removal mechanism
from polluted soil solution, however, contribution of sorption to the overall attenuation
might not be negligible.
5.2.1 Materiais and methods
Working solutions of environmentally representative concentrations -(60 - 6000
pgZ) were prepaï-ed by injecting designated arnounts of VOCs-methanol stock into
capped 60 mL glass serum bottles fiiied with de-ionized distilled water. (The VOC-
methanol stock contained ail of the VOC dissolved at known concentrations.) These
solutions were left for 3 hours to equilibrate before the measurements of the final
concentrations were taken. 35 mL heavy duty glas centrifuge tubes (Krnble Glass Inc.)
were previously treated with Sylon-CT (Supelco-Sigma, USA) to deactivate glass surface
and rninimize loss of sorbates ont0 the vesse1 walls. Haiton till (from the Halton Waste
Management Site, Southem Ontario, see Table 2.2) was air dned, pulverized and passed
through U. S. sieve No.4 (4.75 mm). Mass of - 10.25 g of this soil at 2.4% moisture
content (yields -10.0 g of oven-dried soil) was placed into centrifuge tubes. The set of
three tubes was then completely filled to the top with working solutions having a
panicular VOCs concentration. The tubes were closed with hole-caps lined with 0.005"
PTFE/O. 12" silicone (Kimble Glass Inc.) septa. The exact amounts of soil and solution
added to each tube were determined gravimetncally using an analytical balance.
The amounts of solution and soil used resulted in a water to solids ratio of - 3.6
which is within the reasonable lirnits as recornrnended in the standard test method (ASTM
E 1195-87, 1988) for the chernicals expected to sorb slightly ont0 the soil.
The duplicate control "blank" tubes containhg no soil were handled as described
above, except that no soil was added into the centrifuge tubes. To infer the losses
introduced by the preparation procedure, the concentration measurements were taken
immediately after filling the "blank" tubes. The samples and "blanks" were placed on a
wrist-action shaker (Burrell Corp. Pittsburgh, PA) for 48 hours at lab temperature (24 t 2
O C ) and t hen centrifuged at 2000 rpm for 20 minutes. Mer centrifugation, 0.1 - 1 mL
aliquots were taken from the supernatant using a gas tight syringe to determine aqueous
concentrations. The equilibriurn concentrations were those determined experimentally,
while the initial concentrations were measured fiom the blanks and corrected for the
losses observed der the shaking. The amounts sorbed ont0 the soil were determined from
the difference between the initial and final concentrations, foiiowing the calculation
procedure specified in ASTM standard E 1195-87.
5.2.2 halytical measurements
Analyses and subsequent quantification of the tested volatile organic compounds
were done by gas chromatography. For the description of the procedure and the analytical
conditions, please see section 5.3.2 below.
5 .Z.3 Results and discussion: Batch sorption tests
The results of batch sorption tests for the selected VOCs are presented in Figs.
5.1.1&2 -5.2.1&2. Fig. 5.1.1 shows the linear sorption isotherms, while Fig. 5.1.2 shows
the Freundlich and Langmuir nodinear isotherms for the tested chlorinated aliphatic
volatile compounds:(a) dichloromethane @CM); (5) 1,2-dichloroethane (I,2-DCA) and
(c) trichloroethylene (TCE). Figs. 5.2.1 and 5.2.2 show the linear and the nonlinear
isothems respectively, for the monoaromatic volatile compounds (a) benzene; fi) toluene; (c) ethyl-benzene and (d) xylenes. Data were processed with comrnercially
available program GiaphPad Prismm V.2.0. The best estimates for sorption coefficients
and the results of the regression analyses are summarized in Table 5.1.
It is emphasized that the tested nonpolar chemicals are also very volatile, with high vapor
pressures and high Henry's Iaw constants as well as with relatively low hydrophobicity
that is, low octanouwater partitioning coefficients K, as can be appreciated fiom Table
1.1. Thus, it is likely that some of the chemicals £?om the solution could have been lost
due to leaks and volatilkation. Any such losses would result in an over-estimation of
sorption.
As evident from Figs. 5.1.1 and 5.2.1, ail of isothens are l i e u for the low
concentrations of the tested chemicais. As the concentrations in solution increased (stili in
low mgL), the amounts sorbed ont0 the soil became visibly nonlinear, forming a
hyperbolic-like (saturation) pattern which consequently, was fitted better by the
Freundlich and Langmuir sorption equations, as s h o w in the Figs. 5.1.2 and 5.2.2. This
observation is not unexpected in view of growing body of literature descnbing nonlinear
sorption. Thus, although it is oflen assumed (and reported) that sorption of many
contaminants is linear over a broad range of dissolved concentrations within the aqueous
solubility limit (see Curtis et al., 1986; Chiou, 1989; Schwarzenbach et al., 1995), the
panitioning dnven sorption, of many nonpolar organic chernicals ont0 natural soil (andjor
organic matter in the soil) becomes nonlinear at high solution concentrations (Ailen-King
et al., 1996; Broholm et al., 1999a; Ball & Roberts, 1991; Xia & Ball, 1999). This is
attributed to formation of a separate organic phase at hi& (close to solubility)
concentrations (Hassett & Banwart, 1989), chernical and stmcturd properties of soil
natural organic matter (Huang et al., 1997; Xing & Pignatello, 1997) or saturation of
sorption sites and loss of the afinity for a particular chernical noticeable even at low
solution concentrations (Schwarzenbach et al., 1995; when the Freundlich Eq. exponent
n<l, as observed herein). Nonlinear isotherms will likely generate better fit-lines over a
broad range of tested concentrations in a naturd soil with inherent heterogeneities, but
their use is not required at low concentrations, particulariy if there is no obvious deviation
From linearity (Ball & Roberts, 199 1).
The linear sorption coefficients, Kd in the Table 5.1, observed at selected tested
concentrations are generally low as expected for volatile nonpolar (nonionic) organic
chernicais. They compare well with the published coeEcients obtained with the similar
type of soil (Chiou et ai., 1983; Maraqa et al., 1998; Lee et al., 1989; Pavlostathis &
Mathavan, 1992; Walton et al., 1992; see Table A5.1, Appendix 5) , given the uncertainty
arising from measurernents and undefined hydrophobicity of the low soii organic matter.
With the exception of dichIoromethane, the observed K d values are also in good
agreement with the values of Kd for "intemediate" and "high" organic carbon contents
0.29% < f, < 0.45% of the Halton till predicted (Le. cdculated) from equations of Lyman
(1 990) and Kariclchoff (Wg), as shown in the Table 5.1. For f,= 0.45%, the observed Kd
values are within (k) 50% (factors 0.6-1.5) or closer to the predicted ones, while the
intermediate f,= 0.29%, yields very good match to the observed values for ethyl-bemene
and xylenes (-20%) but over-predicts slightly @y a factor -2.2) for chiorinated aliphatics,
benzene and toluene. When a value off, = 0.14 % is used, the observed K d values are
higher than the predicted by the factors 2-5. (For example, observed Kds for
tetrachloroethene on Borden aquifer, reported by Bail & Roberts, 1991, are 7 to 33 times
higher than predicted by equations used herein; for more details concerning the
calculation, see Fig. AS. 1, Appendix 5).
The cause of the "high" Kd, for DCM is unknown, although some dipole
interaction with mineral surfaces could not be d e d out because of its polanty and high
solubility. Given the overall Iow values of the Kds, noted deviations are considered
marginal and as such they do not significantly affect the extent of sorption for the volatile
chemicals tested.
This cornparison indicates that the partitioning (dissolution) between chemicals
tested in the soi1 solution and soi1 organic matter is not the single dominant mechanism of
sorption of the chemicals tested ont0 Halton till, as postulated by equations based on
linear kee energy relationships (Chiou 1989; Schwarzenbach et al., 1995;). i t has been
suggested by Hassett & Banwart (1989) that the clay rich mineral surfaces affect the
hydrophobic sorption, particularly for the soils with organic matter content below 6-8 %
(for Halton till 0 . 2 3 < 1 , [%]< 0.77), resulting in Kd values higher than predicted from
water solubility or octanoVwater partitioning coefficient. This could be related to the
presence of intact and hydrophobic -Si-O-Si- bonds (Chen, 1979; Xu et al., 1997), which
contribute to sorption of nonpolar (hydrophobic) organic chemicals. However, this type
of attraction diminishes with increased degree of isomorphic substitution and disniption
of the bonds. More negatively charged -SiOH sites are formed which hold strongly
hydrated cations and water, but do not attraa nonpolar compounds (Hassett & Banwart,
1989). Some attempts have been made to account for a contribution of the mineral surface
to the hydrophobic sorption (McCany et al., 198 l), however, many experts (Lee et al.,
1989; Xu et al., 1997) consider that the natural clayey soils and aquifer materials with low
organic carbon (or matter) content are not effective as sorbents of nonpolar (nonionic)
organic pollutants. Another possible source for slightly higher observed than predicted Kd
coefficients could be the losses, which despite the corrections, might have been attributed
to sorption.
Nevertheless, the values of the linear sorption coefficients observed in the present
study are within reasonable limits for the chernicals tested, given their physical properties
and type of sorbent soil. Considering the test methodology and type of soil, it is unlikely
that the sorptioq if taken to be rapid (and non-kinetic), could be higher than that inferred
herein.
5.3 Diffusion of volatile organic compounds (VOC s) tbrough compacted clayey soi1
Diffusion, which is mass transport driven by chernical potential (i-e. a chernical
gradient), has been recognized as a dominant mechanism for contaminant transport in low
permeability soils and engineered waste facilities where a chernical gradient is established
upon disposal of waste. Although its significance has been well recognized and numerous
manuals, handbooks and reviews oEer exhaustive List of procedures for testing diffision
in soil (Page 1980; Rowe, 1987; Rowe et al., 1988; 1995; Shackelford, 1% 1; Shackelford
& Daniel, 1990a & b) there are still surpnsingly few studies of difision of volatile
organic contaminants in saturated compacted soil, with reported diffision rates
determined from experimentation or field case (Table A5.2, Appendix 5) .
The purpose of this part of the presented study is to delineate the processes of bulk
diffision in the pore space and instant (non-kinetic) sorption that are presumed to govem
transport of a group of tow-polarity organic pollutants such as VOCs. In this section, the
experiments and consequent data analysis based on Eq. 1.1 are presented with the
objective of proposing estimates of difision coefficients for the eight seleaed VOCs
derived from short term laboratory tests with compacted Halton till.
5.3.1 Materials and method
Diffision tests were conducted in a single-piece glas diffusion cells, as descnbed
in Chapter 4, section 4.3.2. Run with a finite amount of VOCs in the source solution
reservoir (see Fig. 3.1) at lab temperature (24 f 2 OC), these tests employed two dserent
types of source solutions, as follows:
I aqueous solution of VOCs only (-5 to 3 mgL) placed over 3 cm thick
precompacted Haiton tiil plug. It was intended to gain insight of difision,
independent of the iduence of volatile fatty acids V A S ) and to run the test
as long as necessary to coiied a representative break-through data in the
receptor solution.
2 aqueous solution of VOCs containing one of the VFAs placed over 2 cm thick
precompacted Halton till plug. With this source make-up, the influence of
volatile fatty acids could be assessed, in terms of cornpetition between the
solutes denved From the differences in polanties and concentrations as well as
beginning of degradation due to VFAs reactivity. Expecting the outcome of
rhis test to be afFected by VFAs relatively fast, the short soi1 plug was used in
order to obtain a representative profile in the receptor solution.
The concentrations of VOCs in both source and receptor solutions were monitored with
time. Upon test termination, the soi1 was also analyzed in order to assess the distribution
with depth of VOCs in the pore water. Supplementary control expenments without
Halton till were carried out at the sarne time in the glas cells having the same size and
shape as source solution and in the semm bonles, with intention to assess and account for
the losses fiom the solution.
5.3.2 Analytical measurements
0.1 - 1 ml, of aqueous sarnples was withdrawn from the testing cells using gas tight
or polypropylene disposable syringes and transferred imrnediately into a giass via1 filled
with 4 - 3 mL of 1% methanoVwater stabilizing solution. It was necessary to dilute the
samples in order to prevent the chemical overload ont0 the analytical instruments.
Standards were prepared by diluting - 1 mg/L (each) methanol stock of the tested
chernicals in the 1% methanoVwater solution. Every 4 mL vial (either standards or
unknowns) was spiked with a designated arnount of chlorofon (yielded - 20 pg/L),
which was used as intemal standard in analytical quantification procedure. M e r 3-4
hours of equilibration, 0.8 mL was withdrawn in duplicate fom each 4 mL prep viai and
dispensed into 2 mL autosarnpler vial. Ail of the screw top glass vials were equipped with
hole-caps with PTFE lined silicone septa in order to minimize the volatilization losses.
Analysis of soil sarnples included an extraction of -20 g of soil in 20 g of 100%
methanol. Upon temination of the test, the soi1 plug was sectioned into slices (layers) and
pieces of the each slice were quickly immersed into 40 mL screw top glass vids
containing methanol. Vials were weighed, capped, shaken vigorously and sonicated for
10 min, then left for 24 hours to rest. 0.1-0.8 mL of the methanol extract was withdrawn
from each of the vials and treated subsequently as an aqueous sample. Preparation testing
showed that repeated extractions were not efficient in recovery of additional amounts of
volatile chemicals, thus a single rnethanol extraction was considered suficient for the
analysis. It is not known as to the analytical precision and bias introduced with this
procedure, since neither aged contarninated soil standard nor alternative technique are
available.
Volatile organic chernicals were analped using a Varian 3800 gas chromatogram,
equipped with programmable injecter, Varian Satum 2000 ion trap mass spectrometer
detector and Varian 8200 Autosarnpler. The separations were done on a 30 rn x 0.25 mm
ID Fused silica capillary colurnn coated with a 0.25 mp DB-5 film and a He carrier gas
velocity of 1.3 ml;inzn. Column was programmed to hold for 0.5 min at initiai
temperature of 35"C, foliowed by first ramp of 10°C/min to 100°C, second ramp of
5O0C/min to 2OOoC, and a finai hold of 3 min. Injector and transfer line were held at 200
and 170°C respectively. A mass range of 35-200 m u (mass units) was scanned. It was
not possible to separate m and p xylene isomers on the analytical colurnn DB-5 and
t herefo re the two xylenes were processed and reponed together.
Injections were performed with a Varian 8200 autosarnpler equipped with a 100
,un poly-dimethylsiloxane solid phase micro extraction (SPME) fibre (Supelco, Bellfonte,
PA). The syringe needle with sheathed fiber pierced a 2 mL autosample vial septum
exposing the fiber to headspace during 10 min absorption. M e r this time, which allowed
for the equilibration transfer of the analytes to the fiber, the syringe (with fiber in
retracted position) was moved from the via1 to the injection port where the fiber was
desorbed for 2 min. Peak integration, calibration and quantification were done using
supplementary Varian Saturn 2000 Star Chromatography software with NIST spectra
library. Routine control checks for the instrument performance confirrned very good
reproducibility and high precision and recovery of the standards and unknowns, making
detection of 50 - 100 pgLL feasible and reliable.
5.3.3 Results and discussion: Difision tests
The results of the difision tests for selected VOCs through compacted Halton till
are shown in the Figs. 5.3 -5.10. Results of the control tests performed in the cells
without soi1 plugs are shown in the Figs. 5.1 1 - 5.13. The best estimates of the diffision
coefficients for the VOCs were obtained by solving difision equation for the clay pore
space (Eq. 1.1) using computer program POLLUTE v.6.5 (Rowe & Booker, 1999).
Summary of diffision coefficients uiferred from these simulations supplernented with
other parameters used for modehg of difision is given in Table 5.2.
For the both testing solutions, data points collected in tirne indicated a decrease of
concentration in source solution and diffisive breakthrough into the receptor solution for
al1 of the chlorinated aliphatic solvents: DCM (Fig. 5.3), 1,2-DCA (Fig. 5.4) and TCE
(Fig. 5.5) as well as benzene (Fig. 5 -6) and toluene (Fig. 5.7). The breakthrough of ethyl-
bemene (Fig. 5.8) and xylenes (Figs.
respike the source solution of testing
5.9 & 5.10) was uncertain, so it was decided to
ceU 1 with extra arnount of VOCs. Celis 2 & 3
containing VOCs with acetic and propanoic acid respectively, were terminated d e r 7 1
days, while ce11 1 was lefl running for 206 days.
Monitoring of the stability of the seleaed VOCs in the control tests without soil
indicated variable degree of losses f?om the solutions in the glas cells (Test 1) and glas
senim bonles (Tests 2 & 3). Barone et al., 1992 and Myrand et ai., 1992 reported
occurrence of losses in their studies on diffusion with VOCs, as well.
It is noted that the surface area to volume ratios (AIV) for the control Test 1 and
Tests 2&3 were 0.77 and 1.1 1 respectively. Furtherrnore, the pH of the solutions with
fatty acids (Le. with acetate for Test 2, and with propionate for Test 3) was adjusted to 6,
thus increasing their ionic strength to -1 = 0.08 M. This presence of dissolved salts,
(herein, Na-acetate or Na-propionate) reduces the activity (effective concentration) of
nonpolar (nonionic) organic chemicais in solution (i.e. VOCs), and facilitates their
sorption or "salting out", particularly at 1 B0.1 M. (Pavlostathis & Jaglal, 1991;
Schwartzenbach et ai., 1995). It is considered likely, that the high surface to volume ratio
and the presence of salts in the solution associated with control tests 2&3 contnbuted to
the higher losses of VOCs in these cells relative to losses in the control test 1 [see the Figs
5.1 1 (chlorinated aliphatics), 5.12 (benzene and toluene) and 5.13 (ethyl-beruene and
xylenes)]. The nature of the losses is not known. Sorption ont0 the glas surface is not
d e d out (Barone et al., 1992), yet it cannot be confirmed since desorption tests were not
perfonned. The nature of the data, however, suggests kinetic (time dependent, not
equilibrium) process, possibly abiotic reaction, leaks through septa or sorption. The extent
of (increasing) loss for a pahcular volatile chernical (see data for Test 1, Figs 5.1 1, 5.12
and 5-13)? is consistent with (increasing) values of Henry's law constants (H [kPa
rn3 inoi], ratio of partial or vapor pressures and aqueous concentration or solubility, al1
values listed in Table 1.1, Chapter 1): negligible for DCM (0.1 1) and I ,%DCA (0.26), - 15-20s for benzene (0.55) and toluene (0.67) and - 30% for TCE (1.18), ethyl-benzene
(0.8) and xylenes (0.7).
Bearing this in mind it was likely that similar losses would occur in the diffusion
tests with soil, possibly exhibithg even more tirne dependent "saiting-out" of VOCs from
the source solutions due to (slow and time dependent) diffusion of ions fiom the soil.
Modeling of the difision data confirmed this expectation. Having source and receptor
solution profiles for concentration in tirne, coupled with depth profle for concentration in
pore water at the time of test termination for each of the tested chemicals, it was neither
possible nor justifiable to assign ail mass missing fiom the solutions to sorption ont0 the
soil. A very good fit to the data was obtained by modeling the losses in the source
solutions as first order reaction while keeping sorption ont0 the soil within the limits
inferred from independent batch tests. (For details regarding the modeling concept shown
for dichloromethane and toluene see Figs. A5.2.1 and A5.2.2, Appendix 5.)
Sorption was modeled as an equilibrium linear process, using Kd coefficients
shown in Tables 5.1 & 5.2. It was assumed that, at the time of test termination, the total
amount of each chernical present in the system (soil and pore water) Cm,, was recovered
in the analytical procedure with methanol extraction. Subsequently assuMng that,
sorption is linear, this total recovered arnount of each chernical would be distnbuted
between the soil and pore water, as C,, = & /R = Ctai /[l+(p&)ln], therefore
allowing for back-figuring pore water concentrations, using information on soil dry
density, porosity and & h m batch sorption tests. The depth profiles for each of the
tested chemicals are plotted on Figs 5.3b-5.10b (A sample cdculation for toluene
concentrations in pore water is given in Table A5.2, Appendk 5 ) .
In order to produce a satisfactory fit line to the depth concentration profiles it was
necessary to match both source solution and receptor solution concentration vs. tirne
profiles with a unique set of diffusion and sorption coefficients. Had the sorption ont0 the
soil been credited for entire decrease in source solution concentration, while keeping
difision coefficient at 25 to 30 % of its value in the free solution (which is predicted
impediment due to soi1 tortuosity, as done by Myrand et al., 1992 or Bal1 et al., 1997) it
would have been impossible to obtain equally good predictions in the two reservoir
solutions as well as in soil pore water (for more details, see sample simulation in Fig.
A5.2, Appendu 5). Interestingly, the best estimates of the diffision coefficients, obtained
through iteration and displayed in the Table 5.2 are within 25% of the values in free water
solution, which is reasonable for compacted clay, and owing to independent estimates of
the Kd, it seems that these estimates of diffision are not fomiitous.
It is recognized that the simulation with first order reaction for losses might result
in variable degree of success in fitting the data for the eight dfising volatile chemicals.
As evident fiom the displays, there is very good agreement with theoretical prediction for
DCM, both in reservoir solutions (Fig. 5.3a) and pore water (Fig. 5.3b), 1.2-DCA (Fig.
5.4a &b), TCE (Fig. 5.5a & b), benzene (Fig.5.6a & b) and toluene (Fig. 5.7a & b) for the
Tests 2 & 3. For ethyl-benzene and xylenes the proposeci coefficients result in over-
prediction in the receptor solution. For Tests 2&3 the overall impact is marginal because
of the low initial concentrations, but for the Test 1 this trend is more distinct. Somewhat
lower D = 1.5 x 1 O-'' m2/s (both 2.5 x 10'1° and 1 .S x 1 O-'' m2/s produced good fits with R'
B0.95) and even higher Kd could yield better prediction, however rnass balance
calculation indicates consistently lower mass recovery for the Test 1 than for the Tests
2&3 for al1 chemicals (see Table A5.2, Appendk 5). Total losses (not adjusted relative to
the "blanks") from diffusion test cells implied from mass baiance are within 15 -25 % for
Tests 2 &3 and 40 - 55 % for Test 1, which is opposite from what couid have been
expected from the control tests discussed earlier. Therefore, lower diffusion and higher
sorption would not be justified over simulation of extra losses in the receptor solutions.
