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Comparative studies of aerobic and anaerobic landfills using simulated landfill lysimeters

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Comparative studies of aerobic and anaerobic landfills using simulated landfill lysimeters
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Kim, Hwidong
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xv, 231 leaves : ill. ; 29 cm.

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Average linear density ( jstor )
Heavy metals ( jstor )
Landfills ( jstor )
Leaching ( jstor )
Lignin ( jstor )
Lysimeters ( jstor )
Methane ( jstor )
pH ( jstor )
Solid wastes ( jstor )
Sulfides ( jstor )
Dissertations, Academic -- Environmental Engineering Sciences -- UF
Environmental Engineering Sciences thesis, Ph. D
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bibliography ( marcgt )
theses ( marcgt )
non-fiction ( marcgt )

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Thesis (Ph. D.)--University of Florida, 2005.
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Includes bibliographical references.
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Printout.
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Vita.
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by Hwidong Kim.

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COMPARATIVE STUDIES OF AEROBIC AND ANAEROBIC LANDFILLS
USING SIMULATED LANDFILL LYSIMETERS















By

HWIDONG KIM














A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2005































Copyright 2005

by

Hwidong Kim































This document is dedicated to my parents and loving wife














ACKNOWLEDGMENTS

I would like to thank my advisor, Dr. Timothy G. Townsend, for showing such great patience as a mentor. He gave me this great opportunity to study on the field of solid waste. He also showed me the way of living as an engineer, professor, and a family man. I cannot forget his tears when Mr. Townsend passed away. I would also like to thank my committee members, Dr. Angela Lindner, Dr. Frank Townsend and Dr. Roger Nordstedt, and my other spectacular faculty members, Dr. David Chynoweth, Dr. Gabriel Bitton and Dr. Matthew Booth, who gave me great help.

I wish to thank my colleagues in the Solid and Hazardous Waste Research group, in particular, Brajesh Dubey, Qiyong Xu, Kim Cochran, Steve Musson, Aaron Jordan, Pradeep Jain, Jaehak Ko, Murat Semiz, Judd Larson, and Yong-Chul Jang, a faculty memeber of Chung-Nam University in South Korea. I also thanks goes to my first mentor and graduate advisor, professor, Byung-Ki Hur, a faculty of Inha University in South Korea.

A special thanks goes to my mother as well as my father, who is fighting against disease. Finally, greatest thanks go to my wife, Eunkyoung Choi, for her patience, encouragement, and love.










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TABLE OF CONTENTS

pne

A CKN OW LED G M EN TS ................................................................................................. iv

LIST O F TA BLES ........................................................................................................... viii

LIST O F FIG U RES ............................................................................................................. x

CHAPTERS

1. IN TROD UCTIO N .......................................................................................................... I

1. 1 Problem Statem ent ................................................................................................. 1
1.2 Objectives .............................................................................................................. 2
1.3 Research Approach ................................................................................................ 3
1.4 O utline of D issertation ........................................................................................... 5

2. COMPARATIVE STUDIES OF LEACHATE AND GAS QUALITY OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILL BIOREACTORS ......... 6

2.1 Introduction ............................................................................................................. 6
2.2 M aterial and M ethods ............................................................................................ 7
2.2.1 G eneral D escription of the Lysim eter ......................................................... 7
2.2.2 Tem perature Control ................................................................................... 8
2.2.3 Fabricated W aste Stream ............................................................................. 9
2.2.4 A ir Injection ............................................................................................... 10
2.2.5 Leachate and G as Analysis ........................................................................ 10
2.2.6 Recovery of the A naerobic Lysim eters ..................................................... 11
2.2.7 Prediction of W aste M ass Loss ................................................................. 12
2.3 Results and D iscussion ........................................................................................ 12
2.3.1 pH .............................................................................................................. 13
2.3.2 O rganic Carbon Concentration .................................................................. 14
2.3.3 N itrogen ..................................................................................................... 16
2.3.4 D issolved Solids Content .......................................................................... 17
2.3.5 O xidation Reduction Conditions ............................................................ 18
2.3.7 G as Q uality ................................................................................................ 19
2.4 D iscussion ............................................................................................................ 20
2.4.1 Differences between Aerobic and Anaerobic Lysimeters ......................... 20
2.4.2 The Comparison of Leachate Parameters with Other Studies ................... 21
2.4.3 Im plications for Full-scale Application ..................................................... 22


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2.4.4 Lim itations ................................................................................................. 23
2.5 Conclusions .......................................................................................................... 24