It is also emphasized that soil plug in the Test 1 ce11 rernained oxidized (rusty
eanh color with clear and odorless solutions) during entire period of 206 days without any
measurable changes in heterotrophs (1500-9500 cfu/mL) or sulfate reducers (15000-
150000 c f i h L ) relative to background untreated compacted soil. The plugs in test cells 2
& 3, however had several tiny (1-2 mm) gray or black dots on the contact surface with
cell wall and traces of fou1 odor in the receptor fluid which indicated begiming of redox
change due to fermentation of VFAs and prompted temination of these tests &er 71
days. Both indigenous soil heterotrophs and sulfate reducers grew to 6 x 10' cftu'rnL and 6
x 10j cfWmL, respectively at expense of acetate or propionate as opposed to the
background and Test 1 counts, where the readily degradable organic carbon was not
available. Although the analytical procedures used do not allow the elimination
biodegradation as a potential source of VOCs loss, it is considered that VOCs did not
degrade to any extent. With the very good match between the observed and theoretical
profiles and unique set of modeling parameters obtained for both inert (Test 1) and
potentially bio-reactive soil (Tests 2 & 3), it is, demonstrated that the difision and
sorption could be delineated fiom each other and fkom degradation as well. Under careful
monitoring, a testing set-up as presented in this study, could simulate the response of the
cornpacted soi1 under quite realistic contamination conditions during sufficiently long
period, while practically allowing for degradation to be neglected.
Both diffusion and especially linear sorption coefficients inferred fiom the
presented study are lower than those given in the rare few literature sources (see Table
AS.2, Appendk 5). When used for prediction of contaminant impact (e-g. Eq. 1.1) the
proposed coefficients will consequently, result in higher, thus more conservative values
for the VOCs concentrations in the pore water. Given the uncertainties associated with
field scale and potential health risk fiom these chernicals, it is considered that the range of
values given in the Table 5.2 would result in a reasonable prediction of contaminant
impact in low hydraulic conductivity natural barriers.
5.4. Intrinsic degradation of VOCs under diffusive transport through compacted
clayey soii
Research conducted in recent two decades has provided remarkable break through
in elucidating the pathways and conditions of biodegradation of volatile organic
chemicals. VOCs are biodegradable (Le. subjected to rnicrobially dnven oxidation-
reduaion similar to one taking place in conventional readily degradable organic waste) if
conducive conditions for the reaction exist.
Intensive laboratory and field research confirmed that BTEX are biodegradable,
i.e. oxidized under both aerobic (Gibson & Subramanian, 1984) and anaerobic (Young,
1984; Grbic-Galic, 1990; Cozzarelli et al., 1990; 1994; Krumholz et al., 1996; Karnpbell
et al., 1996; Davis et al., 1999; Gieg et al., 1999; Armstrong et ai., 2000) conditions.
BTEX as relatively reduced organic chemicals undergo oxidation and can potentially
serve as carbon source (electron donor) for the rnicroorganisms. Grbic-Galic & Vogel,
1987, proposed the hypothetical degradation pathways for benzene and toluene mediated
by fermentative/methanogenic rnixed culture. Initial oxidation (hydroxylation) of homo-
aromatic ring was started by water derived oxygen. Benzene was consequently oxidized
to phenol, cyclohexanone, propionic (aliphatic) acid and finally to COi and C&.
Recently, Caldwell & Sufita, (2000), presented evidence for an alternative benzene
degradation pathway via phenol or, possibly even direct benzene, carboxylation to
benzoate. Toluene, as postulated by Grbic-Galic & Vogel, 1987 and Grbic-Galic 1990,
breaks down via ring hydroxylation to p-cresol and p-hydroxy-benzoic acid, as weli as,
via methyl-group hydroxyiation to beddehyde and benzoic acid. Identification of
methyl-cyclohexane and cyclohexane pointed out to the partial reduction (hydrogenation)
taking place under fermentation as well (Grbic-Galic & Vogel, 1987). Toluene
intermediates were reduced to aliphatic acids and subsequently mineralized to CO2 and
Ch. These findings are consistent with methanogenic degradation of naturally occumng
monoarornatic, benzoate (Young, 1984), where volatile fatty acids (mainly acetate, also
propionate and butyrate), CO2 and C& were generated (Ferry & Wolfe, 1976; Keith et
al., 1978; Kaiser & Hanselmann, 1982). More importantly, the generation of identified
intermediates (Grbic-Gaiic, 1990) indicated the syntrophic involvement of a very
complex mixed culture of several metaboiically different bacteria. Methanobacteritim
formicimm and Meihanospirillum hngatei alone were not able to degrade benzoate
(Ferry & Wolfe, 1976) and are considered to serve as temiinal organisms, which utilize
the degradation by-products such as acetate, Hz and CO2. Benzoate degraders were,
however, faster if the culture was pre-exposed to fermentation substrate such as acetate.
Cleavage of the arornatic ring is not thennodynamically favorable (AG'X) under
methanogenesis and thus requires interspecies action for by-product(s) scavenging in
order to be an energy yielding (exergonic) process. It was suggested that anerobes capable
of ring cieavage (e.g. Coprococcus sp. and Streptococcus bovis) or demethoxylation
(Aceiobacierium woodii) as well as many terrestriai lignin degraden could use synthetic
aromatic compounds as growth substrates (i.e. electron donors) efficiently with other
more energetically favorable electron accepton (Kaiser & Hanselmann, 1982). Later
work confinned that each of BTEX could be degraded under nitrate, iron and sulfate
reduction by soil or sediment microorganisms (Kmmholz et al., 1996; Aronson &
Howard, 1997; Kazumi et al., 1997; Nales et al., 1998) at (or fiom) the contaminated
sites, however there seems to exist distinct substrate selectivity in favor of toluene, m&p-
xylenes, and ethyl-benzene with O-xylene and benzene being the least preferred for
utilization. A recent midy of 26 isolates from the indigenous "fuel-hydrocarbons"
degraders, phylogenetically similar to common soi1 microorganisms from the
Pseudomonaî, Raistonia, Burkhoidera, Spphingomonas, FZavobacterium and Bucillus
genera, showed that more than 80% of the isolates could grow aerobically on JP-4 jet
fuel, toluene and ethyl-benzene, while only less than 10 % grew on benzene and O-xylene
(Stapleton et al., 2000).
Chlonnated aliphatic hydrocarbons, (such as the three selected in this study) are
biodegradable, particularly under anaerobic conditions. It has been observed in many lab
and field cases with contarninated soi1 and sediments, that this class of organic chemicals,
which are oxidized (unlike BTEX, which are reduced) undergo reduction by microbidly
mediated reductive dechionnation (Bouwer et al., 1981; Vogel & McCarty, 1985; Barrio-
Lage et al., 1986; van der Meer et al., 1992; Wiedemeier et al., 1998). Many researchers
suggest that the availability of hydrogen (discussed in Chapter 4) as very reactive eiectron
donor and the ability of hydrogen utilizing bactena to compete for it by regulating its
threshold levels while reducing various electron accepton might be crucial for successful
dechlorination (DiStefano et al., 1992; Ballapragada et al., 1997; Dolfing, 2000). Because
of the positive reduction potential of chlorinated compounds as oxidizing agents (i.e.
electron acceptors; Vogel, et al., 1987) and higher amount of energy that could be
released fiom dechlonnation than fiom sulfate reduction or methanogenesis, there might
emerge an energetically favorable niche for spedized hydrogen utilizing bactena
(Dolfing, 2000),"halorespirators", which would be engaged even in metabolic (growth
related, fast and complete) and not just in CO-metabolic (ofien, incomplete)
dechlorinations (Hollipr & Schraa, 1994). It is hypothesized that the optimai conditions
for the halorespirators would be at relatively low H2 concentrations (and consequently
relatively high redox potentiais, higher than required for sulfate reduction and
methanogenesis) with only limited arnount of halogenated compounds, such as those
occurring at the borders of the contaminant plumes (Dolfing, 2000).
The details of dichlormethane @CM) biodegradations are discussed in Chapter 2
and will not be repeated. h brief, DCM is fermented to acetic acid, CO2 and CH4 by
anaerobic moted culture serving, however, as carbon source (Le. electron donor).
Trichloroethene (TCE) and 1,2-dichloroethane (1,2-DCA) could be discharged as such,
but they are frequent as the daughter products of tertachloroethene (PCE) and 1,1,2-
trichloroethane ( I , 1,2-TCA) dechlorination, respectively (Vogel et al., 1987). L ,2-DCA
was dechlorinated (CO-metabolicaily) to ethene in pure methanogenic cultures of
Methmobucterzum thermoautoîrophicum, and Methanococci (deltue and
~hennolithotrophicus) grown on H2 - CU2 (Egli et ai., 1987; Belay & Daniels, 1987). PCA
is dechlonnated through a sequence of reduction steps, producing TCE, DCE isomers
(cis-, -tram or 1,1), carcinogenic intermediate vinyl chloride and finally ethene and
ethane (Burio-Lage et al., 1986). Recently, Magnuson et al. (1998) presented a pure
culture Dehalmoccoides ethenogenes capable of complete dechionnation of PCE to
ethene. PCE (and TCE) dechlorination is aiso frequently observed with mixed cultures
derived fiom different waste effluents, supplying consequently reducing equivalents from
different electron donon. Poor fermentation substrates such as propionic and butyric
acids, which generate low levels of H2 slowly, could at least initially and in short term
increase PCE dechlorination at the expense of methanogenesis (Smatlak et al., 1996;
Fennell et al., 1997). Results Erom the lab tests perfomed by Ballapragada et al. (1997)
indicated, however, that methanogens scavenged 95% of al1 available Hz pool for
methane production under lactate fermentation, while only 5% of Hi was used for PCE
dechlonnation. Hydrogen could also be supplied from toluene s e ~ n g as main (parent)
electron donor (Sewell & Gibson, 1991). Strongly reduced environment, such as
methanogenic, (low oxidationlreduction potential Eh < -0.25 V, high concentration of Hz
10 - 40 nglL) was demonstrated to be conducive for complete reduction of highly
chlorinated compounds (Aronson & Howard, 1997; Bradley & Chapelle, 1999), however
PCE, TCE and VC dechlorination was also observed under sulfate (DiStefano et al.,
1992;) and iron (Bradley et al., 1998) reducing conditions.
Recently Wiedemeier et al. (1999) reveaied a fascinating ability of indigenous
subsurface microorganisms to carry out intnnsic bioremediation at numerous sites across
the US contaminated with BTEX and /or chlorinated aliphatic solvents by oxidation of
organic pollutants u t i l k g the entire variety of available [i-e. both natural
(mineraYitnorganic)] and anthropogenic electron acceptors. This invaluable information is,
however, extracted from nurnerous lab and field studies with sediments fiom aquifers or
similar geological layers havhg relatively high hydraulic conductivity (> 10" d s ) . There
is still surprising lack of information about biodegradation of VOCs in saturated soils and
sediments with low hydraulic conductivity, such as landfill clay liners ald natural
attenuation layers. Based on the findings ffom the contôriated plumes, it may be
hypothesized that VOCs, particularly in combination with fermentation products and
hydrogen precursors, would serve either as carbon source or be CO-metabolized by starved
indigenous rnicrobial population, and as such, they could potentidiy start degrading even
in aquitards. The acclimation (Iag) periods in such low permeability layers might not
seem too long relative to retention time of contarninants in a landfill (designed for safe
and perpetual confinement), but the rates of degradation are expected to be lower than in
the plumes, due to low flow and mass transfer limitations.
The laboratory study of intrinsic biodegradation of VOCs in compacted clay soi1
is presented in following sections of this chapter. Information on diffiision and sorption
rates obtained in separate tests as discussed earlier, is used in conjunction with the test
data reponed in the following to estimate the first order rates of VOCs degradation and
half-lives.
5.4.1 Materials and method
A detailed description of the testing ceU assembly for intrinsic degradation and the
test methodology and materials are given in Chapter 4, (8 4.4.1). Since the objective of
this study was to simulate degradation of synthetic MSWL leachate components in a
laboratory clay liner, the same ce11 assembly, (as shown in Figs 4.5-4.6 and Figs. A4.4-
A4.5, Appendix 4) was sirnultaneously employed to examine the fate of volatile organic
compounds (VOCs) and volatile fatty aads (VFAs).
5.4.2 Analytical measurements
Analyses and subsequent quantification of the tested volatile organic compounds
was done by gas chromatography. For the description of the procedure and the analytical
conditions, see 8 5.3.2 above.
5.4.3 Results and discussion: Intrinsic degradation tests
Results of intnnsic degradation tests under dominant diffisive transport of VOCs
frorn the synthetic KVL Ieachate through cornpacted Halton till are shown in Figs. 5.14.1
- 5.2 1 . 1 and in Figs. 5.14.2 - 5.21.2 for the 3 cm and 5 cm soil plugs, respeaively. Top
graph (a). on each page shows concentration vs. t h e profiles for the source and receptor
solutions for each VOC. The bottom four graphs (b) 1 - 4, on each page show pore water
concentration vs. depth profiles at the four different times at which a set of tests was
terminated. The half-lives (tiI2 = I d A , A = first order biodegradation rates) inferred from
the simulations of difisive mass transport coupled with linear sorption and first order
(biodegradation) reaction (Eq. 1.1) using cornputer program POLLUTE v.6.5 (Rowe &
Booker, 1999) are summarized in Table 5.3.
A description of the tests and the changes registered during the course of the
expenments is given in Chapter 4 (5 4.4.3). The soii in the test ceils was subrnitted to
noticeable physical transformation fiom rusty oxidized compacted plug(s) (Fig. 4.5) at the
beginning, to cracked and expanded plug(s) with visible blackening at the interfaces with
solutions in the end of the test (Fig. 4.6). As noted earlier and illustrated in the Figs. 4.7.1
and 4.7.2 (Chapter 4), soil porosity in the top layers increased gradually fiom 0.34 to 0.55
for the 3 cm soil plugs during the testing period of 162 days, and fiom 0.34 to 0.63 for the
5 cm soil plugs during 271 days of testing. It was evident that these changes were caused
by fermentation of VFAs and funher enhanced by gas generation. (presumably, C02, Hz,
CK, H2S). Routine concentration monitoring in time confirmed breakthrough of al1 of the
VOCs from the synthetic leachate in the source solution through the soi1 into the receptor
solution.
In order to confinn the dinusion under conditions with a supply of contaminants,
two sets of duplicate cells were tenninated at an early stage of testing, when it was
hypothesized that degradation would not yet be evident. The data points for the pore-
water concentration vs. depth, recovered from a set of two 3 cm thick plugs (cells E l and
1-2) terminated at 23 and 29 days, indicate excellent agreement with the theoretical
profiles for al1 of the eight VOCs tested. An equally satisfactory theoretical fit to the
experimentd points was obtained for the 5 cm thick soi1 plugs for another 2 tests (cells 1-
2 and 1-1) which were tenninated afler 33 and 49 days. Thus, the first bottom graph on
the left designated &J(l), in the Figs.: 5.14.1 & 5.14.2 for DCM; Figs.5.15.1 & 5.15.2 for
1,2-DCA, Figs 5.16.1 & 5.16.2 for TCE; Figs. 5.17.1 & 5.17.2 for benzene; Figs.5.18.1 &
5.18.2 for toluene; Figs. 5.19.1 & 5.19.2 for ethyl-benzene; Figs.5.20.1 & 5.20.2 for
m&p-xylenes and Fig. 5.21.1 & 5.2 1.2 for O-xylene, al1 show good agreement between
the experimentai data and theoreticai curves generated using the same diffusion and
sorption rates as inferred previously (paragraph 5.3.3) and aven in Table 5.2. Thus, it
appears that there was no signifiant biodegradation of VOCs d e r 23 to 49 days of
continuous supply of VFAs. For each of the VOCs, a characteristic and distinct diffusion
profile was recovered, and difision (with sorption) was vimially unaffected by
degradation after this period of testing, as represented by dotted or long-dash Iine(s) on
these figures. These profiles of pore-water concentration vs. depth (see Table A5.3,
Appendix 5) show that l i e u sorption (&) and difision (D) can be clearly distinguished
from degradation at the early stage of monitoring.
As VFAs and other components of synthetic KM. leachate diffised through the
clay, it was evident that the indigenous bacterial population in the soil was responding to
the nutnents and a significant amount of readily degradable organic carbon. The soil
staned losing its oxidized rusty (earth) color and was tuming du11 gray, fermentation of
VFAs began and trace of fou1 odor was detected in the receptor solution. The next
termination of ce11 III-1 with a 3 cm plug fouowed 63 days f i e r the start-up. The analyses
of the soil showed a clear decrease in pore water concentrations relative to the eariier
profiles recovered at tirnes of 23 and 29 days for ali three of the chiorinated aiiphatics, as
evident in Fig. 5.14.1 (b)(2) for DCM; Fig. 5.15.1 (i3)(2) for 1,2-DCA and Fig.5.16.1 (b)(2)
for TCE. In addition to aliphatics, benzene, in the Fig. S. 17.1@)(2) as well as, toluene, in
the Fig. S. 18.l(b)(2), had significantly decreased pore-water concentrations at 63 days
than earlier at 23 and 29 days. These concentrations were lower than predicted by
diffusion and sorption alone as seen on the graphs o ( 2 ) by the position of the data points
relative to the doned line. However, ethyl-benzene, in Fig. 5.19.1@)(2) and xylenes, in
Figs. 5.20.1 o ( 2 ) and 5.2 1.1 @)(2), retained the characteristic diffusion profiles and as
such were virtually unafFected.
The same findings were reproduced with another set of 3 cm thick duplicates
(cells IV4 and TI-2) tenninated after 1 17.5 days, as evident in Figs. S. M.l(b)(3),
5.15. ) ( 3 ) S.l6.1@)(3), S. l7.l@)(3) and 5.18.1@)(3). The top interface formed at
depth of 5 - 7 mm for both plugs was black and has evidently low consistency and a
moisture content higher than the rest of the plug. As displayed earlier in the Figs. 4.8.1
and 4.9.1 (Chapter 4) bacterial count has already been increased relative to untreated soil
(Le. within 117.5 days) frorn 9 x 104 to -10' cfu/mL and fiom 7.5 x 10" to 5 x 106 cfu/mL,
for HAB (heterotrophic aerobic bacteria) and SRB (sulfate reducing bacteria)
respectively. It is considered that such evident growth of the indigenous rnicroorganisms
within compacted Halton Till plugs as well as increased ATP (adenosine tri phosphate)
content (Fig. 4.9.1) couid be directly linked to fermentation of VFAs and to the
disappearance of the VOCs obsewed in the pore water concentration profiles. Pore water
concentration data for DCM in Fig. 5.14. Io)@), 1,2-DCA in Fig. 5.15. I(0)(3), TCE in
Fig. 5.17.1 (6)(3), benzene in Fig. 5.18.1 (b)(3), and toluene in Fig. 5.14.10(3), were
forming distinct almost uniform profile with tendency to equilibrate with concentration in
the receptor solution. It seemed that &er certain lag phase of 50-60 days, the soil turned
into a bioreaaor steadily rernoving signifiant arnounts of both chlorinated aliphatics and
benzene and toluene to a residuai value. Ethyl-benzene and xylenes were still unafEeaed
after 1 17.5 days retaining their consistent difision profiles as seen in Figs. 5.19.1 (b)(3),
5.20.1 (b)(3) and 5.2 1.1 (b)(3).
The same patterns were reproduced once again in the final termination of the
remaining set of the two cells @-1 and IV-2) with 3 cm thick plug afler 162.5 days. As
seen in the Figs. 5.14.1(%)(4), 5.15.1@(4), 5.16.1@)(4), 5.17.1@)(4) and S.18.1@)(4),
DCM, 1,2-DCA, TCE, benzene and toluene have been removed steadily inside the Halton
Till plug(s) under the continuous feed from synthetic KVL leachate supplied to the source
solution. Despite some data scatter at this stage of exposure, the concentration in the
receptor solution showed the signs of decline due to degradation, particularly for
chlonnated aliphatics, as noticeable in the Fig. 5.14.l(a) for DCM, in Fig. 5.15.l(a) for
1,2-DCA and in Fig. 5.16.l(a) for TCE. Benzene and toluene removal fkom the receptor
solution is not obvious, aithough it appears that the concentrations are lower than
predicted if modeled with h a 1 (and increased) porosity and without degradation, as
show by doned lines in Figs. 5.18.1 (a) (benzene) and 5.19.1 (a) (toluene). Ethyl-benzene
and xylenes were still unaffected by degradation even after 162.5 days of testing, as seen
in the Figs. 5.19.1, 5.20.1 and 5.21.1.
The number of HAB and SRB, as well as ATP content increased significantly
relative to the background (shown in Fig. 4.8.1 and 4.9.1, Chapter 4). Furthemore, after
163 days of testing, even methanogens were active in entire soil core and both solutions.
The most aggressive methanogenesis was recorded at the interface with source solution,
i.e in the top soil layer(s) (or Iayer 1) with only 2 days of reaction delay with BIOGAS
BART IM incubation. In contrast, it took 10 days for methanogens to react in samples
from other soi1 layers and the source solution, while the methanogens from the receptor
were delayed for 30 days. The positive B A R F responses pointed to well developed
syntrophic activity of methanogenic and sulfate reducing bacteria, charactenstic at low
reduction potential (< -0.3 Y), i.e. at relatively high levels of hydrogen, generated fiom
fermentation of VFAs, at essentially unlimited supply of sulfates and carbonateiC02, al1
of which were established Ui Halton till. Hydrogen and methane were not quantified
because testing apparatus did not allow for collection of gaseous products of
fermentation. The modification of the chromatographie method with SPME did not allow
for monitoring of gas accumulation in solutions as was possible eariier with the
preliminary tests described in Chapter 3 (see Fig. A3.1, Appendix 3). The thriving of
methanogens and sulfate reducing bacteria constitutes an indirect and qualitative evidence
of gas generation. As such, metabolic activity of these bactena is also a sign that the
conditions, conducive for CO-metabolic degradation of some simple xenobiotics such as
VOCs have been established.
Considering this, combined with results of monitoring and consistent and steady
pattern of concentration decrease in time and depth displayed in the enclosed figures, it is
reasonable to infer that intnnsic degradation of chlorinated aliphatics, beruene and
toluene had taking place in the compacted Halton till. As cm be seen in the Figs. 5.14.2,
5.15.2, 5.16.2, 5.17.2 and 5.18.2 this was rearmed by the findings fkom a set of
additional 6 tests using 5 cm thick soil plugs, that were run and terminated in paraiiel to
the tests descnbed above for the 3 cm soil plugs. Al1 observations noted for 3 cm soil
plugs regarding the gradua1 change of porosity of the top layer, formation of cracks due to
gas generation and growth of bacteria, soil blackening associated with lowenng of redox
potential, stand for the 5 cm soil plugs as well (already shown in Figs. 4.5, 4.7.2; 4.8.2;
4.9.2 in Chapter 4 and in Figs A4.4 and A4.5, Appendix 4).
Results of dl tests perfoned indicate that DCM, 1,2-DCA, TCE, benzene and
toluene are removed from the soil due to intrinsic degradation while ethyl-benzene and
xylenes remained unaffected by degradation under the conditions examined. Generally,
these findings are in very good agreement with numerous records compiled by
Wiedemeier et al. (1 999) and Aronson & Howard (1 997) describing intnnsic degradation
in soils and sediments originally contaminated by fuel spills. Nevertheless, it seems that
aflermath of a couple of decades without any treatrnent creates the conditions somewhat
similar to those found in anaerobic waste digesters and MSW landfills, with significant
pool of short chah and volatile fatty acids generated at the spi11 contaminated sites as
presented by Cozarelli et al., 1990; 1994; Wiedemeier et al., 1999.
The conditions created in these experiments are believed to be the best suited for
inttinsic degradation of the chlorinated aliphatics, namely low redox potential developed
gradually through fermentation of VFAs and generation of hydrogen gas is known to be
necessary for reductive dehalogenation. The results presented in this study do indicate
that degradation of the chlonnated aliphatics is at advanced stage, since both data from
pore water concentrations vs. depth and data from receptor solutions have declining
patterns for each of the chlorinated solvents. This outcome was anticipated based on
earlier tests with DCM under virtually the same conditions with sirnilar (somewhat
weaker) synthetic and real KVL leachates and with intact and compacted soil, as
discussed in Chapter 3. The reproducible results with DCM, combined with new and
invaluable evidence from the soil pore-water concentration profiles displayed in this
chapter do bring more weight and reassurance about hypothesized aspects of degradation.
Analyses of the soil by tenninating ("sacrificing") a test at certain times confinned that
the Halton till, local uncontaminated soil used for the construction of cornpacted clay
liner from the real landfill site has a microbial population that is capable of intrinsic
degradation of the three chlorinated aliphatic solvents.
The reduced conditions created in the tests are also conducive for anaerobic
degradation of monoaromatics. Toluene has been recognized as the most degradable of
the BTEX and (apart from being remarkable given the adverse difision lirnited
conditions), its removal from the compacted Halton till plugs did not corne as a total
surprise.
Degradation of benzene, very comparable to degradation of toluene in the same
soi1 under adverse diffusion limited conditions, was surpnsing, and certainly is
remarkable evidence of intnnsic degradation of such recalcitrant and toxic monoaromatic
chetnical. Aithough fiequently reported as degradable under different redox conditions in
vanous lab and field studies, benzene degradation is the most uncertain because of
comparable body of literature reponing its persistence in contaminated soil (Kmrnholz et
al., 1996; Aronson & Howard, 1997; Wiedemeier et al., 1999). Generally, the information
is gathered fiom mixed cultures under insufficiently defined incubation conditions, thus
not much explanation is given as to what might have caused lack of biodegradation of
benzene in a particular case. Apart from the influence of many complicating factors
specific to a case, it is speculated that presence of each and every of the BTEX in the
medium or in contaminated site could induce inhibitory cornpetition on the microbial
culture, which would eventually disappear upon culture's acclimation or upon selective
consumption of easily degradable of the "TEX' compounds (Knirnholz et al., 1996).
Benzene, degradation has been found at numerous fuel and gasoline contarninated sites
(Aronson & Howard, 19971, but was not obvious in a few tests with MSW landfill
leachates (Knimholz et al., 1996). Trial tests with KVL leachate spiked with BTEX,
similar to those described in Chapter 2, indicated slow removal over a penod of four
rnonths. It is noted, however, that benzene was either not deteaed or at very low levels
(-20 pglL) in ocwiondy sampled reai KVL leachate, while toluene and xylenes were
persistent at somewhat higher levels of - 100 - 250 pg4. Regardless of relatively long
lag observed in these trials, it is believed that exposure of the indigenous microorganisms
fiom the real KVL leachate to acetate (Ferry & Wolfe, 1976), or to other (even poor)
fermentation substrates like propionate and butyrate, could be a key factor in benzene and
TEX degradation. By analogy, it appears that the exposure of the indigenous
microorganisms fiom initidy oligotrophic Halton till to steady and unlimited supply of
VFAs was instnirnental in the observed benzene removal in the compacted Halton till.
Upon acclimation and beginning of fermentation under controlled lab conditions, the
uptake of benzene started. It appears that high Ievels of VFAs, s e ~ n g as primary and
virtually exclusive growth substrates, did not have an inhibitory effect on the
consumption of relatively low levels of benzene and toluene. Also, benzene's (and to
lesser extent toluene's) high water solubility and low sorption could have contributed to
its bioavailability under quite limited mass transfer conditions in compacted soi1 relative
to ethyl-benzene and xylenes. Aithough ethyl-benzene and xylenes did not degrade in the
presented study, it is hypothesized that their degradation would have comrnenced had the
tests been lefi running longer than 163 or 271 days. It is not known what could have
caused the observed lack (or retardation) of degradation. It is speculated that some fmors
other than substrates (i.e. any of the "BTEX") could be involved, since these
monoaromatics are generally considered biodegradable and the test conditions and media
are designed to simulate both wbon and nutnent unlimited growth of metabolically
versatile soil rnicrobial population.
Given the importance of the degradation on the overall contaminant impact, an
attempt was made to mode1 the rates of decay for the organic chernicals tested. The
approach devised earlier with the tests presented in the Chapter 3 (paragraph 3.5.2) was
adopted with modeling of the lab intrinsic degradation of tested VOCs. As speculated in
the Chapter 3 with DCM when concentration distribution in soil was not available, now it
can be stated, based on data presented in this paragraph, that degradation, does indeed
take place in the soil core. Because of the higher initial concentrations fed into the source
solutions and somewhat thicker soil plugs than those employed in the tests descnbed in
the Chapter 3, the degradation in receptor solutions, shown in the Figs. 5.14.1&2 (a) -
5.18.1&2 (n) is not as obvious as in the pore water, shown in Figs 5.14.1&2 (b) - 5.
18.1&2 (3). This observation makes it clear that either the lag andlor the half lives for the
degradable VOCs, inciuding ch io~ated aliphatics, was long relative to the duration of
the test, since there was negligible degradation in the receptor solutions. Based on the
recovered depth profiles fiom the tests with both 3 cm and 5 cm plugs degradation on the
concentration of the VOCs could not be measured before 30 to 50 days, and hence this
period can be considered a lower bound for degradation lag [Figs. 5.14.1&2(6)(1);
5.15.1&2(3)(1); 5.16.1&2(b)(l); 5.17.1&2(b)(l) and 5.18.1&2@)(1)]. Findly, the graduai
changes of porosity and appearance of top layer at the interface with source solution are
considered to be direct consequences of degradation. This interface appeared to be the
locus of the most intensive microbial activity and degradation. It was hypothesized that
degradation was already in progress with the fist registered (measurable) increase in
porosity because this increase in porosity could only be linked to generation of gases due
to fermentation of VFAs.
Modeling of intrinsic degradation was done with diffusion and sorption rates
summarized in Table 5.2, and with a challenging task of fitting the concentration vs.
depth profiles descnbed earlier. The best fit to the data was obtained by sirnulating
degradation only in the more porous top layer, which had a thickness that increased
gradually during the course of the expenments, as elaborated in the legend on enclosed
figures. A surnrnary of the varied parameters together with the estimates of the half lives
for the DCM, 1,2-DCA, TCE, benzene and toluene are given in Table 5.3. As c m be
appreciated from the Table 5.3 as well as from the Figs 5.14.1-5.18.1, the first order rates
of intrinsic degradations are quite farit and steady, translating into very short half Iives, of
0.75- 1 day for DCM and 1,2-DCA, 1-2 days for benzene, 2-2.5days for toluene and 2.5-3
days for TCE with 55 day lag period observed for 3 cm soil plugs. As pointed out earlier,
the beginning of VOCs degradation is associated with, (or it is believed not to occur
earlier than the VFAs fermentation induced) change of the porosity of the top layer, thus
the lag of 55 days is adopted for ail of the degradable VOCs, regardless of the fact that in
reality, the lags might vary and be dependant on the properties of each of the VOCs.
For 5 cm thkk soil plugs (Figs. 5.14.2 - 5.1 8.2) intnnsic degradation rates for
TCE, benzene and toluene seem to be slightly slower with longer half-lives. For DCM
and 1,2-DCA the rates are close to those inferred for the 3 cm thick plugs, however due to
data scatter they appear variable in tirne, particularly the rate for DCM due to lack of
information at the latest t h e of 271 days. Degradation in 5 cm thick plugs had a siightly
longer lag of 90 days, relative to 55 observed with 3 cm plugs. As evident from the Table
5.3, these lag periods were estimated based on the registered increase in porosity of the
top layer (layer 1). Generally, it seems that the outcome of degradation under simulated
conditions is not particularly sensitive to the lag period. niere is virtually no measurable
difference in the impact (ie. concentration in pore water) regardless of the duration of the
lag phase within the time fiame of 35 to 90 days, as shown in the Fig.5.14.l.l (a) & (3).
The sensitivity anaiysis performed for lag penods varied from shortest 35 to
recornrnended (and lower bound) 55 days indicated no difference in concentration vs.
depth profile for a 3 cm soi1 plugs in the cells II-1 & IV-2, as shown on the graph (a) in
the Fig. 5.14.1.1 (also see Fig. A5.3, Appendix 5, for details). Negligible difference
between the impacts after 55 and 90 day lags is also reproduced for a 5 cm thick plug of
the Ce11 3-1, as shown in the Fig A5.4, Appendix 5. On a small scale tested, simulated
fast rate of degradation, (Le. short half-Iife) dominates over the degradation lag, and other
mechanisms.
Regarding the rates, it is not clear why the rates of degradation happen to be lower
(i.e. increased half lives) at the top of the 5 cm thick plugs, since the thickness of the
"layer 1" with intensive bio-activity increased to 1.8 cm at 271 days compared with
"only" 0.65 cm thick "layer I"at 163 days. The number of HAB (-109 cji<!mL) is almost
the same for both 3 and 5 cm plugs at 163 and 27 1 days respectively, however both
number or SRI3 and ATP content with 3 cm plug at 163 days are significantly higher than
with 5 cm plug at 271 days, as seen in Figs 4.1 1.1-4.11.2 and 4.12.1-4.12.2 and discussed
earlier in Chapter 4. Regardless of the noted discrepancies in the observed half lives it is
striking that these first order rates are fast with short inferred half-lives even within the
time fiame of 271 days the test was mn. It is understandable that noted and srnail
discrepancies in half iives of severai days cause negligible diEerences in contaminant
impact which are, most of the time within standard deviation of the analytical
measurements, as seen in Figs. 5.14.1 & 2 for DCW 5.15.1 & 2 for 1,2-DCq Fig. 5.16.1
for TCE, Fig. 5.17.2 for benzene and Fig. 5.18.18~ 2 for toluene. (It is noted that the
receptor solution curves generated with consiaently adopted degradation parameters for
benzene in Fig. 5.17.1 and for TCE in Fig. 5.16.2 deviate From the data. Some of the poor
fit could arise From the analytical method, however the results for the replicates in Figs.
5.1 7.2 and 5.16.1 ailow for the use of degradation rates proposed in Table 5.3 .) It appears
that for the tested assembly and scale, degradation is dominant in the top layer. This
should not imply no degradation below the top layer but rather that the rate of degradation
below the top layer is much lower. The infiuence of the degradation in the layers below
the top interface is not prominent if the degradation proceeds at rate, much lower than one
which dominates the impact and dictates the shape of removai in the residual
concentration patterns. The sensitivity analysis performed for DCM and ce11 3-1 data in
the Fig. 5.14.2b (2), with different half-lives in the soil below the "layer 1" ranging fiom
365 to 1 O days combined with a fast rate in the top (half life of 0.75 day), indicates that
only half lives shorter than 50 days could significantly affect (reduce) DCM impact, as
demoostrated in the Fig. 5.14.l.l@) (and A5.4 in Appendix 5). Half-lives of similar order
would probably cause similar effect for the other degradable VOCs. It is generally
accepted (Balba & Nedwell, 1982; Sansone & Martens, 1982; Chapelle & Lovley, 1990)
that degradation and rnicrobial metabolism decrease markedly with depth, thus one
should not use arbitrarily selected rates.
It is noted that the diffision coefficient of D = 8.0 x 1 0 " ~ m'/s and linear sorption
K d = 1.6 cm3.'g (both higher that above), were used for DCM in the earlier work of Rowe
et al. ( 1997) and in Chapter 3 (see Figs. 3.3-3 -4; 3.6; 3 -8-3.1 1). Independent estimates of
linear sorption and diffusion inferred for environrnentaily representative concentrations
were not available for the preliminary degradation tests with DCM. These estimates are
proposed (and discussed, 5.2.3 and 5.3.3) in the current woric, and as more
appropnate, are used for the simulation of the DCM impact. It is recognized that the rates
of DCM degradation (i.e. half-lives) given in the prelirninary tests [Rowe et al. (1997);
Chapter 31 would have to be higher (half-lives shorter) as a consequence of newly
proposed (lower) sorption and (lower) diffision coefficients, if one was to produce
satisfactory fit-lines to then collected data. As stressed, previously used degradation rates
are hypothetical due to then lack of information on DCM distribution in the exposed soil.
The new evidence brought by intrinsic degradation teaing methodology provides the
rnissing information and consequently facilitates the approach to modeling. It is evident
that half-lives for DCM CO-metabolism simulated earlier and now are short, unmistakably
pointing to fast degradation which does dominate over other processes. It is the author's
opinion that this fact, in addition to the tested scale, and inability to actually measure the
eventual differences arising from imposed change of parameters, justifies not re-visiting
previous expenrnents and not re-simulating their outcome with newly proposed rates.
Based on the performed intrinsic degradation tests, it is considered that the VOCs,
supplied at relatively low (mg&) levels, are being CO-rnetaboiized under the established
lab VFAs fermentation mediated by mixed culture of soil indigenous microorganisms.
This simple approach to modeling of the intnnsic degradation using the first order decay
in the compacted soil appears reasonable, given the system dependent rates of
biochemicai reactions under difisive mass transport and srnall scale of the experiments
with only a lirnited number of data-points available for processing. With a few
exceptions, the theoretical lines generated by computer program POLLUTE v.6.5 give
very good fit to the collected data points. Merred half-lives and first order anaerobic
rates. possibly dominated by sulfate reducers, compare well with the values compiled by
Aronson & Howard, 1997 (Table A5.4, in Appendix 5) and appear to be higher but within
the same order of magnitude for the anaerobic first order rates of benzene and toluene.
Degradation of ethyl-beruene and xylenes under various conditions is reponed to be
faster than degradation of chlorinated aliphatics, however moa of the previously
published data refer to petroleum or solvents spills in more peneable soi1 than that
examined here. Short half-lives and fast rates observed in this work with chlorinated
aliphatics are considered to be realistic for the simulated conditions, and it could be
expected that this class of chernicals might have fast field removal rates in the MSWL
leachates under reducing conditions and/or under unlirnited hydrogen supply [as reactive
electron donor when other intennediates are used-up, while highly chlorinated
compounds are electron acceptors (Newell et ai., 1997; Wiedemeier et al., 1999), halo-
and chloro-methanes can be fermented as discussed in Chapter 21, as observed in triais
reported herein with KVL leachate.
Generally, for a MSW landfill, it might be expected that al1 of the tested VOCs,
being totally dissolved and at low initial levels, will eventually disappear or decrease to
safe levels when conducive conditions arise for their CO-metabolism. This aiso implies
different rates of removal and possibly retardation and long retention times, however
given the volatility and susceptibility to biodegradation of these chemicals it is
hypothesized that they would not penia long afler the steady conversion of bulk
dissolved organic carbon (Le. late fermentation and beginning of methanogenesis) from
the leachate has been established. The irreversible removal of organic pollutants including
VOCs is initiated in the waste fiil, thus the levels of the VOCs that could diffise into (and
out of) compacted clay liner could already be reduced due to degradation. It is possible
that conducive conditions could eventually develop for fùrther degradation and reduction
of the VOCs in the course of very long retention times even under adverse mass transfer
conditions in the compacted soil, however the rate of such degradation is expected to be
significantly lower than the rate in the waste fill.
5.5 Summary and conclusion^
Based on the results of the sixteen (16) lab tests performed with compacted
clayey soi1 under dominant difisive transport and continuous supply of fermentable
substrates and volatile organic cherni-cals it can be concluded that DCM, 1,2-DCA, TCE,
benzene and toluene are subjected to intrinsic degradation mediated by soil indigenous
microorganisrns. Afier a lag of 55 to 90 days, removal of contarninants was fast, and
predorninately located at the soiWsolution interface where the best conditions have been
created for gas fermentation, mass transfer and unhindered rnicrobial growth. Half lives
for the degradable VOCs are short and in the range of 1 to 15 days. Based on these
findings From the laboratory tests which simulated successfiilly very adverse conditions
for degradation, it is hypothesized that conducive environment might develop in the field
waste disposal facilities as well. The prospects of field scale intrinsic degradation seem to
be realistic and efforts should be made to empioy every well designed landfill not only as
a safe engineered containment facility, but also as a reliable naturd bio-reactor. It is the
author's opinion that these conditions favorable for the degradation of the volatile organic
chemicals will prevail and possibly even be practically effective mainly in the fennenting
waste fil1 where the fermentation can be initiated and maintained for some time d u h g the
landfill seMce life. When present at low, but hannful levels as sometimes found in
landfill leachates and, more importantly as sole source of carbon and energy, (that is
when VFAs or some other preferential Clenergy source is unavailable to initiate the
microbial activity), the biodegradation of the examined volatile organic compounds in the
oxidized natural confining deposits is, however less feasible. It is noted that in theory,
such degradation is thermodynarnically possible and may be initiated, as demonstrated
under different contamination and hydrogeologicd conditions and compiled for numerous
laboratory and filed cases by Wiedemeier et al., 1999. It is generaily considered that the
mass transfer limitations imposed by compaction and matnx-diffusion will severely
hinder any possible biologically driven reaction. In the absence to the contrary, it is
reasonable io proceed with conservative approach in the design and modeling and either
not consider VOC degradation in deep compacted clay liner(s) or to count only on the
very long degradation lags and very low degradation rates, as discussed in Chapter 6.
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Table 5.1 Sunimary of sorptiun paranieters for the VOCs and Halton Till
Dic hloromethane @CM)
1,2-Dichloroethanc (1,2-DCA)
Trichloroethene (TCE)
Benziic
Toluene
Linear Sori)tion
observçd 95% CI^ for
O. 574.68 1.4 1-2.10
0.7 14.83 0.37-0.5 1
1.26-1.41
0.34-0.44 0.6 14.7 1
0.90- 1 .O7 1.33-1.51
1.17-1.47 1.78-1.M 2.28-2.44
1.99-2.1 1 2.64-2,s
1.60- 1.75 2.13-2.36
Cwl range ItigU
74 - 3333
156 - 4605
9 1-5597
95 - 4377
65 - 6301
63 - 3361
65 - 3447
Freundlich Cd, = & x Cmi.