3. THE FATE OF HEAVY METALS IN SIMULATED LANDFILL
BIOREACTORS, UNDER AEROBIC AND ANAEROBIC CONDITIONS ............ 47

3.1 Introduction .......................................................................................................... 47
3.2 M aterials and M ethods ........................................................................................ 48
3.2.1 H eavy M etal Sources in Synthetic W aste ................................................. 48
3.2.2 Sam pling M ethods ..................................................................................... 49
3.2.3 A nalytical M ethods ................................................................................... 49
3.3 Results and D iscussions ....................................................................................... 50
3.3.1 Changes in Metal Concentrations versus Time and the Percentage of
M ass Loss ......................................................................................................... 50
3.3.1.1 A lum inum ........................................................................................ 50
3.3.1.2 A rsenic ............................................................................................ 51
3.3.1.3 Chrom ium ........................................................................................ 53
3.3.1.4 Copper ............................................................................................. 54
3.3.1.5 Lead ................................................................................................. 56
3.3.1.6 Iron .................................................................................................. 57
3.3.1.7 M anganese and Zinc ........................................................................ 58
3.3.2 Organic Wastes as Absorbents of Heavy Metals ...................................... 59
3.4 D iscussion ............................................................................................................ 61
3.4.1 O verall Com parison of M etal Behavior .................................................... 61
3.4.2 Com parison to O ther Studies ..................................................................... 63
3.4.3 Im plication for D isposal of H eavy M etals ................................................ 65
3.4.4 The Im pact of A ir on M etal M obility ........................................................ 66
3.5 Conclusions .......................................................................................................... 67

4. THE EVALUATION OF LIGNOCELLULOSIC WASTE DECOMPOSITION OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILLS .................................. 95

4.1 Introduction ........................................................................................................... 95
4.2 M aterials and M ethods ........................................................................................ 97
4.2.1 Com position of Fabricated W aste ............................................................. 97
4.2.2 Excavation and Processing of Decomposed Solid Waste ......................... 97
4.2.3 M ethane Y ield D eterm ination ................................................................... 98
4.2.4 Cellulose and Lignin D eterm ination ....................................................... 100
4.2.5 D ata A nalysis ........................................................................................... 101
4.3 Results ................................................................................................................ 101
4.3.1 M ethane Y ield of Raw W aste .................................................................. 101
4.3.2 Solid W aste Excavation ........................................................................... 102
4.3.2 M ass Loss for Individual Com ponents .................................................... 103
4.3.3 Biodegradability of Excavated W astes .................................................... 104
4.3.4 B iodegradability of W ood W aste ............................................................ 105
4.4 D iscussion .......................................................................................................... 106
4.5 Conclusions ........................................................................................................ 108


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5. LANDFILL SETTLEMENT BEHAVIOR WITH WASTE DECOMPOSITION ..... 119

5.1 Introduction ........................................................................................................ 119
5.2 M aterials and M ethods ...................................................................................... 120
5.2.1 Lysimeters ............................................................................................... 120
5.2.2 Application of Overburden Pressure ....................................................... 121
5.2.3 Compression Index and Phase Separate M ethod ..................................... 122
5.2.4 Estimation of M ass Loss ......................................................................... 123
5.2.5 Volume Loss versus M ass Loss .............................................................. 124
5.3 Results ................................................................................................................ 125
5.3.1 Settlement Behavior over Time ............................................................... 125
5.3.2 The Relationship between The Settlement and Mass Loss ..................... 126
5.3.3 Ultimate Settlement ................................................................................. 127
5.4 Discussion .......................................................................................................... 127
5.4.1 Compression Index .................................................................................. 127
5.4.2 Correlation of M ass Loss and Volume Loss ........................................... 128
5.4.3 Application .............................................................................................. 129
5.5 Conclusions ........................................................................................................ 131

6. SUM M ARY AND CONCLUSIONS ......................................................................... 141

6.1 Summ ary ............................................................................................................. 141
6.2 The Implication of This Research ...................................................................... 143
6.3 Conclusions ........................................................................................................ 145
6.4 Future W ork ....................................................................................................... 147

APPENDIX

A. ADDITIONAL PROCEDURES AND CONCEPTS ................................................. 149

A. I Prediction of M ass Loss by Gas and Leachate ................................................. 149
A.2 Estimation of Biodegradable Volatile Solids (BVS) ........................................ 152
A.3 Lysimeter Dismantlement ................................................................................. 153

B. SUPPLEM ENTAL FIGURES ................................................................................... 159

C. LYSIMETER EXPERIMENT RAW DATA AND GRAPHS ............................... 169

C. I Graphs ............................................................................................................... 169
C.2 Raw Data ........................................................................................................... 189

LIST OF REFERENCES ................................................................................................. 219

BIOGRAPHICAL SKETCH ........................................................................................... 231






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LIST OF TABLES

Table page

2-1. MSW components .......................................................................... 25

2-2. Parameters and methods for analysis ..................................................... 26

2-3. Comparison of initial and final characteristics of the aerobic lysimeters............. 27

2-4. Comparison of initial and final characteristics of the anaerobic lysimeters.......... 28

2-5. Comparison of leachate parameters with other aerobic landfill studies .............. 29

2-6. Comparison of leachate parameters with other anaerobic landfill studies ........... 29

3-1. Heavy metal sources in fabricated waste stream ......................................... 69

3-2. Results of statistical analysis of metal leached between aerobic and anaerobic ....69 3-3. The amount of leachate produced and used for analysis................................ 69

3-4. Leachability of As, Cr, and Cu ............................................................ 70