Langmuir Cd, = S,,x bxCJ1 +bxCd
' colculated from Karicklioff ct al. ( 1979) and L y i i ~ n ( 1980) for f, = (0.144.45) %; ' IUkp) = lcin3/g J, CWd Ipg/kg J; ' 95%CI = 95 % confidence interval; *significant devistion from model, ( b a d an tcsting of randorn distribution of residuals around the curve for the nidcl);
Table 5.2 Summary of diîïusion and linear sorption coenicients Cor the VOCs and Halton Till used for rnodeling
Dichloroniethane (DCM)
1,2-Dichloroethane (1,2-DCA)
Triclilorwthcne (TCE)
Benzene
Toluene
Ethyl-knzene
m&p-Xylenes
O-Xy lcne
D, Diffusion coefficient in soi1 pore water, [ma/sl
2.5 x 10-'O
2.5 s 10-'O
2.5 x IO-"
2.5 x 10*1°
2.5 1; 10'''
2.5 r IO-"
2.5 x 1 ~ "
2.5 x 10-'O
B,,, Diffusion coefficient in g las
disk, [ m2/sj ' K4, liriear sorption
codlcient, [cm3/gl
0.0s
0.25
1.35
0,SO
1.SO
2,35
3.00
2.40
' îIme as diffusion coefficient in the frec solution, values taken îroni Yaws ( 199 1); ' from Y 5% CI and from KarickhoB et al. (1 979) and Lyman ( 1 980);
pK, fi = 1 + - , calnilaicd for dry dciisity p = 1.80 g/ciii2 and averagcd porosity n = 0.34 for Halton tilt compactcd in Iaboratory; 11
1 "?+ri o a o o o
-*O-- 21 11 Y ! ? ? ? ? o a o o o
e r e
L L L ;1 j $ j
DCM in the Solution [pg/L]
O IWO 2000 3000
1,2-DCA in the Solution fpg/L)
TCE in the SoIution [pg/Lj
Fig. 5.1. L SORPTION of CHLORINATED ALIPHATICS onto the HALTON TILL: linear isotherms and 95 % confidence interval for K, (a) DCM @) I,2-DCA (c) TCE
eri
I$= 8.0 Ukg n = 0.68 (Freundlich)
I y C, range (74 - 3333 ) &L
DCM in the Solution [pg/L]
1.2-DCA in the Solution (pg/L]
- K,=8.1 L'icg n=0.75 A
Cd range (86 - 4605) pg5
1--
TCE in the Solution [pg/L]
Fig.5.1.2 SORPTION of CHLORINATED ALIPHATICS onto the HALTON TILL: Data with Freundlich and Langmuir isothems (a) DCM (b) 1,2-DCA (c) TCE
- O 500 1 O00
Benwne in ~ h e Solution [pg/L)
O 1000 2000 3000
Et hyl-Bewne in the Solution 1 pg/L J
Toluene in the Solution [ d l
ni&p-Xylcnes in the Solution IpglL]
O 1 O00 2000 3000
O-Xylcne in the Solution [pg/LJ
Fig. 5.2.1 SORPTlON of BTEX ont0 the HALTON TILL: Data with linear isothems (a) benzene (b) toluene (c) ethyl-benzene (d) xylenes; al1 lines shown with 95 % CI for K,
Benzene in the Solution (vg/L]
Eihyl-Benzenc in the Solution [ p a l
O 1000 2000 3000 4000 SOQO
Toluene in the Solution [&LI
iiitkp-Xylenes in the Soluiion [p@]
O-Xylcnc in the Solution [ci@]
Fig. 5.2.2 SORPTlON of BTEX ont0 the HALTON TILL: Data with Freundlich and Langmuir isotherms (a) benzene (b) toluene (c) ethyl-benzene (d) xylenes
Fig. 5.3 DIFFUSION OF DCM THROUGH HALTON TILL (a) source SS and receptor RS solutions @) depth profile
Fig. 5.4 DEFUSION OF I,2-DCA THROUGH HALTON TILL (a) source SS and receptor RS solutions (b) depth profile
TCE in Porc Watcr [mgt]
Fig. 5.5 DIFFUSION OF TCE THROUGH HALTON TILL (a) source SS and receptor RS solutions @) depth profile
Fig. 5.6 DIFFUSION OF BENZENE THROUGH HALTON TILL (a) source SS and receptor RS solutions @) depth profile
Time [daysl Tolucne in Porc Water (mg/L]
Fig. 5.7 DIFFUSION OF TOLUENE THROUGH HALTON TILL (a) source and receptor solutions @) depth profile
Timc [days] Ethyl-Eknzme Ui Parc Wakr [m&]
Fig. 5.8 DIFFUSION OF ETHYL-BENENE THROUGH HALTON TILL (a) source and receptor solutions @) depth profile
Time [days] m8tpXylencs in P m Waicr [mgLI
Fig. 5.9 DIFFUSION OF m&p-XYLENEs THROUGH HALTON TILL (a) source and receptor solutions (b) depth profile
Timc [days) +Xytcne in Porc Wata [mgLj
Fig. 5.10 DIFRISION OF O-XYLENE Tl (a) source and receptor solutions (
ROUGH HALTON TILL b) depth profile
Time [days]
Fig. 5.1 1 CONCENTIUTION OF VOCs in SOLUTION: monitoring stability of dissolved DCM, 1,2-DCA and TCE in time
Fig. 5.12 CONCENTRATION OF VOCs in SOLUTION: monitoring stability of dissolved benzene, toluene and ethyl-benzene in time
Fig. 5.13 CONCENTRATION OF VOCs in SOLUTION: monitoring stability of dissolved m&p-xylenes and O-xylene in time
..... D = 2.5 x 1 ~ " ~ r n ~ / s
K, = 0.05 m'tg n = n, without ciqdation alh U- f & [V-2 (avmgtd)
with &gradation crlls rI-1 & IV-2
Time [daysl
10 2 (b) DCM [mg/L] in the pore water for D = 2.5 x 10- m 1s and
8 Cell 1-2 @ 29 &y
. - . - a witllout degradation 1-1
--- without degradation 1-2
A CeIl 11-1 3 162 &y Cell IV-2 163 &y.r avemgcd for 162.3 &'
S . . . . without dqphtiorl
-- with &gradation
DCM degndstim anrsidcred in the hyer 1 ody and modticd (aioag the dcpth and ai tirne) as: l'stage: t,, = (0.75 - 1 ) &y m che L i y u 1 (0.5 top cm ) afkr 55 &y-trg; 2"ltrgc: @ 140 day m the h y u 1 (0.65 an) $, = 0.75 &y a d D = 4 x 10"' dk pomsity mitidy 034 and ss caiculrted a! dcsigmrted tinw
Fig. 5.14.1 INTRINSIC DEGRADATION OF DCM in HALTON TILL : 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
P SOURCE SOLUTION
A 6 6
4 (a)
- D = 2.5 x 10''~rn'/a K,= 0.23 cmJig n = variable
O Celb 1-1 & 1-2 1 A Cellrii-1 &IL2 0
with &gradation œlIs 11-1 & IV-2
2
1
O 30 60 90 120 150
Time [days]
V CtlLIII-l&III-2 RECEPTOR SOLUTION A o.... .*.. - 0 Celb N-1 & CV-2 i
O
- O
&.--
-IO 2 (b) 1,2- Dichloroethane [rngiL] in the pore water for D = 2.5 x 10 rn 1s and K,= 0.25 cm31g
O M - " * , I 1 r r
O Cell 1.1 cg 23 days
O Ccll I-2 @ 29 days CcUIV-l@ t l 7 h p A CdlII-1 @ 162 days Cet1 11-2 @ 118 dap Cdl IV-2 !a, 163 &y averagcd for 117.5 day avcngcd for 162.5 da.
0 1 2 3 4 5 0 1 2 3 4 5
1.2-DCA âcphUcm anuidatd m tbe L.ya 1 only and modelai (dong the dcprh and in t h e ) as: l'stage: S.,, = 0.75 &y m the Laya 1 (- 0.5 top an ) rttn 55 dry-lag
2*aig+: @ 140dqs m h L . y a l ( 0 . 6 3 m ) ~ = 0 . 7 5 & y u i d ~ = 4 x 1 0 " ~ ; initiai pros* 0.34 and u cdfulakd d a i p k d h c r
Fig.5.15.1 INTRINSXC DEGRADAnON OF 1,2-DCA in HALTON TILL: 3 cm compacted plugs; (a) source & receptor solutions @) depth profiles
0 30 60 90 120 150
Time [daysl
-10 2 b) TCE [mgL], in the pore water for D = 2.5 x 10 m 1s and K, = 1.3 5 cm31g
..... D = 2.5 x 10-'~rn~!s K, = 1.35 cm'/g n = n ~ ,
without dçgradatim LI-1 & N-2
D = 2.3 x 10* '~m~/ s K, = 1-35 anJlg n = ib,
without Il-1 & N-2
-- D = 2.5 x l~'~rn'/s
b
SOURCE SOLUTION
m i
v 19 + m O
O Cell iV- 1 @ 1 17 day~ Ccll IL2 GJ 1 18 day rvcragcd fur 1 17.5 da.
O 1 2 3
O .II d 2 - f - % d)
5 .E 1
ul
A Cell IL1 @ 162 dap Cc11 IV-2 t@ 163 day~ s v q c d for 162.5 day
O Cells 1-1 & 1-2 (a) A Cclh II-1 & II-2 v Cclb III-1 & III-2 RECEPTOR SOLUIlON ....-. - O Cclls Pl-1 & N 2 A.. .*..--
O F
m . . - - without degradation
K, = 135 n =
O with dtgradstion u-1 & IV-2 . *
Fig. 5.16.1 INTRINSIC DEGRADATION OF TCE in HALTON TILL: 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
oc+-----* f t 1 (90 day lag in the 2%tagc)
2 - O Cclls 1-1 & 1-2
RECEPTOR SOLUnON . . - * . . . . O
A CCUS il-1 & iI-2 .. .- v Ceils ~n-1 a tu-2 . .& ' A
1 - O CeNs IV-1 ~t N-2 a n
1 1 1 1
K, = 0.5 cm31g n =
without(.kgm&ion œlfs a-1 & IV-2
- D = 2.5 x 10"~m'!s K,= 0.3 cmSfg n = variable with degradatiai e l b II- 1 gt IV-2
O 30 60 90 120 150
Time [days]
-10 2 @) Benzene [ m a ] in the pore water for D = 2.5 x 10 m 1s and K, = 0.5 crn3Ig
0 CeIl 1-1 4 23 Cd1 111-1 @ 63 &r O Cell 1-2 I@ 29 da'
B CeU W-1 @ 1 l i &Y, A Ctll II-1 3 162 d v Ceii II-2 @ 1 18 days Ccll IV-2 @ 163 dayJ rvcragcd for 117.5 âays avmgcd fot 162.5 day
0 1 2 3 4 0 1 2 3 4
Eknzene dcgdalion anisidatd m the Liyn 1 only and modclcd (dong the dcpth and m tirne) as: lasIage:$, = 1 dayinthefryu l (0 .5 topcm).ftaSS day-1%
2*stage: @! 140 days intbc Liys l(0.65 cm)$,= 1 dayand D = 4 x IO-" m2k pomsity initidiy 034 d u crlcutitcd at d a i d tkncr
Fig . 5.1 7.1 INInINSIC DEGRADATION OF BMZENE in HALTON TILL: 3 cm com~acted ~iuas: (a) source & receptor solutions (b) depth profiles
-10 2 (b) Toluene [mgL], in the pore water for D = 2.5 x 10 rn /r and K, = 1.5 em31g
O CeIl 1-1 (3 23 da^ v a i l m-1 63 &y 0 CeU IV-1 @ 117 di)J A Ce11 11-1 (9 162 day 0 CeIl 1-2 29 dap Ceii II-2 @ 118 day3 Ce11 IV-2 3 163 days
ivemgcd fa 1 1 7.5 dayi avcraged for 162.5 day
Fig.5.18.1 INTRINSIC DEGRADATION OF TOLUENE in HALTON TILL: 3cm compacted plugs; (a) source & receptor solutions @) depth profiles
Time [&YS]
- "J 3 3-0 : E Y
e SOURCE SOLUTION (a) V1 = 0 2 . 3 , O . - w 3 - O 2.0 VI -
-10 2 b) Ethyl-Benzene [mgR], in the pore water for D = 2.5 x 10 rn Is and K, = 2.3 5 cm3lg
b)
5 1.5: c . I Q) C 1.0 P) N c
O Cell I-1 9 23 days (b Cdl1-2 @ 29 day
Fig. 5.19.1 INïRINSlC DEGRADATION OF ETHYL-BEN- in HALTON TILL : 3 cm compacted plugs; (a) source & receptor solutions @) depth profiles
__c-4-C K, = 1.62 cmJig n = II-, _-/--- -
)r ---- CC---
b * ' n a 2 o . o ~ , - - - - - Z ? ,O d 1
a
. O Cellrr 1-1 & 15 * A Celb U-1& II-2 1 O Cclh III-1 & III-2
0 Cells IV-1 & IV 2 RECEPTOR SOLüTION _--- _----
---. D = 2.5 x 10 '~~rn~ l s IV-2 K, = 1.62 cm'lg n = n,
D = 2.3 x I ~"~rn'/s IV-2
10 2 (b) rn&p-Xylenes [rng/L], in the pore water for D = 2.5 x 10- m Is and K,., = 3.0
4a
5 1.5 c . - m
1 . 0 - E: 4a - )r X 0.5 I
..... no degradntim Ceil1-2 . . . . . ~ d t g r r d r t i ~ n ~ c ~ m - 1 '"" n o ~ ~ C e l l f V - 1 - " " no degradation Ce11 U-1
--- no &pdhm Cd1 1-1 --a no degradru011 CcU II-l --- no dcgdauan Ccll W-2
C I - O Celb 1-1 & 1-2 A Celb 11-1 crt II-2 v Celb i?4& Ut-2 O Celb IV-1 & IV-2 RECEPTOR SOLlJTION _---- - _--- _----
_--- _---
Fig 5.20.1 INTRINSIC DEGRADAnON OF m&p-XYLENEs in HALTON TILL : 3 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
a ---- P --- ___--- - 8 a
0.0- $ u 1 l I n. h
Time [days]
Q)
5 c .a
0 1 c 0 - X
-10 2 b) O-Xylene [mfl], in the pore water for D = 2.5 x 10 m 1s and K, = 2.41
Fig. 5.2 1.1 INTRINSIC DEGRADATION OF O-XYLENE in HALTON TILL : 3 cm cornpacted plugs; (a) source & receptor solutions (b) depth profiles
L
O CeIb 1-1 & I-2 A Cells 11-1 & ii-2 v Celb 111-1 & 111-2 - KI Celb IV-1 & IV-2 __--
U C E p i o R YKMOz-_--------- ---- d _--- ----- * c i b
--- D = 2.5 K 10"~m~!s K, = 2.41 anJfg n = n,
D = 2.5 K 10' '~rn~~s K, = 2.41 d l g n = n,
8 * 0 0 - CH---c- a A 0
O 1 - I 1 1 I
10 2 O>) DCM [mg/L] in the pore water for D = 2.5 x 10- m 1s and K, = 0.05 cm3/g
e Cctl 1-2 @ 33 dar O Ccii 1.1 @ 49 days
I Cell4-2 6jî 20 1 daqr Ccll3-2 6J 202 da)% Ccll2-2 @ 204 da- ivemgcd for 202 &y
0 1 2 3 4 5
DCM dqadaiim amsidaed m tht hyer 1 d y d modeled (Ilmg thc and in tirne) as: 18stage: $,= 5 &y m the hyer 1 (top 0.5 cm d y ) rfta 90 dry-1%
2*aagc: g 163 dayr in tbe ï ~ y u 1 (1.5 cm): t,,= 5 days and D = 4 x 10"~rn'ls
Fig. 5.14.2 INTRINSIC DEGRADATION OF DCM in HALTON TILL : 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
DCM in pore water Lm@]
-7- /-- (a)
DCM in pore water [rng/L]
A
Data . . . . 55 ciay lag (Layer 1 only) - 90 day lag (Layer 1 onIy) - - - 90 day lag :
tll, 0.75 d (LL) 365d (bel~w L 1)
-- 90 day lag t1,2= 0.75d (Ll) 1OOd @elow Ll)
--- 90 day lag t l r 0 . 7 S d (L 1) SOd (be10w L 1)
---- 90 day lag t 112= 0.75d (L 1) 10d (below L 1)
A Cells II-1 & IV-2 @ 162.5 days t,, = 0.75 d in Ll - 55 days lag
Fia. 5.14.2.1 Variation of Urtrinsic degradation parameters :
13 1 - - - - - - tln t,, = = 0.75 0.75 d d in in L1 L1 - - 1 5 35 day day lag lag
- (a) effect oflag duration on DCM imapact elaborated
for the cells II-1 & N-2 @ 162.5 days shown in Fig. 4.14.1@)(4)] (b) effect oflag duration and half-lives on DCM impact elaborated
for the ceU 3-1 @ 128 days shown in Fig. 4.14.2@)(2)]
- - O 30 60 90 120 150 180 210 240 270
Time [days)
n T 7 A T
d 5 ~ $ p p p ~ 9 : * g ' : U E i r p * 6
-10 2 (b) 1,2-DCA[mg/L] in the porewater for D = 2.5 x 10 m 1s and K,=0.26 crn31g
A C e I 1 2 - l f ~ 2 7 1 ~ O Cc11 4- 1 ig 271 days
. D = 2.5 K 1 O.'~&/S K, = 0.26 m'lg n = n, without dqpb ion cell4-1
D = 2.5 x 10~ '~rn~ ls K, = 0.26 cm'lg n = n, without &gradation
2 4 - O .- Y
s d
2 3 : 0 5 C 2 - .-
Fig. 5.15.2 INTRINSIC DEGRADATION OF 1,ZDCA in HALTON TILL: 5 cm compaded plugs; (a) source & receptor solutions (b) depth profiles
3 - SOURCE SOLUTION (a)
O Celb 1-1 & 1-2 A Cells 2-1 & 2-2
6 U 0 I r
CI
nh
V Ctlfs 3-1 & 3-2 ctll4-1 RECEPTOR SOL~TION
O Cclls 4-1 & 4-2 - initidiy D = 2.5 x 10"~rn~'s K,= 0.26 cm'/g n = variable with dcgdation ccll4-1 (sec tefi)
c ---. D = 2.5 x 10-'~rn'fs O . -.. - -A O
L ' 1 1 I;, = 1.35 cm'/g n = a, O without &gdation 4-1
-
Time [days]
- SOüRCE SOLUTION 1 -
10 2 (b) TCE [rnfl] in the pore water for D = 2.5 x 10- m l s and K, = 1 .35crn31g
I CcU4-2@201day3 Ccil3-2 @ 202 day3 Ceil 2-2 (5J 204 day~ r v q e d for 202 days
O 1 2 3 I
7 a: I
..... uithout &gradahoa 2-1 --- without degradation 4-1
-- ~ i t h dr+~n 4-1
TCE degdation anrridned m the hycr 1 only and modclcd (dong the depth and in the) as : 1' stage: $r, = 20 d m the L*ya 1 (top 0.5 cm d y ) a f k 35 &y-1%
2*s(.gc: @go d inthetryer 1(toQ l.O~m)$~= 15d
J**. @165d inihc~y~~(l.~cm)n=~.55orrralai~~=4xl~'0m2~smd~,=l5d:
4brgc: $ UO d in the Lycr 1 ( 1.8 cm) n = 0.63 a u almlala& Du = 5 x ~O"~rn'ts and t,, = 15 d: initial potasity 0.34 anci as dnilrtcd a! daipfd tima
Fig. 5.16.2 INTRINSIC DEGRADATION OF TCE in HALTON TILL : 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
SOL'RCE SOLLION (a)
7
b
D = 2.5 x 10"0m2/s K, = 0.5 m31g a =
wiihout dqpdaiaa al ls 2-1 & 4-1
D = 2.3 x 10"~rn~/r O Celb 1-1 & 1-2 K, = 0.5 m'tg a = n, A Celh 2-1 & 2-2 without dqdat ioa v Celb 3-1 & 3-2 cclls 2-1 & 4-1 (3 Cells 4-1 & &2
RECEPTOR SOLüTiON initially D = 2.5 x 10"~m'~s K,= 0.5 cmllg n = =wiablc withdegrsdriiori alis 2-1 & 4-1
Time [days]
-10 2 (b) Benzene [mg&] in the pore water for D = 2.5 x 10 rn 1s and K, = 0.5 cm31g
90 &Y for dl lina Ll = 1.0 cm @ 128 dayl
CellC2@201&yS Cdl3-2@202 dayl CM 2-2 @ 204 drys rvaagal for 202 Clay
O 1 2 3
A CeIl 2-1 -270 dayl CeIl 4-1 @ 27 1 dayl ivmrgcd for 270.5 d a y
&mrne dtgrahtida cornidercd m ibc kycr 1 d y and modeIlcd (dong ibc depth and in time) a: 1'stngt:t,,=7drys9ithe~~1 (1.0cm)rftt90diy-Ii& ~"'stagc: 165 bys in thc t y c r l(1.5 an) b= 7d0ys and O = 4 x 10"Orn~~~
id*: @ 230 hys mLbc h y a 1 (1.8 cm)Gn= 7&ys rd D = 5 x 10'~~m'lr. porosity: initiaily 034 and as calculaicd at dcsigr~trd t ims
Fig. 5.17.2 INTRINSIC DEGRADATION OF BENZENE in HALTON TILL: 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
D = 2.5 x 10-'~rn'/s E;~ = 1.5 cm'fg n =
without degradation alls 2- t Bt 4-1
- initially D = 2.5 x 10"~rn'h K, = 1.5 cm'ls n = variable
Time [daysj
10 2 (b) Toluene [mg$] in the pore water for D = 2.5 x 10- rn Is and K, = 1.5 crn3lg
Cell4-2 @ 20 1 &y4 Ce11 3-2 @ 202 days Ce11 2-2 @ 204 days avcngcd fot 202 days
A Cd1 2- 1 (a 270 &y C d 4- 1 @ 27 1 dry3 avcraged for 270.5 day
O 1 2 3
I I -
Tolucrie degndalion amsîdaed in I& iaycr 1 d y a d modelai ( dong the dcpth and in time) as: l'stage: t,,= 15dayinthcLqcr 1 ( t o p 0 . 5 c m ~ ) a î k U d a y - 1 % 2*stage: @) 165 d.). P h hycr 1 (1.5 un): g= 10 days iod D = 4 x 10-'~m'1'r.
3"aige: @ 230 dar in& hyu 1 (1.8 an): rin= 10 days and D = 5 x 10 '~~rn~fs porwi- initidly 0.34 iad m dculatcd u desipkd tùnts
Fig. 5.18.2 INTRINSIC DEGRADATION OF TOLUENE in HALTON TILL: 5 cm cornpacted plugs; (a) source & receptor solutions (b) depth profiles
RECEPTOR SOLüTiON
Time [days]
10 2 (b) Ethyl-benzene [rng/L] in the pore water for D = 2.5 x 10' m Is and K, = 2.35crn"gj
Fig. 5.19.2 INTRiNSIC DEGRADATION OF ETHYL-BENZENE in HALTON TILL : 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
-10 2 (b) m&p-Xylenes [ m a ] in the pore water for D = 2.5 x 10 m /s and K, = 3 .O
. . . . . w ~ o a ~ c l l 4 - 2 ----• no dtgradation CefI 2-1
--- no degnhtion Cc11 3-2 --- no degradation C d Cl -- lm dcgnditioa CeIl 2-2
Fig. 5.20.2 llUïRNSIC DEGRADATION OF m&p-XYLENEs in HALTON TILL : 5 cm cornpacted plugs; (a) source & receptor solutions @) depth profiles
O 30 60 90 120 150 180 210 240 270
Time [days]
3.0 - SOURCE SOLLITION
c O 2.0 . m eC 3
e O Cetisl-1&t-2 (a) . I ---
V Celh 2-1 & 2-2 D = 2.5 x L O * ' ~ ~ ' I S b) c A Cclls 3-1 & 3-2 K, = 2.4 crn3ig n = n,, 1 .0 -
C1
I , 0 Cclls 4-1 & 4-2 x" f D = 2.5 x 10"~m'/s
10 2 (b) O-Xylene [rngL] in the pore water for D = 2.5 x 10' m /s and K, = 2.4
-- 00 dcgnditioa ccu 2-2
K, = 2.4 cm3tg a = a, 8 0.2
Fig 5.2 1 -2 INTRINSIC DEGRADAnON OF O-XnENE in HALTON T L L : 5 cm compacted plugs; (a) source & receptor solutions (b) depth profiles
0.0- A, __--- n a
- RECEPTOR SOLUTION _ _-__-------
CHAPTER 6 PREDICTION OF CONTAMINATION IMPACT FOR SELECTED
ORGANIC CHEMICALS MIGRATING FROM A HYPOTHETICAL LANDETLL
6.1 Introduction
In this chapter ditfusion, sorption and first-order rate of degradation parameters
deduced from the laboratory experiments are used to examine the potential contaminant
impact of a hypothetical landfill on an underlying aquifer. For a simple design case, 1-D
transport of DOC-VFAs (dissolved organic carbon as volatile fatty acids),
dichloromethane @CM), benzene and xylenes has been simulated using computer
program POLLUTE .v 6.5 (Rowe & Booker, 1999). The barrier design studied involved a
primary leachate collection system (PLCS), a compacted clay liner (CCL) and a
secondary leachate collection system (SLCS) overlying a naturai till confining deposit.