3-5. Comparison of cumulative mass of metal dissolved in leachate and adsorbed on
lignocellulosic materials.................................................................. 70
3-6. Comparison of average metal concentrations of the aerobic and anaerobic
lysimeters with MSW leachate and regulatory levels ................................. 71

3-7. Comparison of characteristics of CCA-treated wood used for Jambeck (2004) and
this study................................................................................. 71

4- 1. Methane yields, VS and mass fraction of the lignocellulosic materials in raw
waste .................................................................................... 109

4-2. Comparison of methane yields of MSW with other studies.......................... 109

4-3. Biodegradable volatile solid (BVS) of organic fraction of the raw waste........... 109

4-4. The physical characteristics of excavated waste ....................................... 110

4-5. Overall methane yields of waste layers of the lysimeters 2 and 4 .................111II


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4-6. Summary of cellulose and lignin content of the wood samples ............................... 112

5-1. (Ca)min and (Ca)max values of lys I through 4 ........................................................... 132

5-2. k values of aerobic and anaerobic lysimeters ........................................................ 132

5-3. Comparison of compress indices between current study and other studies ............. 133

A- 1. Actual mass loss and predicted values of the aerobic and anaerobic lysimeter ...... 155 A-2. Mass and density of wastes excavated by depth ................................................. 155

C-1. pH of the aerobic and anaerobic lysimeters ...................................................... 189

C-2. Conductivity of the aerobic and anaerobic lysimeters ......................................... 192

C-3. Alkalinity of aerobic and anaerobic lysimeters ...................................................... 194

C-4. Total dissolved solids (TDS) of aerobic and anaerobic lysimeters .......................... 196

C-5. Total organic contents (TOC) of aerobic and anaerobic lysimeter ........................ 197

C-6. Chemical oxygen demand (COD) of aerobic and anaerobic lysimeter ................... 199

C-7. NH3 -N concentrations of aerobic and anaerobic lysimeters ............... 201

C-8. Sulfide concentrations of aerobic and anaerobic lysimeters ................ 203

C-9. Volatile fatty acids (VFA) of lysimeter 1 ............................................................205

C- 10. Volatile fatty acids (VFA) of lysimeter 2 ..........................................................206

C-11. Volatile fatty acids (VFA) of lysimeter 3 .................................................. 207

C- 12. Volatile fatty acids (VFA) of lysimeter 4 ...................................................... 209

C- 13. Fabricated waste in lysim eters .......................................................................211

C- 14. M etal concentrations of lysim eter 1 ....................................................................212

C- 15. Metal concentrations of lysimeter 2 .................................................................. 213

C- 16. Metal concentrations of lysimeter 3 .................................................................. 214

C- 17. Metal concentrations of lysimeter 4 ................................................................ 216

C- 18. ANOVA results of metals and organic absorbence ......................................... 218




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LIST OF FIGURES

Figure pae

2- 1. Schematic of the lysimeter ................................................................ 30

2-2. The composition of fabricated municipal solid waste for this research............... 31

2-3. Comparison of pH between aerobic and anaerobic lysimeters versus time .......... 32

2-4. Changes in COD of aerobic and anaerobic lysimeters versus time ................... 33

2-5. Changes in BOD of aerobic and anaerobic lysimeters versus time ................... 34

2-6. Changes in VFAs of aerobic and anaerobic lysimeters versus time (A) acetic acid
only and (B) acetic acid, propionic acid and butyric acid ............................ 36

2-7 Changes in the ratio BOD/COD of the aerobic and anaerobic lysimeters over time ..37 2-8. Changes in ammonia concentrations versus time........................................ 38

2-9. Changes in TDS of the aerobic and anaerobic lysimeters versus time................ 39

2-11. Changes in sulfide and pH versus time ................................................. 41

2-12. The changes in sulfate and sulfide versus time in the presence of dissolved
oxygen ................................................................................... 42

2-13. The changes in air injection rate and gas concentrations of aerobic lysimeter ....43 2-14. Changes in gas concentrations of anaerobic lysimeter 4 .............................. 44

2-15. Cumulative biogas vs. days in aerobic and anaerobic lysimeters.................... 45

2-16. Changes in gas concentrations, pH and gas generation rate after air injection into
lysimeter 3 ................................................................................ 46

3-I. Changes of Al concentrations over time.................................................. 72

3-2. Changes of As concentrations over time ................................................. 73

3-3. Changes of Cr concentrations over time.................................................. 74



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3-4. Changes of Cu concentrations over time ................................................. 75

3-5. Changes of Pb concentrations over time ................................................. 76

3-6. Changes of Fe concentrations over time.................................................. 77

3-7. Changes of Mn concentrations over time................................................. 78

3-8. Changes of Zn concentrations over time ................................................. 79

3-9. Distribution of As over a C-pH diagram ................................................. 80

3- 10. Potential- pH diagram of Cr ............................................................. 81

3-1 1. Distribution of Cu over a C-pH diagram................................................ 82

3-12. Adsorption of metal on solid wastes..................................................... 83