Attention is focused on the Iag period(s) for degradation (i.e. "the time needed for the
degradation to stan") and the rates of degradation (i.e. half-lives) for the selected organic
contaminants in the observed system and the influence that these parameters might have
on the contaminant migration and consequent impact on the ground water pollution.
6.2 Bypothetical hydrogeological setting and choice of contaminants
A schematic of the hypothetical landfill examined is given in the Fig. 6.1. Data on
the geometry and hydraulic parameters are summarized in Tables 6.1 and 6.2. It is
assumed that this hypothetical landfill cm accommodate -20 m thick refuse at apparent
density of - 750 kg/m3. The primary leachate collection systern (PLCS) consists of a
penneable, 0.3 m thick coarse-stone drainage layer with collection pipes placed at the
landfill base. The bottom of the landfill is underlain by a 1 rn thick compacted clay liner,
with porosity of 0.34 and hydraulic wnductivity of kccL = 10''~ d s . SLCS (or HCL,
hydraulic control layer, Rowe et al., 1995), another permeable and initially saturated, 0.3
m coarse-stone drainage layer (with collection pipes) is placed undemeath the clay Liner.
This engineered barrier is assumed to rest on a natural tiil deposit having the same
hydraulic conductivity as the clay liner (Le. = 10-'O m/s). The hypothetical landfill is
examined as a major potential source of pollution for 1 m thick sand aquifer with porosity
0.3 and horizontal Darcy velocity of 1 m/cr at the up-gradient edge.
At the beginning of the disposal activity, the PLCS is assumed to be controlling
the leachate head (mound) to 0.3 m above the (compaaed) clay liner. Gradually this
fûnction becomes impaired due to the clogging of the permeable drainage blanket (see
Fleming et al., 1999) and the PLCS fails to keep such low and hydraulically favorable
leachate head over the clay liner. However, the fluid removal from this engineered system
remains fùnctional. Initially, during the normal operation in the PLCS the rate of build-up
of the leachate mound is controlled by the rate of infiltration through the landfill cover
(taken as 0.15 mia) and the downward Darcy velocity through the clay liner. The leachate
mound height (initial 0.3 m and final 15 m) together with other parameters given by
geometry of the hypothetical hydrogeological setting is used to calculate hydraulic
gradients and downward Darcy velocities, as listed in Table 6.2. As can be seen, the
initial (vertical) velocity is zero. This is diaated by the position of the PLCS base and the
(constant) level of the hydrostatic head 6.6 m above the top of the aquifer in the Fig. 6.1.
The PLCS is modeled with a senice life of 100 years as mggested by the Ontario
Regdation 232/98 (MOE, 1998). It is assumed that the mound rises linearly to its
maximum within 10 years upon the failure (clogging) of the PLCS (Le. after 110 years),
resulting in the unfavorable advection (downward velocities), which could facilitate the
contaminant transport and increase the ground water pollution. In order to rninimize this
adverse condition ansing from the leachate mounding, the downward flow from the clay
liner is de-coupled and contarninated fluid removed fiom the SLCS. The rates of
(horizontal) removal from the SLCS and the venical velocities in low k-strata are
calculated from the continuity of flow in the SLCS [so that "flow-iny' (vcc~ + vnll)xLx W
equals "flow-out", vxcsx Wxhscs] and their values and directions are given in the Table
6.2. By analogy, the Darcy velocity in the aquifer at the down-gradient edge of the
landfill, v,, ,,, is also calculated taking the continuity of £low in the aquifer into
consideration. For the purpose of this study oniy, the flow generated upon failure is
assumed to proceed in the direction of the existing flow in the aquifer and under
conservative mixhg conditions with zero hydrodynamic dispersivity (a = O, Eq. 1.1).
For al1 impact simulations, the rnass (in the waste or source) of each species
examined was considered to be finite. As a result of difisive-advective transport (and
even in combination with attenuating mechanisms) at such initiai boundary condition, the
aquifer generally becornes contaminated. This Uiflowing contamination will be
discharged further at the rate controlled by flow rate in the aquifer (va, ,3. The maximum
concentration, C m , occurrhg at a particular time at the dom-gradient landfill edge (i.e.
at the landfill length of L = 500 m) is recognized by the program POLLüTE features and
subsequently checked against the maximum allowable Limit (O. Reg. 232/98) for selected
organic pollutants. If the selection of the engineered components and their performance is
satisfactory, the recorded Cm never reaches (or exceeds) the designated lirnit as the
concentration of contaminant gradually drops due to diminishing mass input from the
source (Rowe et al., 1995).
This hypothetical landfill provides reasonable protection against chlonde
contamination if the SLCS, started at 100 years, stays in operation for at Ieast 260 years
(i.e. if the fluid is removed fkom this engineered layer at max rate of - 60 d a , see Table
6.2). Chlonde, as a conservative species subject to advective-diffusive transport only, will
be neither retarded by sorption nor destroyed by decay, thus its contamination usudly is a
determinant of the impact on a clean aquifer. For this case and with initial chloride
concentration of Co = 2.5 g/L, the maximum impact is C,, = 121 mg/L, which is
allowable based on the reasonable use policy as defined in O. Reg 232/98 (MOE, 1998,
also see Table 1.2, Chapter 1). Thus, this "design" could be considered as adequate and
examined funher for the selected organic contaminants.
VFAs (expressed as DOC), dichloromethane @CM), benzene and xylenes are
selected as representative of organic contaminants in the municipal solid waste landfill
leachate. VFAs-DOC is an indicator of major readily degradable contaminants, othenvise
very mobile and not retarded by sorption. DCM and benzene are modeled as priority
micro-pollutants associated with health risk, exhibiting almost no or very low sorption
(retardation) ont0 soi1 respectively. Xylenes are chosen as quite cornmon although not
ïisk-associated (micro-) contaminants in landfill leachates charactented with moderate
sorption (as volatile organic chemicals) and low initial concentration. Sorption of DCM,
benzene and xylenes is modeled as linear with coefficients deduced previously and given
in the Table 5.2 Chapter 5 .
The upper bound for diffusion coefficient(s) of D = 5 x 10 -'O m2/s, inferred fiom
the laboratory difision and intrinsic degradation tests (see Table. 4.1, Chapter 4 and
Table 5.2, Chapter 5) is used for simulating the impact of al1 selected organic
contarninants. This value results in higher (more conservative) concentrations, but is also
considered a realistic design parameter given the large-scale heterogeneities. The initial
concentration of a particular contaminant in leachate is either taken from the existing
regulations or chosen according to the other sources or experience. For DCM, the
recommended Co listed in the 0. Reg 232198 is 3.3 mg/L, and is conservative and
simulated as such. The initial concentration given by O.Reg 232/98, for benzene is only
Co = 0.02 mg/L (i.e. 20 pg/L). This gives an impact of only 0.28 pg/L (without sorption
and degradation). To explore the possible implications of a higher concentration (e.g. due
to accidental disposal of hazardous waste in a MSW landfill) and given the very low and
stringent DWO of 5 pgZ (both MOE, 1998 and NPDWRs USEPA, see Table 1.2) for this
carcinogen its rnodeling was also conducted for a higher Co = 1 mg/L.
The MOE does not have specific drinking water objectives for VFAs. Thus, for
the purpose of this study, dissolved organic carbon DOC, is considered an appropriate
surrogate. The drinking water objective @WO) for DOC (recognized as aesthetic, Le.
non-health related parameter) is set to 5 mg4 by Ontario MOE, 1998. Since the landfill
standard (O. Reg 232/98) does not give initiai waste-fiIl concentration for VFAs or DOC,
a value 2.4 g/L DOC as VFAs was assumed to be consistent with the levels tested in the
intrinsic degradation expenments. (This was seiected to simulate a typical level in KeeIe
Valley Landfill Ieachate). This initial concentration corresponds to about 17 g/L COD
which is high at the terminal (methanogenic) stage of landfill operation but, as such, it
will yield conservative impact prediction. The initiai concentration for xylenes in not
available in the regulations and the Ievel of 1 mg& is used to be conservative in this
hypothetical case. It is noted that concentration of xylenes in KVL Ieachate is 200 - 400
pgL when detected. The dnnkhg water objective for xylene is taken as O. 15 r n f l (MOE
1998; 0. Reg. 232/98).
The objective of the following study is to give an indication of the influence of the
degradation lag periods and degradation rates on groundwater contamination. The
hypothetical landfill is intentionally exarnined as a straightfoward case. As such, it
features a set of combinations for which the interference of the many recognized
influentid parameters (see Rowe et. al, 1995) could be successfully eliminated and
subsequently consideration of the parameters more favorable than considered herein (e-g.
lower Co, or thicker natural aquitard) would result in the obvious lower impact. In that
case, the simulated changes. to be presented, could be assigned to the variation and
uncertainties due to degradation parameters.
6.3 Biodegradation parameters
The major parameters needed to mode1 advective-diffisive transport with linear
sorption and first order degradation simulated herein, are iisted in Eq. 1.1 (Chapter 1) and
their characterization and estimates are often associated with laborious procedures and
uncertainties. As noted previously (in Chapters 4 & S), the independent estimates were
made for (linear) sorption and diffusion coefficients fiom separate ancillary tests, and
subsequently these estimates were re-examined in long-term intrinsic degradation tests,
with intention to infer degradation rates for tested chernicals and conditions. Variation of
sorption and diffusion coefficients can be bounded within the reasonable range prescribed
by LFERs (linear free-energy relationships) and many other reponedly valid empiricai
expressions (Lyman et al., 1982; Schwanenbach et al., Mackay et al., 1995). However,
use of biodegradation parameters in large-scale simulations and bio-remediation projects
js regularly associated widi large variations and uncertainties regarding the reaction rates,
particularly due to, still persiaing lack of field data and referent rates for coupled
processes. Perhaps the most difficult to estimate and ail1 the most uncertain are the
parameters related to: biomass size, growth kinetics, type of degraders in a particular
cornrnunity and substrate degradation kinetics. In the "lumped" first-order contaminant
degradation kinetics approach adopted in this study, only one parameter describing the
reaction [Le. rate of contaminant degradation, (A), or half-We, defined as In2hate A] was
considered. [See Eq. 1.1. Approach implies that a non-growing uniformiy distributed
biomass is an "agent" carrying out the (first order) reaction of contaminant removai at
rate A]. The influence of listed factors is not forgotten, yet the approach in which this
influence is neglected is so frequently and successfully used in impact predidion (Rao et
al., 1993) therefore the approach is also used here. Furthermore, the first order
degradation reaction is mathematically convenient and also recognized by the US EPA
Composite Modei for leachate migration with Transformation Products (EPACMTP)
protocol (Aronson & Howard, 1997).
The first order reaction rates, expressed in ternis of the reaction' s characteristic
and popular feature, half-lives, were deduced from simulated laboratory intnnsic
degradation tests as presented in Chapters 4 and 5 . The reported half-lives ranging from
0.75 to 15 days (for ail chemicals) are considered very short, pointing to the fast reaction,
otherwise known to be difision-limited and slow (McMahon & Chapelle, 199 1 ; Scow &
Hutson, 1992; Ventraete & Top, 1999). It is stressed again that these rates refer to the
stage of expenment when the soil consistency changed due to fermentation and gas
release, however the mass transfer limitation imposed by test set-up did not vanish. Other
than these inferred values, no reported degradation rates were found in the literature
searched, which would characterize the degradation of seleaed chemicals under
conditions similar to those relevant to the bamer system (Le. in compacted clay).
Generally, very few documents deal with rates of degradation in the soil (under
environrnentally representative conditions), wit h exception of recently published
remarkable compilation of anaerobic first order rates by Aronson & Howard (1997).
Table A5.4 in Chapter 5 gives the excerpt of these rates for BTEX, DCM, I,2-DCA and
TCE, summarized for different environmentai conditions based on numerous laboratory
and field studies. It is evident fiom the Table A5.4 that al1 sources confinn degradation of
the chlorinated aliphatic chemicals under tested (anaerobic) conditions. The mean in situ
and lower limits for dichlorornethane haKlife of only 108 and 1083 days (0.3 - 3 years)
are pointed as noteworthy Although found to be degradable at relatively fast (i-e. short
half-lives) and site specific rates, BTEX are also indiscriminately reported as non-
degradable in significant number of cases. The rates available from this compilation, (see
Table A5.4) were mainly deduced fiom lab and field
sediments with high hydraulic conductivity, however
metabolic terminal processes, and as such they could
cases dealing with soils and
they are classified based on
be considered for modeling.
Aronson & Howard (1997) have aiso reported half-life of 976 days (- 3 years) as lower
limit for anaerobic degradation of acetic acid in ground water. Chapelle & Lovely (1990),
based on 77 day lab incubation, iderred first-order turnover rate constant of k~ 0.025 y'',
which relates production of CO1 to acetate consumption in clayey sediments. [0.025 y"
translates to half-life of - 28 years, but is not arialy applicable to the rates defined in Eq.
1.1, herein, however this rate, reported as c 0.000068lday is probably taken as the first
order rate by Aronson & Howard (1997)l. In their work, Chapelle & Lovley (1990) also
cautioned that this laboratory inferred rate, (as well as rates reported for more
transmissive materials tested), is fast and likely 2 to 4 orders of magnitude higher than the
field rates estimates inferred fiom geochemical modeling and total carbon balance in
examined sediments. Balba & Nedwell(1982) reported very fast turnover rates for acetate
(0.78 h-'). propionate (0.5 1 K') and butyrate (0.58 K') in cores of intact surface (O - 4 cm)
of marine sediments fiom the Colne Point saltmarsh, üK, yet the rates were practically
zero at 18 - 22 cm depth from the surface.
Surpnsingly, very little information is available regarding degradation rates andfor
half-lives of the contaminants in landfill leachates and in waste generally. Bleiker (1992)
reporied first order rate ki = 0.412 month -' (0.14 y -') for COD for leachate from Brock
West MSW landfill, Ontario. Based on some achial figures (numbers) available corn
Robinson (1995) it was possible to infer half-lives for degradation of COD, TOC and
even the VFAs or interest, compiled for different landfills in the UK. The half-lives [all in
pars] are short and dl refer to the acido- or acetogenic stage of landfill operation, as
follows:
t , ! ~ = 1.2 (TOC); t l ~ = (0.5 - 0.75) (COD); tf,2 = 2.5 (acetate); Ir,? = 3.4
(propionate) and = 1 (butyrate), ail simulated for Stangate site;
t p ~ for COD: 0.6 (Aveley site); 0.24 (Bassett site) and 0.8 (Chape! F m ) .
It is noted that many other data in the Robinson's report clearly indicate fast
removal of bulk organic contamination (usually plotted as COD) from leachate recorded
in large landfills with very high waste input rate. COD drops rapidly within first 2 - 5
years and tends to level off at - 10 -15 % of its initial amount for much longer penod of
time.
Half-lives ranging £tom (10-20) years for DCM, and (25 -50) years for benzene
are recommended for impact modeling by Rowe (1994), based on data on leachate quality
compiled for several MSW landfüls in Southem Ontario.
It appears based on literature sources as weU as experimental study presented in
this work that al1 of the seleaed contarninants will degrade in ground water, contaminated
aquifers, soi1 slumes, leachate and even in compacted clay when the conducive
conditions for microorganisms develop. Yet, the optimism arising fiom still srnall number
of successfÙ1 cases brings also the uncertainties regarding the beginning of degradation
and reliable field-scale rates of degradation, given the complexity and unfavorable
environment al conditions. Acknowledging the recognized implications of scale-up
(S turman et al., 1 994) and ubiquitous heterogeneities in the "subsurface" (Vroblesky &
Chapelle, 1994; Cushey & Rubin, 1997; Haggerty & Gorelick 1998;) the following snidy
only examines the cases of contaminant impact of the hypothetical landfill (characterized
in 5 6.2) emerging from the variation of degradation lag periods and degradation rates.
The approach to simulations relies on distilled information and is generally conservative
assuming relatively fast to moderate rates in the landfill and slow rates in the confining
layers undemeath, with an objective to elucidate the influence of the hypothetical rates on
ground water pollution.
6.4 Results and discussion
The results from the theoretical simulations of the DOC (representing VFAs)
impact on the hypothetical aquifer are shown in Figs. 6.2.1 and 6.2.2. A characteristic
breakthrough pattern (already observed with receptor solution data and curves) with the
maximum concentration in the aquifer reached afker certain time is common for al1 cases
examined. Fig. 6.2.1 demonstrates the effect of degradation lag and different degradation
rates considered in the waste fül only. Fig. 6.2.2 incorporates the influence of DOC
degradation in compacted clay andor naturai till cornbined with DOC degradation in the
waste fill simulated for (a) 5 year lag and (b) 10 year lag.
For the case when only degradation in the waste is considered, three different lag
durations are examined as seen in the Fig. 6.2.1 on the graphs (a), @) and (c). Each (2, 5
and 10 years) of them, is very short relative to the required 260 years of service needed
for this hypotheticai landfill, however they might be representative of the time required
for the commencement of anaerobic fermentation and methanogenesis. Based on data
from many different municipal landfills in the üK compiled by Robinson (1995), COD
drastically drops to - 10 % of its initially ofien very high (- 50 g/L) value within the first
2-4 years. The simulations in the Fig. 6.2.1 indicate that impact of DOC without
degradation of 103 mgL, exceeds the max ailowable lirnit of 2.5 rng/L, however when the
degradation even in the waste fil1 oniy, is taken into account the DOC concentration
becornes significantly reduced. As expected, very short halGlife (0.14 year) originated for
the fast rate taken from Bleiker (1992) produced the lowest impact, as opposed to the
impact resulted for the longest simulated haKlife of 30 years. Generally, the effect of the
different lag periods is not great at slow degradation rates [e.g. compare 16.6 mg'L for a 2
year lag, 19.8 mg/L for 10 year lag both simulated for tin = 30 years, dash-dot-dot lines in
Fig. 6.2.1 (a) and (c)]. A h , the difference between the impact of 1.2 mg/L for 2 year lag
and the impact of 5.2 ng/L for 10 year lag, both simulated with fast rates at tin = 0.14
years only appears more prominent because of high removal relative to the conservative
case of 103 mg/L (Le. without degradation). As noticed previously on a small scale (with
3 and 5 cm and 163 and 271 days, see Fig. 5.14.2. l), it seems that lag period within
realistic time frame (-of 5-10 years) does not significantly dominate the impact. It is
rather a degradation rate that dictates the impact in the aquifer, and for this hypothetical
Iandfill, the objective for DOC of 2.5 mg/L is met if the half Me in the source (waste fill)
is r (0.14-0.5) years pig.6.2. l (a) and 6.2.1 @)], and assurning unfavorable case without
degradation in the soi1 layers. These rates are fast, yet COD t r , ~ = 0.14 year has been
observed in the field (Bleiker, 1992) and TOC t , ~ = 1.2 years inferred fiom Robinson
(1992) exceeds this range. [It is noted that the top text line in the legend on each graph
corresponds to the lowest impact. The subsequent text iines correspond to increasing
impact up to the maximum (case without degradation)].
The situation regardhg the impact in the aquifer improves if the degradation in
the soil is considered as can be appreciated from Fig. 6.2.2. Only the lag periods of 5 and
10 years are considered in combination with variable rates in the waste fil1 (source) and
soi1 layers. Generally, the modeling approach was to simulate degradation with the fast
rates in the waste fi\, yet not faster than with haif-lives ranging from 0.14 to 0.5 years,
since such rates have produced satisfactory impact without the need to account for
degradation in the soil beneath the waste fill. The lowest impact of Cm, = 380 pg/L (thick
dotted lines, graph a in Fig. 6.2.2) is produced with the short half lives of 3 years in both
waste fill and entire one meter compacted clay liner, while the half-life in the rest of the
bamer (till) is set to 1000 years. Such fast rate of degradation in the compacted liner is
unlikely, however, the impact simulated with 30 year-half-life in the compacted clay liner
and 100 year-half-life in the till generates satisfactory impact of Cm, = 1 .O mgL.
An impact of C,, = 1.9 mgR. (not shown) is produced with the shon half-life of 5 years
in the waste fil1 and laboratory-inferred 1 day half-Me at 0.1 m (Le. 10 cm) top of the
CCL only. When degradation was considered in the 0.0 1 ni (i.e. 1 cm) of the 1 m thick
compacted clay liner CCL, the impact was higher and slightly over the pehssible limit
Cm, = 3.1 mg/L. as shown in the Fig. 6.2 2(a) with dash-dot-dot line. The same trend is
captured for 10 year lag and slower rates of degradation in the waste and soil as shown in
the Fig. 6.2.2 (5)
However, other combinations of the short half-lives (Le. fast reaction rates) could
produce low and likely realistic DOC impact in the aquifer. As seen in the graph 6.2.2 (a),
t r n = 5 years in the waste in combination with very conservative (slow) t112 = 100 y in the
soil (entire bamer, CCL+SLCS + till) below yields satisfactory 1.3 mg4 (short-dash).
The same holds for the other combinations shown in the top graph (a) simulated for 5
year lag and bottom graph (3) simulated for 10 year lag in the Fig. 6.2.2. The details are
given in the legend and curves with impact less than the MOE ümit are shown below the
horizontal dot line. Some of the uncertainties associated with very slow degradation rates
in the soil are resolved with the satisfj4ng impact (c 2.5 mga) as seen with the simulated
t 1 ,Z = 500 years [short dash, graph (a)], and 1 O00 years [(thick dot, graphs (a) and fi)].
There is still no evidence that these extremely slow rates could corne into effect, however
degradation might proceed at comparable rates (Chapelle & Lovley, 1990), once the
reduced DOC (VFAs) contamination slowly df i ses from the landfill towards the
aquitard.
6.4.2 DCM
Results for the DCM impact simulated for the hypothetical landfill are show in
Fig. 6.3.1 and 6.3.2. Generally for the conservative case without degradation in the soil
and halElives shorter than 10 yean in the waste fill at initial concentration Co = 3.3 mg/L,
the impact of DCM is below the MOE k t of 12.5 pg/' for the two simulated
degradation lags of 5 years in the graph (a) and 10 yean in the graph (b) as given in the
Fig. 6.3.1.
This seems as quite feasible degradation rate, and one could speculate that such
conducive conditions would develop (such as VFAs fermentation and anaerobic, reduced
environment) in the waste fil1 and consequently facilitate drastic DCM removal at the
source. In case the degradation in the soil is considered for both 5 year- and 10 year- lag
penods and at rates simulated in the Fig. 6.3.2 (a) and (b), DCM concentrations in the
aquifer will be reduced further al1 dropping below the reasonable use policy limit of 12.5
pg L, as show in the legend. In the graph (a) more conservative scenario is presented
with degradation in the compacted clay liner only assurning relatively slow rates, i.e. long
half-lives of 50 and 100 years, while the half-Iives in the waste fil1 ranged from 1 to 10
years.
Sirnilar to the case with DOC simulation, the laboratory inferred half-life of 1 day,
simulated only for the top 0.1 rn of the clay liner with half-life of 5 years in the waste fill
after 10 years lag, resulted in low impact of 1.5 pgiL (not show). The impact was
slightly higher, for the case when the same laboratory inferred hdf-life of 1 day was
simulated for degradation in the top 1 cm of the clay liner (other parameters unchanged)
as aven with dash-dot line with maximum of 4.2 pgLL in the Fig. 6.3 -2 (a). It is believed
that the simulated rates in the waste fill with half-lives ranging fiom 3 to 10 years are
realistic (likely even "over-conservative"), particularly in view of the evidence denved
from the KVL leachate (See Chapter 2 and 3) under anaerobic conditions. This is also in
agreement with the rates compiled by Aronson & Howard (1997) who reported that
DCM, as well as 1,2-DCA and TCE were always degraded relatively fast (0.3 - 3 year
half-life for DCM) in the tested systems and under strongly reducing conditions.
The results for DCM degradation with very long cometabolic half-lives (100 -
500 years) in the entire bamier are shown in Fig. 6.3.2 (3). Although conservative, these
rates are entirely hypotheticai.
6.4.3 Benzene
Results for the benzene impact in the hypothetical aquifer are s h o w in the Figs.
6.4.1 and 6.4.2. The effect of degradation lag and rate of degradation observed for DOC,
is generally valid for benzene as well. Because of uncenainties associated with a
persistence of benzene in contaminated soil, both lag penods and the half-lives simulated
for this midy are taken to be longer, which translates into less favorable impact. As,
remarked by Aronson & Howard (1997), benzene as much as the rest of the TEX, may
exhibit quite fast degradation under anaerobic conditions, however, each of them is also
reported as non-degradable. Effect of the lag penods on benzene degradation is given in
the Fig. 6.4.1. As evident fiom the graph (a), the permksible level of 1.25 pg/L is reached
(and exceeded) for the degradation lag of 10 years and half-life longer than 10 years in
the waste fill and without degradation in the soil. Faster degradation rates (Le. haif-[ives
shorter than 10 years) wiil produce satisfactory low impact in the aquifer. If the longer
(and less favorable) degradation lag of 50 years is simulated, at half-life of 50 years,
impact reaches 5.1 pgZ in the aquifer, as seen in the Fig. 6.4.l(b). For the more
conservative 50 year lag, benzene concentration will be at allowable lirnit even at fast rate
with haif life of 0.5 years (" 1 -275" pgA, not shown)]. Similar to the observation made for
DOC, if degradation is not considered, the Cm, = 22. pg4 in the aquifer, highly exceeds
the allowable iirnit. Aiso, the difference in impact at the two simulated lags (at 10 and at
50 year half-life between 3.3 and 5.1 pga) is marginal and probably very difficult to
distinguish. The simulated lag periods are considered conservative and appropriate when
uncertainties of benzene biodegradation are recded (Aronson & Howard, 1997), thus it
might be speculated that this hypothetical landfill would not be safe for benzene
emissions if the half-life is 25 (as recomrnended by O. Reg. 232/98) in the waste fil1 only.
When degradation in the soi1 layers (CCL andior till) is taken into account, the
simulated impact decreases. It would suffice to have very slow rates of degradation, with
half-lives ranging from 25 - 50 years in waste (and CCL) and 1000 years in the till
deposit to reduce the impact in aquifer to the limit of 1.25 pg/L, as s h o w in the Fig.
6.4.2(a), for Cm, = 0.96 and 1.1 pg/L (long-dash and dash-dot lines) and @) with Cm, =
1.7 pg L (long-dash line) for the two conservative lag periods, 10 and 50 years,
respectively. (Half-life of 1 day simulated for the top 1 cm of CCL interface and 100 year
half-life in the waste fil1 resulted in 0.44 and 1.3 pg/L respectively for the 10-year and 50-
year lag penods.)
These simulations indicate that even very slow degradation could potentially
result in acceptable impact, however these rates are hypothetical and as remarked for
DCM, there is still a question as to whether degradation of benzene is possible in naturai
confining deposits, particularly if some other more readily degradable organics such as
VFAs are absent.
6 4 4 Xylenes
Results of the simulation of xylenes impact on the hypothetical aquifer are given
in the Fig. 6.5. As evident h m the graph, even under the worst case considered (i-e.
without degradation), Cm, = 10.9 pgL, and much less than Cal= 75 pg/L (MOE, 1998,
O. Reg. 232/98), and as such xylenes do not pose a threat to ground water quality if
disposed in the hypotheticd landfi1 at Co = 1 mg/..
6.5 General remarks on transport and degradation simulations
Based on the results from the impact simulations presented in this Chapter, it may
be hypothesized that anaerobic degradation provides sufficient degree of attenuation to
the transport of DOC (VFAs), DCM and benzene. Aithough the desired level of
attenuation could be accomplished with very slow degradation, the uncertainty associated
with the environmental factors seems to dominate the prospects of degradation in the clay
liner and natural confining layer. It is possible that the degradation of selected
contaminants could proceed at simulated fast and rnoderate rates, but mainiy in the waste
fil1 (landfill). DOC (Le. WAs) would degrade "readily" afler 2 to 10 year lag at 2 to 5
year half-life (given the sub-optimal conditions in landfill environment), and its presence
and unhindered removal at the source might be crucial for initiation ancilor successful
degradation of other less wanted contaminants. Considenng the unlirnited amounts of
WAs, it could be hypothesized that DCM and benzene would also be cometabolized in
the waste at favorable fast rate, e.g. (5 - 10) and (10 - 25) years respectively at 10 year
lag. This conservative scenario already provides significant amount of irreversible
attenuation of the impact of these pollutants in ground water, however given the stnngent
regulations and other adverse conditions encountered in the real hydro-geological setting,
it might be necessary to engage all resources of the natural eco-system in degradation, at
least to reasonable extent. It is shown that the very slow rates of degradation (having half-
lives fiom 100 to 1000 years) in the compacted clay liner tested for this hypothetical case
could significantly reduce the impact of contaminants released from municipal solid
waste landfills. Recogninng the limitations imposed on the reaction in the field arising
fiom many unknown inhibiting factors from "un-optimized" leachate medium as well as
from confhing stress due to the (nevertheless fermenting) waste fil1 it is not expected that
the rates of degradation in the compacted clay liner could ever become as fast as inferred
fiom intrinsic degradation experiments presented in Chapten 4 and 5 . However,
considering that a one-meter clay liner is placed between the very trasmissive drainage
layers (Le. primas, and secondary leachate collection systems) it is possible that at some
(yet realistic) time when readily degradable organic pollutants diffuse in, the redox
conditions in the soi1 will be gradually changing at the onset of biodegradation. In such
setting there might not be obvious expansion and excessive fluidizing of the compacted
liner, yet if even srnall amounts of the fermentation gases are relieved into the
transmissive layers the degradation could possibly continue at very slow but
environrnentally significant rates. As pointed out earlier in the Chapter 5 there is not
sufficient information to determine the field degradation rates, thus design approach still
has to remain conservative although the degradation of organic pollutants in the
compacted clay, tested in this study, might be possible. While degradation of degradable
pollutants such as low levels of diffusing VFAs is likely, pmicular caution should
however, be exercised if degradation of micro-pollutants, such as DCM and benzene, is
considered in any layers with low hydraulic conductivity, because such degradation might
be dictated by the presence of other (more readily) degradable organic chernicals or
intermediates of anaerobic fermentation.
6.6 References
Aronson D and PH Howard, 1997, Anaerobic biodegradation of organic chernicals in
groundwater: A sumrnary of field and laboratory studies, Fiai Report,
Environmental Science Center, Syracuse Research Corporation, North Syracuse,
NY, htp://esc.symes. codAnaerobicRpt. hm
Balba, MT and DB Nedwell, 1982, Microbial metabolism of acetate, propionate and
butyrate in anoxic sediments from the Colne Point saltmarsh, Essex, U. K.,
Journal of General Microbiology, 128, p 141 5- 1422
Bleiker, DE, 1992, Landfill performance: Leachate quality prediction and settlement
implications, Master's. Thesis, Department of Civil Engineering, University of
Waterloo, Waterloo, Ont., Canada
Chapelle, FE and DR Lovley, 1990, Rates of microbial metabolism in deep coastal plain
aquifers, Applied and Environmental Microbiology, 56 96), p 1 856 - 1874
Cushey, MA and Y Rubin, 1997, Field-scaie transport on nonpolar organic solutes in 3-
D heterogeneous aquifers, Environmentai Science & Technology, 3 1 (5 ) , p 1259-
1268; also see Correspondence, ES &T, 1998, 32 (17) p 2654-2656
Fleming, Ut, RK Rowe and DR Cullimore, 1999, Field observations of clogging in a
landfill leachate collection system, Canadian Geotechnical Journal, 36 (4), p 686-
707
Haggerty, R and SM Gorelick, 1998, Modeling mass transfer processes in soi1 columns
with pore-scale heterogeneity, Soi1 Science Society of Amenca, Journal, 62, p 62-
74
Lyman WJ, WF Reehl and DH Rosenblatt, 1982, Handbook of chernical property
estimation methods, McGraw-Hi11 Book Company
McMahon, PB and F'Ei Chapelle, 1991, Microbial production of organic acids in
aquitard sediments and its role in aquifer geochemistry, Nature, 349, p 233-235
National Pnmary Drinking Water Regulations (NPDWRs), USEPA Office of Ground
Water and Drinking Water, Current dnnking water standards,
www. epa.gov/safewater/mcI. htmI
Ontario Drinking Water Objectives (revised), 1994, Ontario Ministty of the
Environment, 8 Queen's Printer for Ontario, 1999
Ontario Regulation 232198, 1998, made under the Environmental Protection Act,
Extract from the Ontario Gazette, vol 13 1-22, Management Board Sectretariat, O
Queen's Printer for Ontario, 1998
h o , PSC, CA Bellin and ML Brusseau, 1993, Coupling biodegradation of organic
chemicals to sorption and transport in soils and aquifers: Paradigms and
paradoxes, in DM Linn (ed.) Sorption and degradation of pesticides and organic
chemicals in soils, p 1-27, SSSA Special publication No 32, Soil Science Society
of Arnerica, hc . and Amencan Society of Agronomy, Inc. Madison WI
Robinson, 8, 1995, A review of the composition of leachates from domestic wastes in
landfill sites, Report prepared for the üK Department of the Environment, Under
Contract Number PECD 711 O/23 8
Rowe, RK, RM Quigley and J R Booker, 1995, Clayey bamer systems for waste
disposa1 facilities, E & M Spon., An Imprint of Chapman & Hall
Row e, RK, 1995, Leachate characteriration report, prepared in CO-operation wit h Golder
Associates Ltd., Fenco MacLaren Inc., M M Dillon Ltd., and Groundwater
Research Ltd. for Intenm Waste Authority Ltd.
Schwaizenbach, RP, PM Gschwend and DM Imboden, 1995, Environmentai organic
chemistrv, A Wiley-Interscience Publication, John Wiley & Sons, Inc.
Scow, KM and J Hutson, 1992, Effect of diffusion and sorption on the kinetics of
biodegradation: theoretical considerations, Soil Sci. Soc. Am. J., 56, p 119-127
Sturman, PJ, RR Sharp, JB DeBar, PS Stewart, AB Cunningham and JH Wolfram,
1994, Scale-up implicatins of respirornetrically determined microbial kinetic parameters,
in Hinchee et al., (eds) Applied biotechnology for site remediation, 301-304,
Verstaete, W and E Top, 1999, Soil clean-up: lessons to remember, International
Biodeterioration and Biodegradation, 43, p 147- 153
Vroblesky DA and FEI Chapelle, 1994, Temporal and spatial changes of terminai
electron-accepting processes in a petroleum hydrocarbon-contaminated aquifer
and the significance for contaminant degradation, Water Resources Research, 30
(9, p 1561-1570
Table 6.1 Layer data for hypothetical landfil1
Compactcd clay liner
HCWSLCS
Naturil attenuation layer
froiii Table 4.2 and Table 5.
DOC 1 DCM 1 Bcnzcnc DOC 1 DCM 1 Benzercnc Xylencs
full rnixing assumed
2.5 - 5.0
(Chiipiers 4 and 5);
Table 6.2 Change of Darcy velocities in the layen during operation of the hypothetical landfiil
Remark
-- - - -- - . . .
1 Nornial operation of PLCS. leachatc hcad 0.3 ni ,
Ciraduil failure of PLCS, leachate head nses (0.3 - l O) in. vclociticsl (linearly) increase and change
direction: omration of SLCS mduallv starts SLCS works at maximum in order to accoiiinioâiite the wak iin~act: niound and velocities at mü?iimiim 1 Peak iiiipact succcssfully attenuated, opration of
Darcy vclocities: in tlic coiiipacied clüy liner: vertical va.; in the ti11 vcrtiçi~l v,,,~ in ttic SLCS v s l . ~ ~ Iiorizoiitd (rcinoval); in the aquifcr v , ~ horizontal
' 4 vertical dowiii~iirds, '? vertical upwards, + horizontal rsiiioval out of SLCS; uelocities wlculated based on the continuity of Darcy flow taking into account the clwnge in Icacliatc tiead (0.3 -lOj ni. landfil1 gmiiictry and k of 10'" (iiilsl in boit1 coiiipacied and natural soi1 layen (for details. see Fig. 6.1) w
. t l / t=O.14~ , &=1.2& \ 1 -- tin = 1.2 Y. = 2.0 mg'L \ \ ---. t i n = 3 y. &= 3.3 mgiL
-. - $, = 5 y, C,= 4.7 mg/L
- - t,,= 30 y, Cm= 16.6 mgi t
2.5 mwL MOE limit - no &gmd&onC, = 103 mgiL
'(a) .O
0.1 C 1 O 200 400 600 800
Time Cyears]
. . . t l ~ a 0.14 y & = 5.2 m g L
- - t i n - 1.2 y, kX = 5.9 r n ~ L --- t1,2 = 5 y. & = 8.3 mg&
- - 1, = 30 y. C, = 19.8 & 2.5 m f l MOE limit
- no degradatioa = 103 mg'L
Time bears]
Fig. 6.2.1 Impact of DOC in hypotheticd aquifer: variation oflag penod: DOC degradation considered ody in the waste fil: (a) 2 year lag (b) 5 year lag (c) 10 year lag (Co = 2.4 gL; D = 5 x 10-'O m% )
l
. t,, = 3y ( wastc & CCL); t,R = 1000 y (tiI1) C, = 0.38 mg&
-- Ga 5 Y ( m X GR = 30 Y (CCL); t,, = 1 Oû y (till); C, = 1.0 m a
- - t,, = 5 y (waslr), 100 y (CCL & till) C,= 1.3 m g L
--- . t in = 1 y ( W h tin = 500 y (CCL gt till)
C, = 2.4 mgiL
- - ly(-),Gn= 1 d ( 1 anCCL), C, = 3.1 mg&
2.5 mwL MOE limit - no âepddol r . C, = IO3 mgL
I
100 ? i (b) . . . $ n ' 0 . 1 4 y ( ~ - C C L )
q n = 1OOû y(til1) C,= 138 pgL -- tins 3 y (wask A 1 cm CCL),
ftn= 100 y ('cd of CCL & till); C, = 1.5 nt@ - - ttir= 3 y (wastt + 0.1 m CCLX
$,= 100 y (rcst of ma); Cm= 1.7 mgL --- t, ,=3 y(wasleAO.lmCCL);
t,, = 250 y (m of banicr); C,= 3.4 mg&
- - ttn= ly(wask),$n = 1 d ( 1 cmCCL), C,= 5.4 mg%
2.5 mg/L MOE lUnit - no degndrtion C, = 103 mgrL
I
Fig. 6.2.2 Impact of DOC in the hypothetical aquifer: Variation of half-lives in the waste fiii and soi1 layers: (a) 5 year lag @) 10 year lag
-10 2 ( Co = 2.4g/L7 D = 5 x 10 m /s)
Time h e m ]
Fig. 6.3.1 Impact of DCM in the hypothetical aquifer: variation oflag period DCM degradation considered in the waste fil1 ody a) 5 year lag b) 10 year lag
-10 2 ( D = 5.0 x 10 m Is, K,=o.os cm3fg )
Time [years]
. . . t l n = l ~ ( ~ X t l n = 5 y ( C C L )
(2,- = 0.5 p#L
- - t,, = 1 y (u*), t in = 50 y (CCL)
& = 2.7 Cign
-- t i n = 5 y ( < n a c ) . t l R = I d ( l m C C L )
& = 4.2 c18/L --- t i n = 1 O y ( w ~ ) , t IR =100y(CCL)
&=8.1 - - t l n = 15 ~ ( w W ) . t1/2 = 1000 y(CCL)
ç- = 12.0 Cign
12.5 pg/L MOE limit - wdegmhan & = 1 2 9 ~ 1 & &
O 250 500 750
Tirne [years]
. . . t i n = 10y(unne).t ln= lOoy(CCL&till)
+= 1.8
-- t l n =5y(YMC) . t lR=500~ccL&t i i l )
& = 5.6 pgL
-- t ln=Sy(wrac) . t l iL= ld(1cmCCL)
= 6.5 )rg/L
--- t = IO y (wrstt). t = 1000 y (CCL & till) i n i n
c,-=9.i pgL
i 2.5 p#L MOE limit - n0-0n,&=129 WL
Fig. 6.3 -2 Impact of DCM in the hypothetical aquifer: Variation of half-lives in the waste fil1 and soil layers a) 5 year lag b) 10 year lag
10 2 ( D = 5.0 x 10- m /s, K, = 0.05 cm3fg )
!
_ - - -
------_
$ , = 5 y.C,= I.4pg/L I -- t,,= IOy.C,= 1 . 6 w L i l
î,, = 2 5 y, C, = 2.6 pgL (
--- t,, = S0y.C- = 5.1 pgL 1 I - - fi, = 100 y, C, = 9.2 pgd 1
1.25 lrglL MOE limit - no &gradation. Cm = 22 pg/I. l
Fig. 6.4.1 Impact of BENZENE in the hypothetical aquifer: variation oflag penod Benzene degradation considered in the waste only (a) 1 O year lag (b) 50 year lag ( Co= 1 .O mgL, D = 5.0 x 10*1° mZ/s, K, = 0.5 cm31g )
Time [years]
cc---__ ---- - z=, - - -- . . . . - * o . . . .
$, = 100 y (m), t,,= ld (1 CUI CCL) - a . - C, = 0.44 Crgn
-- t,, = 25 y (waste), t,n = 1Oûû y (CCL & tiil) C, = 0.96 pgL
- - = 50 Y (w"ste + CCLh t,, = 1 O00 y (till) c, = 1.1 CLgn
--- $a = 100 Y ( W h Gn = 1000 y (CCL & till) c , = 4.3 Cign
- - t,, = IOû y (wastc), $, = 5000 y ( CCL & till)
Fig. 6.4.2 Impact of BENZENE in the hypothetical aquifer: Variation of of haIfilives in waste fill and soi1 layers (a) 10 year lag (b) 50 year lag
10 2 (Co= 1.0mga, D = 5 . 0 x 10- m /s, K,= 0.5 cm3@)
0 soo Io00 C, = 6.3 CLgn
1.25 pgl. MOE limit T h e [years] / - n o W r n C , = 2 2 p g L
CHAPTER 7 CONCLUSIONS AND RECOMMENDATIONS
7.1 Summary and Conclusions
Tests were conducted to examine the potential for biodegradation of organic
contaminants in laborator- compacted soii. Three volatile fatty acids (Wh, acetate,
propionate and butyrate) as representative of bulk organic contamination and eight
volatile organic compounds [VOCs, B E X : (benzene, toluene, ethyl-benzene and
xylenes) and DCM (dichloromethane), 1,2-DCA (1,2-dichloroethane) and TCE
(tnchloroethylene)] representing organic micro-pollutants in synthetic leachate were
tested as they diffise fkom reduced synthetic leachate through sixteen compacted clay
plugs during testing periods of 163 and 271 days. Results of this study demonstrate that
volatile fatty acids, DCM, 42-DCA, TCE, benzene and toluene are subjected to intnnsic
degradation mediated by indigenous rnicroorganisms in the soil. Removal of ethyl-
benzene and xylenes was not observed during the course of this expenment.