3-13. The comparison of aerobic and anaerobic lysimeters in respect of total mass of
metals adsorbed on lignocellulosic materials........................................... 85

3-14. The comparison of metal concentrations adsorbed on organic (newspaper and
cardboard) and plastic waste............................................................. 87

3-15. Fate of heavy metals thermodynamically occurred in aerobic (oxidizing) and
anaerobic (reducing) conditions ......................................................... 89

3-16. Comparison of concentrations of metal leached between aerobic and anaerobic
lysimeters .................................................................................. 90

3-17. Changes in cumulative mass of meta released over a mass loss, % ................. 92

3-18. Comparison of As, Cu and Cr leaching trend of the lysimeters to other study ....94 4- 1. The dry weight differences between predicted and measured remaining mass...113 4-2. Comparison of dry weights between raw and decomposed lignocellulosic wastes. .114 4-3. The changes in the percentage of waste components after decomposition; (A) raw
waste components and (B) decomposed waste (aerobic) ........................... 115

4-4. Changes in cumulative methane volume of lignocellulosic materials over time..116 4-5. Methane yields and weight differences of lignocellulosic materials among raw
and two lysimeters (A) all lignocellulosic materials; (B) wood only.............. 117

4-6. The comparison of dry masses measured and predicted by gas generated and
BMP assay................................................................................ 118



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5- 1. The changes in settlement, cumulative gas (C02) and pH over time................ 134

5-2. The changes in settlement, cumulative gas (CO2 and CH4) and pH over time......135 5-3. Settlement behaviors and compression coefficients of aerobic and anaerobic
lysimeter over a period of time......................................................... 136

5-4. Relationship between settlement and overall mass loss of the aerobic and
anaerobic lysimeters..................................................................... 137

5-5. Relationship between percentage of settlement and mass loss....................... 138

5-6. Correlation of logarithm of mass loss of the aerobic lysimeters over time ......... 139

5-7. Different k values of anaerobic lysimeters at lag and log phases .................... 139

5-8. Settlement prediction of the aerobic lysimeters ........................................ 140

A-i1. Schematic of mass loss by waste decomposition...................................... 156

A-2. Waste mass loss by TOC and gas generation .......................................... 158

B- 1. Schematics of aerobic and anaerobic lysimeters used for this research ............ 159

B-2. The carriage system ...................................................................... 160

B-3. A schematic of the temperature control system........................................ 161

B-4. Schematic of gas volume measuring tool; before gas measurement, fill tap-water
up to the top scale........................................................................ 162

B-5. The nation-wide composition of discarded municipal solid waste in 2003......... 163

B-6. The composition of municipal solid waste in Florida in 2000....................... 163

B-7. (A) 'Blue water phenomenon' observed from gas collection system of aerobic
lysimeters; (B) a hole on copper tube caused by corrosion of Cu.................. 164

B-8. Solid samples excavated from one of the aerobic lysimeter ......................... 165

B-9. Decomposed papers were commingled together (aerobic lysimeter) ............... 166

B-l10. Not well degraded office paper (aerobic lysimeter).................................. 167

B-i 1 Wood blocks excavated from aerobic lysimeter ...................................... 167

C- 1. The change in COD of the lysimeters over the percentage of mass loss ........... 169





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C-2. The change in BOD5 of the aerobic and anaerobic lysimeters over the percentage

of m ass loss ............................................................................................................ 170

C-3. The change in ammonia of the aerobic and anaerobic lysimeters over time ........... 171

C-4. The change in fluoride of the aerobic and anaerobic lysimeters over time ............. 172

C-5. The change in chloride (CI) of the aerobic and anaerobic lysimeters over time .... 173 C-6. The change in sulfate of the aerobic and anaerobic lysimeters over time ............... 174

C-7. The change in calcium (Ca) of the aerobic and anaerobic lysimeters over time ..... 175 C-8. The change in sodium (Na) of the aerobic and anaerobic lysimeters over time ...... 176 C-9. The change in biogas produced from the aerobic lysimeters ................................... 177

C-10. The change in biogas produced from the anaerobic lysimeters ............................. 178

C-1 1. Al concentration versus pH in leachate from the lysimeters ................................. 179

C- 12. Cr concentration versus pH in leachate from the lysimeters ................................. 180

C-13. Cu concentration versus pH in leachate from the lysimeters ................................. 181

C-14. Mn concentration versus pH in leachate from the lysimeters ................................ 182

C-15. Pb concentration versus pH in leachate from the lysimeters ................................. 183

C- 16. Zn concentration versus pH in leachate from the lysimeters ................................. 184

C-17. Change in methane yields of the waste layer 2-1 and 2-2 ..................................... 185

C-18. Change in methane yields of the waste layer 2-3 and 2-4 ..................................... 186

C- 19. Change in methane yields of the waste layer 4-1 and 4-2 ..................................... 187

C-20. Change in methane yields of the waste layer 4-3 and 4-4 ..................................... 188













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Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy

COMPARATIVE STUDIES OF AEROBIC AND ANAEROBIC LANDFILLS
USING SIMULATED LANDFILL LYSIMETERS By

Hwidong Kim

December 2005

Chair: Timothy G. Townsend
Major Department: Department of Environmental Engineering Sciences

Many proposals suggest that air injection into bioreactor landfills enhance waste composition; several potential benefits of air addition have been hypothesized, yet little has been proven about the overall performance of aerobic landfills compared with current anaerobic landfills. Utilizing research conducted with six-foot tall stainless steel simulated landfill lysimeters, complete with fabricated wastes, this Ph.D. dissertation compares aerobic and anaerobic landfills with respect to gas and leachate quality, fate of metals, settlement behavior and biodegradation of lignocellulosic materials.

Through air injection, a large enhancement of waste decomposition was observed. More than 90% of the maximum chemical oxygen demand (COD), biochemical oxygen demand (BOD) and total organic carbon (TOC) concentrations decreased within 100 days. During the methanogenic phase in the anaerobic condition, concentrations of ammonia increased by an amount four times greater than the initial concentrations. A large change of ammonia was not observed from the aerobic lysimeters.



xiv








The fate of metals leached from the various metal sources including cathode ray tube (CRT) monitor glass and ground CCA-treated wood were explored. Metal leaching trends observed varied from anaerobic to aerobic lysimeters; the average concentrations of As, Fe, Mn, and Zn in the anaerobic lysimeters proved significantly greater in concentration than observed in the anaerobic lysimeters. Furthermore, significantly greater concentrations of Al, Cu, Cr, and Pb were detected in the aerobic lysimeters as compared to the anaerobic lysimeters.

Using leachate and gas measurements, mass losses from the aerobic and anaerobic lysimeters were estimated. Mass removed from the wastes was primarily converted into gas; after the water was removed from the lysimeters, the mass of waste excavated from each lysimeter was compared with the estimated loss mass. For wood waste, no great influence on air addition was observed through cellulose/lignin analysis. Methane potential of lignocellulosic materials other than wood waste resulted in great differences of biodegradation between aerobic and anaerobic lysimeters.

The landfill settlement behavior occurring in aerobic and anaerobic simulated

landfills was mathematically analyzed. The logarithm of mass loss was linearly correlated with the percentage of settlement. With this relationship, the secondary settlement of bioreactor landfills could be mathematically modeled using the first-order exponential function.












xv













CHAPTER 1
INTRODUCTION

1.1 Problem Statement

Landfills remain the predominant method for managing municipal solid waste

(MSW) in the U.S. Although modem engineered landfills protect the environment from groundwater contamination and in some cases gas emissions, they are most often operated in a fashion where only a small amount of the disposed waste is permitted to biodegrade to a more stabilized state. This results in large amounts of undegraded waste being stored for many years in the future; their management will continue to demand resources and may pose a long-term environmental risk.

Alternatively, many innovative and more environmental-friendly strategies for operation of MSW landfills have been proposed (Stegmann, 1983; Barlaz et al., 1992; Komilis et al., 1999). Among these techniques, leachate recirculation has been found to be the most practical approach for enhancing waste decomposition and stabilization in landfills (Reinhart et al., 2002). This process stabilizes landfilled waste more rapidly because of the increased moisture content and the more effective distribution of nutrients and microorganisms in the landfill. This result creates a very favorable environment for the existing anaerobic organisms responsible for waste degradation. If controlled, methane produced can be utilized as a resource. This technique has changed the concept of a landfill from a historical garbage dump to a bioreactor, where various biochemical reactions are managed in a controlled fashion.





1





2


Air addition has been suggested as another means, in concert with leachate recirculation, to achieve rapid landfill stabilization. It has been reported that waste decomposes more rapidly in aerobic systems relative to anaerobic systems (Read et al., 2001). Additional reports suggest that air injection may stop the production of methane (one of the most serious greenhouse gases), change the leachate quality for the better, reduce the amount of volatile organic compounds (VOCs), and improve the degradability of anaerobically recalcitrant materials (Grima et al., 2000; Read et al., 2001; Lee et al., 2002; Reinhart et al., 2002). Some of these potential benefits have been investigated at the lab scale (Stessel and Murphy, 1992), and some positive outcomes have been reported from field studies (Read et al., 200 1; Lee et al, 2002). However, in order to apply this new technique successfully to full-scale operating landfills, further investigation is necessary. While anaerobic bioreactors have been heavily simulated in previous studies, there are few cases involving the simulation of aerobic landfills. It is also rare to find side-by-side simulations on the same waste stream under the same field conditions comparing aerobic and anaerobic systems.