Each of the tested chernicals exhibited the difisive breakthrough, considered to
be unafEected by degradation at least for &st 50 days of testing. M e r this period, the
growth of selected group of soil rnicroorganisms, such as HAB (heterotrophic aerobic
bacteria) and SRB (sulfate reducing bacteria) was evident and approached a maximum of
(2 - 8) x 10' cfu/g and (3 - I l ) x IO' cfug, respectively (Le. on average 1000- and 100-
fold higher than initial count in untreated soil). As time elapsed the sign of intense
microbial activity became more evident, particularly in the upper centimeter of the soi1
plug in contact with source of organic substrate-contaminants. This upper interface was
gradually fluidized by gases generated from VFAs fermentation, its compacted stmcture
loosened and favorable conditions arose for enhanced mass transfer and microbial
growth. It is considered that manifest biodegradation of al1 contaminants was largely
localized in this reduced and "fermenting" soil layer. However, regardless of eariy
observed changes in soi1 appearance, sigdcant increase in bacterial count and ATP
content, as well as confhned methanogenesis, the consumption of VFAs was smaii and
degradation did not become measurable until &er a somewhat long lag of 140-180 days.
This is attnbuted to high concentration of VFAs (2.4 gL as DOC, total VFAs expressed
as dissolved organic carbon) introduced to this initiaily oligotrophic micro-environment,
as well as severe limitation to exchange the incoming carbon and nutrients through
cornpacted clayey soil with small pore size. Once the microorganisms acclimated the
fermentation advanced and the acids were removed from the soii at very fast rates,
resulting in short half-lives of 0.75 - 5 days as simulated for DOC (total VFAs expressed
as dissolved organic carbon).
Degradation of the liaed VOCs becarne obvious earlier than VFAs and after 55 to
90 days, DCM, 1,2-DCA, TCE, benzene and toluene were cometabolized very fast, with
half-lives ranging fiom 1 to 15 days, simulated oniy in the top of the bio-active clay layer.
Based on the observed resuits it cm be concluded that soil indigenous microbial
population could initiate and carry out biodegradation of VFAs, (taken as buik organic
contamination), as well as the most of the VOC teaed (i.e. three chlorinated aliphatics,
benzene and toluene) without any man-induced intervention. The reaction, modeled as
first order with very short half-lives, appears to be most dominant at the top contact
interface between soil and source of carbon and nutrients Frorn the synthetic leachate.
These experiments clearly show that remolded and cornpacted soil is a source of
microorganisms that are capable of irreversible and terminal degradation of high levels of
volatile fatty acids as well as some priority pollutants.
In separate short-terni tests it was inferred that al1 of the tested chemicals difise
through the same laboratory compacted clayey soil at rates with diffusion coefficients of
D = (2.5 - 5.0) x IO-'* m2/s and negligible to low linear sorption with Kd = (O - 3) crn3lg
(more exactly: zero Kd for VFAs; 0.05 for DCM; 0.25 for 1.2-DCA, 1.33 for TCE; 0.5 for
benzene; 1.5 for toluene; 2.35 for ethyl-benzene and 2.5 - 3 for xylenes) for this soil
which had a organic carbon fiaction f, of (0.29 - 0.45) %. ïhese coefficients have been
successfùlly reassessed in long terni intrinsic degradation tests
Based on the findings from the laboratory tests, which simulated very adverse
conditions for degradation, it is hypothesized that an environment conducive to
degradation might develop in actual waste disposal facilities as well. It is recognized that
field conditions in the real landfill and field scale compacted clay liner will be more
adverse than simulated in the tests due to the heterogeneities of the large system and
limited means of controllhg the inputs.
exerts to the compacted clay liner will
Presence of the waste fiil and confining stress it
impose additional limitation to the already slow
and non-optimized degradation. However considering d e and perpetual confinement, the
prospects for field sa le intrinsic degradation of organic chernicals tested seem to be
realistic and effort should be made to explore and define the factors that could facilitate
such degradation. Since this thesis represents oniy the first step in attempting to
understand the processes, it is difficult to speculate on the outcome of degradation of
selected contaminants under different conditions bearing in mind the complexity of a
land fil1 and rnicrobial ecology of its underground surroundings.
It is hypothesized that the presence of volatile fatty acids is crucial factor and
driving force of intnnsic degradation of less reactive and recalcitrant micro-contaminants.
Based on the observations gathered from long-term intrinsic degradation tests with soil
and constant supply of the VFAs it is believed that the acids as readily degradable and
preferential substrate are needed to initiate and maintain the conditions conducive for
cometabolism of other organic compounds. The long-term diaision test performed with
VOCs only conflrmed quite stable and high levels of al1 chernicals in the soil pore water
as well as persisting oxidized conditions after 206 days of testing, which could be
attributed to the absence of the fatty acids. Volatile organic compounds alone present at
low mgil could not initiate growth and activity of indigenous rnicroorganism in the soil,
thus remaining unconsumed in the system subjected to diffision slightly retarded due to
sorption.
Intrinsic degradation of VFAs and particularly VOCs in compaaed clay liners and
confining deposits is expected to be very slow, not only because of long periods of
acclimation required for conducive conditions to arise, but also because of recognized
severe mass transfer Limitations imposed to the reaction. Such biodegradation is site
specific and could be affected by hydrogeologic setting, in particular by the proximity of
discharge-recharge zones and availability of naniral electron acceptors and as such it
should not be considered in isolation f?om the landfili. When judging whether
degradation in low hydraulic conduaivity layen below the municipal solid waste landfi11
is feasible many factors regarding a particular design proposal such as initial contaminant
mass, landfill size and infiuence of other engineered components have to be taken into
account. In theory and based on the experimentation reported herein, the long acchation
periods and slow degradation rates could potentiaily be very effkctive when a perpetud
protection against the specified emissions is required, and sigrilficant portion of
contamination could be, removed even 60m "containment only" facilities such as
compacted clay liners placeû at the bottom of municipal solid waste landfills. However,
there is not sufficient documented evidence ffom field liners and landfiil facilities thus it
rernains for design engineers to proceed with conservative approach and either not take
degradation of organic pollutants into account at ail or consider only long degradation lag
periods in conjunction with slow degradation rates in confining layers, as ilhstrated in
this thesis.
7.2 Recornmendations for future work
The following recommendations for fùture work are developed based on the
findings of the thesis and in recognition of its lirnited scope:
investigate fate of volatile fatty acids alone under confining stress applied in the
large s a l e experiments with the primary objective of elucidating the effeas of
concentration and inhibition arising from release and accumulation of
intermediates and their interaction with soil. This would require a careful
monitoring of released intermediates and characterization of changes of soil
structure, as weU as of the indigenous soil microorganisms engaged in the
particular steps of degradation.
following a thorough assessrnent of VFA degradation, examine the potentid for
degradation of other organic contaminants commonly found in landfili leachate,
both in the simplified and real environmental media and settings;
investigate the effect on Uitrinsic degradation of a "sacrificiai" soi1 layer with a
hydraulic conductivity and pore size much higher than compacted clay liner, with
the objective of assessing whether such a layer could potentially be used to initiate
degradation and maintain it at fast rates before the contaminant reached the liner,
investigate the potential of intrinsic degradation for more complex settings
including sophisticated engineered dfision-advection barriers, cornposed of
synthetic liners, pre-fabricated earth liners with bentonite and organo-adsorbing
additives;
examine the degradation in real municipal solid waste landfills and attempt to
assess the effect of the uncertainties arising from heterogeneities and scaling-up
associated with coupled processes.
Abiotic
Aerobic
Aliphatic
Anaerobic
Aquifer
komatic
Archaeobactena
ATP
Autotroph
not living or not biotic
living, active, or occurring only in the presence of oxygen
of or pertaining to an organic compound that has an open chah structure (ais0 acyclic)
designating a compound denved from a saturated cyclic hydrocarbon
in the absence of oxygen; not requinng presence of or not capable of using molecular oxygen for growth
water-bearing stratum of permeable rock, sand or grave1
of or pertaining to a carbocyclic organic compound that contains the benzene nucleus
a group of bacteria compnsing the third pnmary domain (kingdom) Archaea, a class of unusual bactena that phylogeneticaily are neither prokaryotes nor eukaryotes; they have some characteristics of prokaryotes (such as absence of nucleus and ce11 organelles) and some characteristics of eukaryotes (such as specific initiation of protein synthesis) and some characteristics that are unique to them (such as composition of ce11 wall and the type of membrane lipids); archaeobactena include thermoacidophiles, extreme halophiles and methanogens and may represent some of the earliest forms of Living cells
adenosine-(5')-triphosphate ; the high energy compound that functions in many biochemical systems (if hydrolysed to either - monophosphate or diphosphate the reaction is accompanied by the release of large amount of energy which is used to drive a variety of metabolic reactions)
ce1 or an organism that uses carbon dionde as its sole carbon source and that synthesizes al1 of its carbon containing molecules f?om carbon dioxide and other small inorganic molecules
Basal medium a medium that supports the growth of a range of nutritionally undemanding chemo-organotrophs
' Ternis and explanations taken h m Stenesh, 1989. Diction- of biochemistry and molecular biology. 2* edi tion
Carbonyl group the group CHO, occurring in aldehydes and ketones
Carboxyl group the Eree radical -COOH of an organic acid
Carboxylation enzyme catalyzed reaction in living system by which a molecule of carbon dioxide is introduced into an organic molecule
Carcinogenic describing a chernical substance or type of radiation that can cause cancer in exposed animals or humans
Catabolisrn biological breakdown (decomposition) of materials into their simpler components performed mainly by bacteria and fun@ (opposite of anabolism)
C O-enzyme an organic compound that fùnctions as a CO-factor of an enzyme, i.e. electron carrier which serves as a donor and acceptor of either electrons or electrons and protons in an electron transport system (e.g. respiration)
Colony
Commensalism
a group of contiguous cells that grow in or upon a solid medium and are denved from a single ce11
a close and permanent association between two populations of organisms in which one population benefits without damaging or benefi ting the other
Cometabolism biotransfonnation of a compound by a rnicroorganisrn that is incapable of using the compound as a source of energy or growth
Datum (pl. data) an experimental finding; a fact; a measurement
Dehalogenation removd of halogen (Cl, F, Br) atom from a molecule
Diffision limited (controlled) reaction-descriptive of a reaction in which the rate of reaction depends solely on the frequency of molecular encounters as a result of diffision
Disproportionation (or dismutation): a chernical reaction in which a single compound serves as both an oxidiring agent and as a reducing agent and gives rise to two or more compounds by gain or loss of electrons (e-g. conversion of two molecules of pymvate plus a molecule of water to one molecule each of lactate, acetate and carbon dioxide, see fermentation)
Dehydrogenation the removal of hydrogen from organic compound
DOC
Electron acceptor
Electron donor
Electronegative
Endergonic
Enzyme induction
Eubacteria
Eukaryotes
Eutrop hic
dissolved organic carbon, a measure of organic compounds that are dissolved Ui water, in the analytical test for DOC, a water sample is filtered, (to remove particulate matter, i.e. separate total fiom dissolved O rganic carbon) acidified (t O remove inorganic salt s of carbonates and bicarbonates) and then chemically converted to CO2, which is then measured to compute the amount of organic material dissolved in the water
s m d inorganic or organic compound that is reduced in a metabolic redox reaction in order to complete an electron transport chain; an oxidant; an oxidizing agent
small inorganic or organic compound that is oxidized to initiate an electron transport chain; electrons are derived from it in a metabolic redox reaction; a reductant; a reducing agent
describing the tendency of an atorn or a group of atoms to gain elearons; having negative charge; having an excess of electrons
reactions (process) that must consume (absorb) fiee energy in order to go to completion, the uphill reaction; has a positive fiee energy change, AG > O
a protein molecule produced by living cells that functions as a cataiyst of biochemicai reactions; number and type of reactions determined by specificity of enzymes; classified into six main groups: oxido-reductases, transferases, hyrolases, lyases, isomerases and ligases
a process whereby an inducible enzyme is synthesized in response to an inducer
prokaryotes that are distinct from archaeobacteria; the term used originaily to denote "tnie" baaeria as opposed to other rnicroorganisms; now used to designate dl bacteria other than archaeobacteria
narne refers to members of the dornain Eucarya, but is not used in any formai taxonornic system; a higher organism (unicellular or multicelluiar) that maintains their genome within a defined nucleus
describing system e ~ c h e d with excessive amount of nutnents (NO{ and PO:> where rnicroorganisms can grow to produce a large biomass; also refers to body of water (lake) which is deficient in oxygen
Exergonic reaction that has a net release of free energy, a downhill reaction; has a negative kee energy change (AG < 0)
Exponential decay the mode of radioactive decay that can be descnbed by the equation N = N&'L<, where N is a number of radioactive atoms present at time t, No is the number or radioactive atoms onginally present, e is the base on natural logarithm and 2 is the decay constant; the same equation also describes pseudo-first order chernical reaction
Exponential growth the growth of cells in which the number of cells (or the ce11 mass) increases exponentially and the growth at any time is proportional to number of cells (or ceii mass) present; the exponential growth rate constant equals the reciprocal of the doubling time, expressed as number of generations per hour
Fermentation metabolic reaction (process) in which organic chemicals are disproportionated and one part is used as electron donor and other part as electron acceptor, comrnon with highly oxidized (energy rich) organic substrates in the absence of oxygen; contrary to respiration(s) which is sustained in presence of external electron acceptor
Free energy that component of a total energy of a system that can do work under conditions of constant temperature and pressure; known as Gibbsfree energy (G) and expressed by thermodynarnic fûnction G = H - TS, where H is enthalpy, T is absolute temperature and S is the entropy
Free energy change AG, Gibbs free energy change, amount of useful energy liberated or taken up dunng a reaction (difference between the free energy of formation of products and reactants); at standard conditions (reactants at I .O mol, pH 7 and I atm pressure) denoted AG'
Glutathione widely dispersed tripeptide that serves as a CO-enzyme and is also thought to function as an antioxidant in protecting the sulfhydryl groups or enzymes and other proteins; glutathione-S-pansferme (mercapturic acid) refers to a large group of transferase-bound substances formed by the detoxification of xenobiotics
t , ~ , the time required for one-half of either the mass or the number of atoms of a radioactive substance to undergo radioactive decay (radioactive trrJ; the time requûed for one-half of the mass of a substance to be either metabolized or excreted by an organism (biologicai il); the tirne required for one-half of the mass of a reactant to undergo chernical reaction, for a fist order reaction
Half-saturation
Heterotroph
Hydrop hilic
Hydrophobic
Hydroxylation
Homoacet ogenesis
Kinetics
Methanogen
Methylation
Methylotroph
tm = Zn 2/k = 0.693fi, where tlo is the half-life and k is the reaction constant
(Mchaelis-Menten) constant Km, kinetic constant numerically equal to the substrate concentration that yields one-half of the maximum velocity of the reaction at the sahirating concentrations
organisrn that derives its carbon nutrition and energy from breaking down variety of different and complex organic materials and that synthesïzes aii of its carbon-containing bio-molecules from these compounds and fkom srnall inorganic molecules, (opposite frorn autonophl
the reaction of a substance with water in which the elements of water CH, OH) are separated: (a) the breakage of a molecule into two or more smaller fragments by the cleavage of one or more covalent bonds of acid denvatives; the elements of water are incorporated at each cleavage point such that one of the produas combines with the H of the water while the other product combines with the hydroxyl group of water; (b) the formation of the undissociated form of a weak electrolyte through the reaction of the ion of that electrolyte with either proton or hydroxyl ions
polar; also descriptive of the tendency of a group of atoms or of a surface to become either wetted or soivated by water
nonpolar; also descriptive of the tendency of a group of atoms or of a sunace to resist becoming either wetted or solvated by water
the introduction of a hydroxyl group (OH) into a organic compound
generation of acetate f?om autotrophic reduction of carbon dioxide with hydrogen
the science that deals with the rate behavior of physical and chernical systems; kinetics coefficient (rate) is a rate constant that depends on the concentration of either reactant or a product
a methane-producing achaeon
introduction of methyl group -CH3 into organic compound (opposite fiom demethyIation)
an organisrn that can utilize as its sole carbon source either one carbon compounds (such as m e t h e or methcmol) or carbon
Mineralization
Mixotrop h
NAD
Oligotrop hic
compounds that contain no carbon to carbon bonds (dimethyl- ether)
breakdown of organic materials into horganic materials brought about by microorganisms
bacterium that uses simultaneously inorganic and organic energy sources and/ or inorganic and organic carbon sources
nicotine adenine dinucleotide (reduced NADH + K, oxidized NAD'), a CO-enzyme, form of the vitamin nicotinic acid for pyridine-linked dehydrogenases active in ATP synthesis
descnbing system with low content of nutrients, where the microorganisms are not able to form a significant biomass; also refers to lakes and other ecosystems, which are low in nutnents and saturated with dissolved oxygen
Oxidation/reduction coupled reaction of electron transfer: in strict chernical sense, oxidaiion is the Ioss ojelechons by atoms or compounds, reduction is the gain of electrons to atoms or molecules; in broader sense: oxidation is addition of oxygen (in chernical reaction) and removal of hydrogens (dehydrogemtion, in biochemical reaction); by analogy reduction, i.e. addition of electrons is frequently accompanied with addition of hydrogen atoms (hyukogetiation). Oxidations are aiways accompanied with simultaneous reductions.
Phosp horyiation introduction of phosphate group into an organic compound through the formation of esther bond between the compound and phosphoric acid; step in A ï F synthesis; storing energy in the phosphate (high-energy) bond (breaking of the phosphate bond releases energy) substrate-level-phosphodation a process in w hic h high-energy bond, which traps some of the total free energy released during oxidation, is fonned on the substrate which is being oxidized, characteristic to fermentation and contnbutes only a small portion of the total energy conserved in high-energy bond; example: (synthesis of ATP in giycolysis );(this energy yielding process is not linked to respiration, i.e. an electron transport system)
Prokaryotes name not used in any forma1 taxonomie system, but previously used interchangeably with "bacterid', more recently used to denote one of the three domains of Me, Prokqa; a simple unicellular organism (such as bacterium) that lacks a discrete nucleus surrounded by a nucleic membrane and that maintains its genome dispersed throughout cytoplasm
Protonate to add protons to a group of atoms or to a compound (opposite from deprotomte)
Reducing equivalents a m u r e of reducing power equal to either one electron or one hydrogen atom
Respiration energy yielding process of coupling oxidation of organic chemicais with reduction of extemal (terminal) electron acceptors; aerobic if electron acceptor is molecular oxygen, anaerobic if electron acceptors are nitrate (NO3), femc ion ( ~ e ~ 3 , haiogenated organics, sulfate (~04~3, carbon dioxide (CO2) or bicarbonate (HCO33
Sediments soi1 particles sand, clay or other substances that settle to the bottom of a body of water, also geological strata (gravels, muds, clays and shales, rocks) compnsing of deposits of particulate matenals in some area by mechanical means (glaciation, consolidation) or accumulation of caicareous shells by marine organisms; in broader sense means a formation, but sometimes also a material that contains very little clay
as matrix: an assemblage of mineral particles of various sires, shapes and chernical characteristics, together with organic materials in various stages of decomposition and living soi1 population; often implies structure and texture or fine grained particulate material; as profile refers to topsoil (A horizon, main source of plant nutrients) subsoil (B horizon, zone of clay formation) and weathered bedrock (C horizon, the deepest layer); used interchangeably with sediments or clay rich and fine grain deposits
Thermodynamics the science that deals with the interconversion of different foms of energy and with the spontaneous direction of processes; involves the study of heat, work and energy, their interconversions and the changes that they bring about; classical (also energetics or equilibrium) thenndynmics deals with the bulk properties of macroscopic systems at equilibrium, considering the initial and final state of a system and its surroundings
VOC volatile organic compounds; substances that contain carbon atoms and which have a minimum vapor pressure of 0.13 kPa (as defined by WHO and the USEPA) at standard temperature (293 K) and pressure (101 kPa)
APPENDIX 1 - SUPPLEMENT TO CHAPTER 1
Fig-Al. 1 Chernical fonnulae and structures of tested VFAs and VOCs ....................... 268
VOLATILE FATTY ACIDS : srnaIlest of the straight chah alkmoic acids (C,&Od
acetic (ethanoic) CH3COOH [64- 19-71 CAS Registry number
propionic (propanoic) CH3CH2COOH [79-09-41
butyric Putanoic) CH3C&CH2COOH [ 107-9241
MONOAROMATIC VOLATILE COMPOUNDS : BTEX
BENZENE C& [7 L 43-21 CAS Registq Number
TOLUENE C&CH3 (108-88-31
CH3 XYLENE isomers C&(CH3)* oc& 0 meta ortho [9S-47-61 108-38-3 13-dimthyi-h~ene 1.2-dimethyl-benzene
para [ l û6-42-31 1,4dimethyI-benzene
DrCHLOROMEmANE CH2Ci2 [75-09-21 CAS Registry Nurnber
TFUCHLOROETWLENE (TCE) CHCI= CClz [79-O 161; chmical name l , l J - trichioroetbylene
Fig. A l . 1 Chernical formulac and structures of tested VFAs and VOCs
APPENDIX 2 - SUPPLEMENT TO CHAP'IER 2
Fig.A2. 1 Degradation of DCM in the KVLL at 24OC - Batches 1,2, 3,4, 5 & 6: Data with Zero-order, First-order and Growth-linked models fit-lines . . . . . . . . . . . 270
Fig.A.2. 2 Removal pattern and kinetics lines upon a single addition of DCM in KVLL: Batches 7 & 8 at 24'C and 10°C with fit iines to Zero-order, First- order and Growth-linked models .... ... ... . ........................ . . . . . . . . . . . . . 271
Fig.A.2. 3 Degradation pattern and Michaelis-Menten lines upon a single addition of DCM in soil-KVLL suspensions - Batches 1 , 7 ,8 & 9 at 24°C and 10°C ...... 272
APPEMDJX 3 - SUPPLEMENT TO CHAPTER 3
A3. 1 Copyright release note .............................................................................. 274
A3. 2 Using POLLUTETM to estimate difision coefficient (adapted fiom User's Guide) ....................................................................................... 275
Fig. A3. 1 Monitoring gas generation in Source and Receptor solutions: Difision/degradation tests with synthetic leachate and Samia silty clay ........................................................................................................ 276
A3.1 COPYRIGHT RELEASE NOTE
Date: Fri, 9 Feb 2001 16:22:23 -0500 From: "Ryan, Karen" <[email protected]> To: 'Leila H' <[email protected]> Subject: FW: Copyright permission
Dear Ms. Hrapovic: Permission is granted for use of ASCE copyrighted material as per your letter (copy attached). This one time grant is subject to the following conditions:
(1) A credit line must be added to the material being reprinted must include title, author, copyright date, and publisher and indicate that the material is reproduced by permission of the publisher (ASCE).