1.2 Objectives

The main objective of this research was to compare aerobic and anaerobic landfills using simulated landfill lysimeters. In the early development of anaerobic bioreactors, several fundamental simulated landfill experiments were performed that have provided much of our understanding of such processes to date (Pohland, 1980). This research presents the results of parallel aerobic and anaerobic simulated bioreactors. Several different parameters of concern were investigated: leachate and gas quality, settlement, heavy metal fate, and decomposition of lignocellulosic materials. The following were specific objectives of this research:





3

o To compare leachate and gas quality between aerobic and anaerobic bioreactor

landfills,

o To explore the fate of heavy metals leached from the fabricated wastes in

aerobic and anaerobic bioreactor landfills,

o To explore the decomposition of lignocellulosic wastes in anaerobic and

aerobic bioreactor landfills, and

o To evaluate the loss of mass versus the loss of volume in aerobic and anaerobic

bioreactors for use in future settlement model development.

1.3 Research Approach

Four stainless steel lysimeters were constructed: two were operated aerobically and two were operated anaerobically. These lysimeters were designed and constructed as part of a previous research experiment (Sheridan, 2003). After operating the aerobic and anaerobic lysimeters for 1 and 2 years, respectively, one aerobic and one anaerobic lysimeter were dismantled. Waste samples were collected and characterized. The remaining aerobic and anaerobic lysimeters were kept in operation so that waste stabilization could be completely researched; the results of this extended operation will be presented elsewhere.

To compare leachate and gas quality between the aerobic and anaerobic bioreactors, two pairs of simulated landfill lysimeters containing fabricated wastes were operated as aerobic and anaerobic bioreactors. The fabricated wastes were loaded into the lysimeters, compacted, and mixed with water and seed (either anaerobic sludge or aerobic compost). Leachate generated by the lysimeters was collected and analyzed for leachate quality parameters. A mixture of collected leachate and deionized water was added back to the





4


lysimeters to compensate for the amount of leachate lost by leachate collection. The gas volume and composition were monitored using a gas totalizer and gas chromatography.

To explore the fate of heavy metals leached out of the fabricated wastes under aerobic and anaerobic conditions, heavy metal-containing wastes (e.g., CCA-treated wood, cathode-ray tube (CRT) glass and pieces of sheet metal) were mixed with the other fabricated wastes before loading into the lysimeters. After loading and compacting the fabricated waste, leachate generated by the lysimeter was collected and analyzed for copper, chromium, arsenic, lead, aluminum, zinc, manganese and iron. The change of heavy metal concentrations in the leachate over time was monitored. After the lysimeter work was completed, the wastes excavated from two columns were analyzed for heavy metals in order to compare heavy metal concentrations absorbed on solid waste to those released from the lysimeters through the leachate.

To explore the decomposition of lignocellulosic wastes in aerobic and anaerobic landfill environments, lignocellulosic wastes including paper and wood blocks were prepared. They were included in the fabricated waste and loaded in the lysimeters. After the lysimeter study was completed, lignocellulosic wastes were excavated and separated. Biochemical methane potential (BMP) assays were used to evaluate the degree of biodegradation of each lignocellulosic waste. In order to evaluate the impact of air addition on wood waste decomposition, cellulose, lignin and BMP of raw and excavated wood blocks were compared with respect to cellulose and lignin concentrations and BMP values.

To simulate landfill settlement in aerobic and anaerobic conditions as a function of waste mass loss, overburden pressure was applied to the stainless steel lysimeters using a





5


hydraulic cylinder and hand pump. To correlate mass loss and volume loss, a lab-scale experiment was designed where waste was decomposed in simulated landfills in the laboratory with both mass loss and volume loss being measured. A difficulty with using lab experiments to simulate landfill settlement is that it is hard to simulate true landfill conditions, especially, the large overburden pressure. In this research, the experiments included the application of overburden pressure to make the laboratory condition closer to the field conditions.

1.4 Outline of Dissertation

The dissertation is presented in six chapters. The current chapter presents the

problem statement, objectives and research approach. Chapter 2 presents the comparison of gas and leachate qualities between aerobic and aerobic simulated landfills. Chapter 3 presents the fate of heavy metals in aerobic and anaerobic simulated landfills. The evaluation of biodegradation of lignocellulosic materials is presented in chapter 4. Settlement behavior with waste decomposition is presented in chapter 5. Chapter 6 presents a summary, conclusions and recommendations for failure work. Background and other analytical procedures used for this research are presented in appendix A. Supplemental s are presented in appendix B. All other tables and s pertaining to leachate data and BMP are presented in appendix C.














CHAPTER 2
COMPARATIVE STUDIES OF LEACHATE AND GAS QUALITY OF AEROBIC
AND ANAEROBIC SIMULATED LANDFILL BIOREACTORS

2.1 Introduction

The operation of municipal solid waste (MSW) landfills as bioreactors; for the purpose of rapid landfill stabilization has historically been proposed as an anaerobic process. Conditions within the landfill are controlled to accelerate the activity of the anaerobic microorganisms responsible for waste decomposition. The addition of air has also been proposed as a method to enhance landfill stabilization (Stessel and Murphy, 1992), and recently this technique has gained more attention (Read et al., 200 1; Reinhart et al., 2002). In addition to an enhancement of waste decomposition that is more rapid than anaerobic operation, a major benefit often cited for air addition is the reduction in methane emissions relative to anaerobic landfills (Borglin et al., 2004). These studies also find that the overall strength of leachate (with respect to readily degradable carbon compounds and oxygen demand) is lower in aerobic systems, offering a potential advantage with respect to leachate treatment.

Research examining the relative differences in leachate quality between aerobic and anaerobic systems is very limited. Though the performance of aerobic bioreactor landfills has been simulated in several studies (Agdag and Sponza, 2004; Warith and Takata, 2004), these studies are often limited with respect to their ability to control several key parameters, and their lack of a complementary anaerobic system for comparison purposes. This chapter reports the results of research performed to examine the characteristics of



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leachate and landfill gas that result from aerobic and anaerobic operation of identical MSW streams. The experiments conducted involved a technique long employed in the study of landfills: waste-filled columns constructed and operated to simulate landfill processes, referred to here as lysimeters (Pohland, 1980). The columns were designed and operated to control several parameters not traditionally simulated in such experiments, such as temperature and overburden pressure. The objective was to compare leachate and landfill gas quality between each type of system so that similarities and differences can be better understood and to assist in future decision-making, design and operation efforts. Several complementary objectives were evaluated as part of this experiment and they are described in greater detail in Chapters 3 (fate of metals), 4 (comparison of decomposition) and 5 (comparison of settlement).

2.2 Material and Methods

Four lysimeters were used in this research, and each consisted of a stainless steel column and a carriage system component. The original design and construction of the lysimeters used for this research were described previously by Sheridan (2003). Two were operated aerobically (lysimeter I and 2) and two were operated anaerobically (lysimeter 3 and 4). Three parameters, temperature, air addition, and overburden pressure, were controlled in an effort to simulate actual aerobic or anaerobic bioreactor landfills.

2.2.1 General Description of the Lysimeter

A schematic of each lysimeter type is presented in Figure 2-1 (see Figure 13-1 for additional detail). The 6-ft stainless steel main body contained 5 front ports, 2 back ports and I valve at the bottom for leachate collection. The front ports were used for air addition (in the case of the aerobic lysimeters). The carriage system component was designed to support a hydraulic pressurizing unit installed at the top of each lysimeters





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for the application of an external load to the fabricated waste. The carriage system consisted of a hydraulic cylinder, carriage, steel shaft, and steel plate. A small port located on the top flange was used as a pathway for liquid addition. Perforations in the steel plate allowed added liquid to percolate into the waste (see Figure B-2 for a detail of the carriage system).

2.2.2 Temperature Control

The temperature at the center of a full-scale landfill usually remains constant

because the garbage and cover soil serve to insulate the system (McBean et al., 1995). In a laboratory environment, however, the heat produced by biologically degrading waste is not sufficient to maintain a temperature close to those normally encountered in a landfill. Thus, a temperature control system was designed and constructed (Figure B-3).

The temperature of each lysimeter was measured using a type T thermocouple wire (SRT2O1-160, Omega) fixed on the outside of each lysimeter. Two temperature controllers (MC240, Electrothermal) were utilized in series to maintain desired temperatures without extreme fluctuations. The lysimeters were insulated with 5-cmthick fiberglass and bubble insulation to minimize heat loss. Prior to operation, the lysimeters were filled with tap water and the temperature controllers were tested by measuring the temperature of the water.

The temperature of the aerobic lysimeters was maintained at a constant 55soC for the entire operating period. The anaerobic lysimeters were started at 350C and at day 400, the temperature was increased from 35'C to 55'C at a rate of 2'C per day. Although 550C is in the optimum range for thermophilic anaerobic waste decomposition (Rittmann and McCarty, 2001) and is often encountered in landfills (Watsoncraik et al., 1994;





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Townsend et al., 1996), 35C was used as the starting point because the anaerobic seed used was from a mesophilic digester.

2.2.3 Fabricated Waste Stream

The waste stream fabricated for this research was based on typical MSW

composition estimates previously reported for the U. S. and Florida (see Figure B-4 and B-5). For simplification purposes, several minor components, such as textiles and tires, were excluded from the fabricated waste stream. A greater portion of commingled paper was allotted as a substitute for those excluded materials. The relative amount of office paper, cardboard and newsprint in commingled paper (4.6 : 2.6 : 1) was again estimated from previous published data (FDEP, 2003 and USEPA, 2005). Figure 2-2 presents the fabricated waste stream composition used. Table 2-1 presents a description of each component, the source, and the method of sample preparation. Commercial grade dog food (Pedigree, USA) was used as the food waste portion of the fabricated waste stream. To support complementary research on the fate of certain heavy metals in aerobic and anaerobic landfill environment (chapter 3), a part of the wood waste fraction was comprised of CCA and a part of the glass fraction was comprised of leaded cathode ray tube (CRT) glass. Detailed waste components and their sources are presented in Table 21 and Figure 2-2.