Best of hck with your thesis.
If you decide not to use this material, please advise the grantor (ASCE) within 30 days. SincercIy,
Karen A. Ryan Business and Administrative Manager Copyrights and Permissions American Society of Civil Engineers 180 1 Alexander Bell Drive Reston, VA 20 19 1 [email protected] FAX: (703) 2954278Phone: (703) 29342 12
--Original Message- From: Perry, Jackie Sent: Thursday, February 08,2001 4145 PM To: Ryan, Karen Subjcct: FW: Copyight permission Karen; Cm you please help this person with copyright? Thanks. JAckie
- 4 r i ginal Message- From: 1. hrapovic [mailto:[email protected]~ Sent: Thursday, Febniary 01,200 1 454 PM To : jvpeny @asce.org Subject: Copyright permission Dear Ms. Peny,
My name is Leila Hrapovic and 1 have just mmpleted the work on Ph.D. thesis at the Universi& of Western Ontario. One chapter of the thesis contains the signincant portions of the article publiskd in the Journal of Geotechnicai and Geoenvironmental Engineering, ASCE, vol. 123. No. 12, p. 1085- 10% titled: Anaerobic Degradation of DCM Difiking through Clay. authored by: R Keny Rowe, Leila Hrapovic, Naim Kosaric, and D. Roy Cullimore. The thesis couid not be accepted and bound without the ASCE permission regardhg the copyright release which shouid be enclosed in the in the thesis Appendiu. Would p u please grant this permission and advise on the conditions of its use. In the meanwhile. accept my thanks and greetings. Leila Hrapovic, graduate studenç University of Western Ontario,Department of Civil and Environmental Engineering; hndon, Ontario, Canada, N6A 5B9; tel. (519) 661-211 1 ext 88338: k.x (519) 661-3912
275
A3.2 Using POLLUTETM to estimate diffufiion coellicient (adapted form User's Guide)
On the toolbar select Dewsit datq
TWe 1 Type in the titie with dcscripCim and file ~ M K 1 I
Nmnber of h y t n 1 SpaciS. the numba of hyas with dia& diffuJive pmpdes 1
Top Boundary : FlïNITE W
D - Y v w
h p i a c e Tndonn panmeten
Initial source umccntration
S M Ducy velocicy tbrough the Iaws), ZERO for pure difîhion
Aîccptdcfauhviluqsiactticcyucsrtisfactorymmoatcucs
1 Volume of leaduie m l l d
Botîom Bomiduy: FKED 0üTFU)W VEl.DClTY
On the toolbar select Layer Data
Refen to h i t e m;us of -a @a!&) diffirJuig thmugh layen;
i.e. in the SOURCE SOLUTION aimpartmcnt of a tuting ceIl
Specify Uiitid concentration C. ofthe spccia (usualfy) at tirne rcfo
SpccifL, a m ifthe peak concatdon is d e d eariy
Tkcn u H, (hcight of the source soluticm) for diEusion cats
Zao f o r ~ k i ~ ;
R d i n to the case whni depait is undcrlain by an aquifcr @ermeable
base); i.e. in the RECEPTOR SOLUTION cornpartment of a ccil
Tlkm u diamdm of the irrtifig œlt
U d l y u t t o ls~ithasnoinnuenctocithercsuhs
Taken as H,, @ci@ of r ibocptor solution) for diauion test
Takm as 1 (solution in mqtor conqmtment)
Zao if the arnounb removeci for sunpling are ncgligible
Nmber of mbiayers
thicknem
On the toolbar select Erecute
Usa! primuily in the output of tht calculaial concentrations w i h @th; spccify
Specity tht torrl thiduKs of erdi Iaycr witb dinernit diffiisive p p d e o
CONCENTUTIONS 1 Wcul.tc miamhant conccntntim u r d d depthr and iims wing the I
On the toolbar select OutDut
- O- CEIL I SOURCE *v. . CE1.L 2 SOURCE
-0- BLANK CONTWOL -4- CELL 1 RECEYrOR
I ai l= 1 1 3 days spi11 Ce11 1 Rcccptor
Time [days]
Fiy A3.1 MONlTORlNG GAS GENERATlON in SOURCE and RECEPTOR SOLUTIONS: Difision/degradation tests with synthetic leachate and Sarnia silty clay
FigA4.1 M o n i t o ~ g stability of VFAs concentrations in working solutions: (a) Halton Till suspensions (b) distilled water .............................................. 278
Fig. A4.2.1 Diffision of acetate through Halton Till: (a) source & receptor solutions (b) depth profiles; Variation of D = (2.5 - 4.5) x 10'" m2/s
....................... and mass balance calculation for Test 1 s h o w in Fig. 4.2 279
Fig. A4.2.1.1 Diffision of acetate through Halton Till: Variation of diffusion coefficients for Test 1 shown in Fig. 4.2: Supplement data and statistics ................................................................................................. 280
Fig. A4.2.2
Fig. A 4 2 3
Fig. A4.3
Fig. A4.4
Fig. A4.5.1
Fig. A 4 5 2
Difision of propionate through Halton Till: (a) source & receptor solutions @) depth profiles; Variation of D = (1 .5 - 2.5) x 1 0 " O m2/s and mass balance calculation for Test 1 shown in Fig. 4.3 ....................... 281
Diffision of butyrate through Halton Till: MIuence of D = (diffision coefficient, 0.9 - 2.5 x 10-'O m2/s), 110 (half-life), and Kd (linear sorption) on the best fit with mass balance calculation for Test 1 shown in Fig. 4.4 .................................................................................... 282
Schematic of cell assembly used in laboratory intnnsic degradation test: (a) 3 cm soi1 pIugs (b) 5 cm soi1 plugs ............................................. 283
Intrinsic degradation of organic chernicals fiom synthetic KVL leachate through compacted Halton Till: Ce11 1-2 (with 5 cm thick soil plug) after 33 days: (Iefl) front view of and (righl) back view of the ce11 .................................................................................................... 284
Variation of diffision coefficients elaborated for Fig. 4.13.1 : Intrinsic degradation of VFAs (as DOC): 3 cm soil plugs: (a) receptor solutions (b) depth profiles with recornmended D [mYs] ........................ 285
Variation of diffusion coefficients elaborated for Fig. 4.13.1 : Intnnsic degradation of VFAs (as DOC): 5 cm soi1 plugs: (a) receptor solutions (b) depth profiles with recommended D [m-1'1 ........................ 286
Time [dm]
Fig.M.1 Monitoring stability of VFAs concentration in working solutions: (a) Halton ta suspensions; (b) distilled water
KVL synthetic leachate
tank
KM, qnt hetic leachate
tank
Fig A4.3 Schematic of the ce1 assembly used in laboratory intrinsic degradation test: (a) 3 cm soi1 plugs and @) 5 cm soi1 plug
Fig. A4.4 lntrinsic degradation of organic chernicals fiom synthetic KVL leachate in compacted Haiton till: Cell 1-2 (with 5 cm thick plug ) afier 33 days: (lefr) fiont view and (right) back view of the cell (Note black spots due to formation of metallic sulfides on predominantly nisty soil. Plug holds intact, except for the disturbance induced by handling ai the top few millimeters of the interface with source solution. Slight increase in turbidity of the source solution indicates beginning of the microbial activity. The receptor solution remains clear.)
@) VFAs caiculated as DOC [&], in pore water
O Cell l- t @i 23 d q s
Cell 1-2 ria 29 &y3
- mnc - O .O 61 I O1 ILI if2 161 la7 360 4 9 1 63.1 119 tdl 939 1011 114.1 1139 1169 llt O 12)s 142 7 1563 1619 ia O
0 Ce11 IV-1 1 17 &y3 A Cell II-1 @ 162 day Ctll 11-2 @ 1 18 dry^ Cc11 Pi-2 @ t 63 bp rvaagcd for 117.5 &y rvmged for 116.5 dnp
Fig. A 4 5 1 Variation of difision coefficients elaborated for Fig. 4.13.1: INTRINSIC DEGRADATION OF VFAs (as DOC): 3 cm soi1 plugs; (a) receptor solutions and (b) depth profiles with recommended D [m2/s]; (data averaged for clarity)
Time [daysl
(b) VFAs calculated as DOC [#LI, in pore water
Fig. A 4 5 2 Variation of diffusion coefficients elaborated for Fig. 4.13 -2: INTRINSIC DEGRADATION OF VFAs as DOC: 5 cm soi1 plugs; (a) receptor solution and (b) depth profiles with recornmended D [m21s] range; (data averaged for clarity);
APPENDK 5 - SUPPLEMENT TO CHAPTER 5
Table A5.1
Table A5.2
Table A5.3
Table A5.4
Fig. A5.1
Fig. A5.2.1
Fig. A5.2.2
Fig. A5.3
Fig. A5.4
Summary of linear sorption coefficients (&) for natural soil and selected VOCs reported in literature ......................................................... 288
Cornparison of the vaiues of diffusion coefficients fiom this study and ............................................................................... values fiom literature 292
Calculation of toluene pore water concentrations using retardation coefficient R ............................................................................................. 293
Summary of First-order anaerobic rates and half-lives for selected .................................................. VOCs, (after Aronson & Howard, 1997) 294
Sorption of VOC on Haiton Till: cornparison of observed and predicted ........................................................................................................... Ky 295
Diffision of DCM through Halton tu: Variation of sorption coefficient, Kd and rate of loss in solutions (for Fig. 5.3) .......................... 296
Variation of K d and rate of loss for TOLUENE (for Fig. 5.7): (a) Fig. 5.7 D = 2.5 x 10'" m2/s, Kd = 1.5 crn3/~, loss in SS 1 1 , ~ = 180 days (b) Fig. 5.7 D = 2.5 x IO-'' m2/s, Kd = l .S no losses in SS (c) Fig. 5.7 D = 2.5 x 10"~m'/s, Kd = 2.5 cm3Ig, no llosses in SS ........................... 297
Variation of lag period and its effea on DCM impact in pore water: elaborated f?om data and fit lines showen in Fig. 5.14.l(b)(4) .................. 298
Variation of half-life in the pore water (soil) and its effect on DCM impact: elaborated from the data and fit-lines s h o w in Fig.
Table A5.1 Summary of linear sorption coeîricients ( K d ) Cor natural soi1 and selected VOCs reported in literature
Chernical foc 1%1 0 .3
2.59
0.44
0. 1 1
0.0 15
1.9
1.49
0.66
2.25
0.7
1.57
0.02
0.5-0.8
0.0 1
Soil description
Mwrlette B mil - fine-loaniy, mixed B, horizon;
(sand: 38.8%; silt: 3 1.6%; clay: 29.6%) pH 5.4; CEC 16.3
Marlette A soi1 - fine-lmy, rnixeû A horizon;
(sand: 56.6%; sill: 22%; clay: 21.4%) pH 6.4; CEC 14.6
Sî. Clair soi1 - fine illitic B, horizon;
(sand: 21%; silt: 34.9'?/0; clay: 44.1%) pH 6.72; CEC 22.6-18.3
Oshtemo d l - coarsc-loamy, mixed B, horizon;
(sand: 89.3%; sill: 4.4%; clay: 6.3%) pH 5.84; CEC 3.5-4.5
Citma Rewamh & Education center sandy aquifcr, FL; Kd calculated from R = 1.4
(sand: 96.5%; silt: 1.7%; clay: 1.8%) pH 7.4; CEC 3.5-4.5; 0 = n = 0.3; p = 1.8
Woodbum silt lowm (sand: 9.b. silt: 2 1%; clay: 68%); CEC 14;
Cwptina silt loam (sand: 7.7%; silt: 62.5%; clay: 29.9%) pH 5.33; CEC 1.15
McLaurin sandy lowm (sand: 74.9%; silt: 20.4%; clay: 4.7%) pH 4.92; CEC 10.15
Owkville A (sand: 94.5%; silt: 3.5%; clay: 2%)
Owkville B (sand: 94.5%, silt: 4%; clay: 1.5%)
Pipestone (sand: 95%; silt: 3.1%; clay: 1.9%)
Ringe clvyey till (coarse silt + sand: 50%; fine silt: 43Y0; clay: 7%) pH 7.8; CEC 6.56
Sarnia gîicio-lacustrine clvyey till (siind + gravel: 16%; silt: 44%; clay:40%)
Sarnia glucio-lacustrine clvyey till (sind + gravel: 16%; sill: 44%; clay:40%)
Lee et al. 1989
Angley et al. 1992
Chiou et al. 1983
Walton et al. 1992
Maraqa et al. 1998
Broholm et al. 1999
Myrand et al. 1987
Johnson et al. 1989
Table AS. 1 continued
Totuene
Soi1 description
Marlette B soi1
Mwrlette A sail
St. Clair mil
Oshtemo mil
C i t r u y Rea & Educwtion Center snndy wqufer, FL.; & çalculaîcd fram R = 1.7;
Captina mil
McLaurin mil
Ringe clayey till
Sarnia clayey till
Silty clay I (fine sand:S6%; silt: 32%; clay: 12%) pH 7.83; CEC 70; Kd :fmm ksorption
Silty chy II (fine sand:5%; silt: 70%; clay: 25%) pH 8.05; CEC 42; &: frorn desorption
Siltu clay III ( a d :sili: clay not deteniiined) pH 7.05; CEC 18.6; & : fmm dçsorption
Course synd IV-A; & detennined froni desorption test
(coarse sand: 40%; fine sand: 39%; fmction >2mm and <0.001 mm: 2 1%) pH 6.8; CEC 34.4
C o w m swnd IV-B; determincd frorn desorption test
(coarsc sand: 56%; fine wid: 24%; fraction >2mrn and c 0 . 0 1 min: 20%) pli 6.6; CEC 32
Sarnia glucict.lwcustrine clwyey till (sand + gravel: 16%; silt: 44%; chy:40%)
Reference
Lee et al. 1989
Angley et ai. 1 992
Walton et al. 1992
Broholm et al. 1999
Myrand et al. 1987
Pavlostathis &
Mathavan , 1992
Johnson et al. ï 989
Chernical
Et hyl-
h e m n e
Xyleneu)
(W
m&p-Xa
0-x
PX
Table A5.1 continued
Soil description
Marlette B soi1
Mariette A mil
Sî. Clair soi1
Oshtemo mil
Citrus Re& & Educ. Center svndy wquifer, PL.; l& calculakd from R = 1.7;
Woodbum silt loam
Riage clayey till
Sarnia glucio-lacustrine clayey till
Silty clay II
Silty clay III
Coarae sand IV-A
Coarse wnd W B
Cirtms Rea & Edu. Center sandy wquifer, FL.; calculated from R = 2.0;
Cirtms Re& & Edu. Center nandy uquifer, PL; & calculated from R = 1.7
Ringe clayey till
Ringe clayey till
Captina soi1
McLaurin mil
k et al., 1989
Angley et al., 1992
Chiou et al., 1983
Bmholm et al, 1999
Johnson et al. 1989
Pavlostathis &
Mathavan, 1992
Angley et al. 1992
Bmholm et al. 1909
Walton et al. 1992
Table A5.1 continued
TCE
Soil description
Mariette B mil
Marieîîe A soi1
Silty clay 1; & dcterrnined from desorption test;
Silty clay II: Kd determincd from desorption test
Silty clwy 111; Kd determin4 from desorption test;
Silty sand (sand: 56%; silt: 32%; clay: 12%); CEC 70
Sarnia glacio-lacustrinc clayey till
Sarnia glacio-lacustrine clayey till
Orange silty clay loam (sand: 23%; silt: 42%; clay: 35%); CEC 15.5
Dark gray silt loum (sand: 17%; silt: 65%; clay: 18%); CEC 29*8
Reference
Lee et al. 1989
Pavlostathis &
Mathavan, 1992
Pavlmthis &
Jaglal, 1991
Myrand et al. 1987
Johnson et al.
1989
Bal1 et al. 1997
Soil description is prwrited as characterizcd in the cited teferences.
Kd [cm3/g/gl linear sorption (partitioning) coefficient; unless othcrwir noid. al1 coeffcient demmined from batch quilibrium sorption tests
K, = - x 100% ; orgitnic carbon/water partitioning coefficient or "adsorption welficient" (Lyman et al.. IY82);/, % of' o r p i c carbon content in soi1 f,
CEC [meq/100 g] cation cxctiange capacity
pK', R = i + - , retardation coelficient; n 1-1 = porosity; p ~ g l c r t ~ ~ l = dry detisity; 0 1-1 = volun~eiric moisture content = n for 100% saturation I l
Table A5.2 Coniparisoii of the values of diffusion coefficieiits Fruni this study and values frorn literature
Vlilucs from literaturc Reuulty fram this study
IWM
1.2-NA
TCE
iknzcnc
Tolucnc
Elhyl-
btnzcnc
n&p-XI
O-Xylenc
' Howc & hrc -
I
n el al., 1999b;
Soi1 description (see Table A5.1 for
additional infonnatioa)
Halton till', same as in this study
Not found
Sarnia clay2p= 1.6; n=0,37; k<5x IO-"
Samia clay'p= 1.77; n4 .34; k=(8-0.8)~ 1 0 . ' ~
OSCL' p=1.22; n=0.53; k=(2-30)xl0-'~
DGCL' p= 1.15; n=0.56; k=(2-30)x 1 0 " ~
Mass recovcry [ %)
Sarnia claf
Sarnia claf
Ringe lil15 p= 1.96; n=0.27; k=2.7xlod
Saniia claf
Sarnia claf
Ringe tills
Sarnia clay6; p=1.68; n=0.39; k not reponed
Sarnia clay2
Ringe til15
Ringe til15
iünge til15 I
krll cc al., 1997; '~rot i
l'able A5.4 S u n î m i i ~ of First-order anaerobic rates and half-lives kr the selected VOCs, (afier Aronson & Howard, 1997)
Benzcnc
1 Range. al1 studies
Man, al1 studics
Mcan, laboratory siudies
Range, fieldlin situ studics I Mcan, fieldlin situ studies
Range, NO3 redircing studies
Mean, NO3 rcâucing studics
Range, Fe(ll1) reducing studics I 1 Mean, Fe(1ll) rcducing studies
1 Range, S 0 4 rcducing studies 0-0.047 nd- 15 0.0 16 43.5 04,052 nd-13.5 0.005 139
Mcan, S 0 4 rçducing studies
Range, mct hanogeriic studics
Mean, met hanogenic studies
I Suggcnted range of rates: lowcr bt nican licldlin situ litiiits
top numbcrW in the box = w dm1 order Wnelin mvcrubic nte, Li, (dup 'J rrpur(td ur givrn in Amruun & Ilmrd, 1997; n t - are d r r i v d Rwn ihc tt.(. with dlffercat typa o f d or sedimrnts, wll briirvtd tu be froni wqulfrini or r a b wilh mhtlvcly hlgh hydnullc conducîivlty and lm &y content botlam nwnbcris) = b m hwlf iifc Iduys) cwlrulwted au #,,A = tn th,, ud &cd Cor cinrltr; nd - no( dc~radabk; na - no4 uvdhbk
Iog K, predicted
OCuidrbog. 1979)
Table 1.1 log Koc - btKair- 031
DCM 1.25 1.2-DCA 1.79 TCE 2.53 B 2.13 T 2.69
EB 3.15 m-Syl 3.2 pSy l 3.18 +Sv t 3.04
DCXl
I .2-DCA
TCE B T EB m-Syl p S y l 0-Xyl
obwned K oc
Koc =Kd *lWtor
Iaw foc hi foc pltnm foc
0.14% 0.45% 0.29%
0 ~ d a t r f o r t a t t d c h , c m i ~
- kmr rc@w&ioa for Iü data and f, = 0.45 %:
~ o b s n v e d = 0.49 K g d i c t c d + 1-26 . ~'=0.790
- . lincar regusion for /,=û.45%, DCM not inclu&d: K, obsavcd = 0.72 K,prrdicttd + 0.64. ~ ' 4 . 9 7 4
Kd = foc @Koc/ 100 (Lyman 1990)
low foc hl foc interm foc
o k m â log Koc
low foc hl foc intcrm fa 0.14% 0.45% 0.29%
Fig.AS.1 Sorption of VOC on Halton Till: Comaprison of observed and prediaed Ks
Fig. 5.3 DIFFUSION OF DCM THROUGH HALTON TILL (a) source SS and receptor RS solutions (b) depth profile
Fig. A5.2.1 DIFFUSION OF DCM THROUGH HALTON TILL: VARIATION of SORPTION COEFFICENT, K, and RATE of LOSS in SOLUTIONS (for Fig. 5.3)
Time [days]
Fig. 5.7 DIFFUSION OF TOLüENE THROUGH HALTON TILL (L) source SS and receptor RS solutions (R) depth profile
O 1 2 3
Tolumc in Porc Watcr [mgLI
Fig. A5.2.2 Variation of K, and rate of loss for TOLUENE ( for Fig. 5.7): -10 2 a) Fig. 5.7 D = 2.5 x 10 m /s K, = 1.5 cm31g, loss in SS t,, =18O days 10 2 b) Fig 5.7 D = 2.5 x 10- m /s K,= 1.5 crn3/g, no lossesin SS
-10 2 c) Fig. 5.7 D = 2.5 x 10 rn /s K' = 2.5 cm3/% no losses in SS
DCM in pore water [mglL]
35 day lag I
Fig. A5.3 Effect of variation oflag period on degradation of DCM: elaborated nom data and fit lines showed in Fig. 5.14.1 b(4)
[ml t ~n = 0.75 d n Liytr I onîy l
lag S 5 d q x lag 45 days 1% 35 day3
DCM in pore water [m@]
--
A Data 55 day lag (Layer 1 oniy) 90 day lag (Layer I only)
- - - 90 day lag : t 0.75 d (L 1) 3654 (below L 1)
-- 90 day lag t i/r O.?Sd (LI) lûûd (be10~ Ll )
- - - 9û day lag t,,z= 0.75d (Ll) SOd (klow L1)
---- 90 day lag t in= 0.75d (Ll) 10d @~Iow Ll)
- - -
t ln-0.75dinL 1 only
iag 3 5 days lag 9û day
Fig. A5.4 Effect of variation of half-life in the soi1 on degradation of DCM: elaborated from the data and fit-lines showed in Fig. 5.13.2b(2)