Mixed fabricated waste samples were created and loaded into the columns as four distinct fractions to prevent waste component stratification in a particular place in the column, (composition of the fabricated waste fractions and their weight are summarized in appendix Q. Prior to loading, 6 inches (15.3 cm) of river rock was placed at the bottom of each lysimeter, and a geotextile was placed between the rock and waste. Each waste fraction was then loaded and compacted until it occupied 25 % of the depth of the





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lysimeter. Two liters of DI water were added along with the compaction of each waste fraction. After loading, 11 L of additional water was added from the top of each lysimeter. The goal of adding water was to bring the waste in each lysimeter, at the beginning of the experiment, to field capacity. A capacity of 58% was targeted as this was the field capacity measured for this waste under the initial compaction conditions of the lysimeter. The waste was compacted to a density of 30 lb/ft3 dy (480.6 kg/m3 dry).

2.2.4 Air Injection

Two computer-controlled pump drives (Model No. 7550-10, Cole-Parmer) were

used for air addition. Air was saturated and warmed prior to injection to keep moisture in the waste from evaporating. Air was injected on the ports located at the side of the aerobic lysimeters using a manifold from day 1 to day 164 and changed to the most bottom port from day 164 to the end of a test period. A flow rate of 70 mL/min was found to be suitable for control purposes and to maintain low exit gas oxygen concentrations. The flow rate was adjusted several times during the experiment when oxygen concentrations in the exit gas became less than 1% to maintain aerobic conditions.

2.2.5 Leachate and Gas Analysis

Leachate samples were collected on a weekly basis. Leachate was analyzed for sulfide and dissolved oxygen immediately; analysis for pH, alkalinity, and conductivity was carried out within one hour after collection. After this initial analysis, 15 mL of leachate was preserved with sulfuric acid and placed in acid-rinsed high-density polyethylene (HDPE) bottles for later analysis of chemical oxygen demand (COD), total organic carbon (TOC), volatile fatty acids (VFA) and ammonia. For metal analysis, 50 mL of leachate was preserved with concentrated nitric acid and stored at 4'C. The remaining leachate was recirculated back to the top of the lysimeters. Deionized water





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was added to make up for the amount of leachate used for analysis. Table 2-2 summarizes the parameters and methods used for each analysis.

Biogas samples generated from both the aerobic and anaerobic lysimeters were collected and analyzed for methane, carbon dioxide, and oxygen. For the aerobic columns, the gas volume was measured using a wet-tip gas meter. For the anaerobic lysimeters, the gas was gathered in 5-L and 10-L air-sampling bags, and the volume contained in the bags was measured using the water-gas replacement method (see Figure B-4). A LANTEC GEM 500 (SAIC, San Diego, CA) gas meter was used for gas analysis for both the aerobic and anaerobic lysimeters. Additionally, gas samples collected from the anaerobic lysimeters were analyzed for CH4 and CO2 using a gas chromatograph equipped with a GS-Carbon plot column (Agilent Technology, Palo Alto, CA) to confirm the measurements analyzed by LANTEC GEM500 gas meter.

2.2.6 Recovery of the Anaerobic Lysimeters

Since both of the anaerobic lysimeters (lys 3 and 4) remained in an acidic condition (pH < 6) for 500 days, 100 g of sodium bicarbonate was added as a buffer to the top of each lysimeter on day 300. The pH of the top part of the lysimeters changed to neutral, but the pH of leachate collected from the bottom port remained low (5 to 5.5). The pH of the leachate from lysimeter 4 increased to pH 7 from day 400. Since only minimum changes in leachate pH of the lysimeter 3 were observed after buffer addition, additional sodium bicarbonate was added to the bottom port rather than to the top of the lysimeters; a total of 1 OOg of sodium bicarbonate was added (20 g each were added on days 420, 453, 469, 532, and 555 again). Only a temporary increase in pH was observed after this addition. As a next step in increasing pH, lab air was injected into lysimeter 3 on day 627. Before air injection, the methane concentration of the lysimeter 3 was 35%, and the pH of





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leachate was 6.11. Lab air was injected with 70 mL/min from the bottom of the lysimeter for five days. The changes in pH and output gas qualities were monitored on a daily basis. The impact of this addition is discussed in the results section of this chapter.

2.2.7 Prediction of Waste Mass Loss

As described in the following sections in this chapter, the aerobic lysimeters more quickly stabilized the waste in comparison to the anaerobic lysimeters, and thus their period of operation was shorter (379 days vs. 741 days). In an effort to normalize the leachate measurements among the different columns, the biogas data, the leachate data and the initial content of the waste was used to estimate the percentage of waste decomposition for a column at any given time. The detailed procedure for this is presented in appendix A, but, in short, the cumulative volume of biogas measured at any given time (CH4 + C02 for anaerobic columns and CO2 for aerobic columns) was used to calculate the mass of initial waste degraded at that time. This was adjusted to account for the mass of organic carbon solubilized in the leachate. The mass of waste estimated to be degraded at a given time was divided by the estimated total potential mass loss in each column (this total potential mass loss was estimated from measured methane yields of the raw waste; see chapter 4 for details).

2.3 Results and Discussion

The data presented for the aerobic and anaerobic lysimeters in this dissertation represent operation periods of 379 and 741 days, respectively. At the end of each operation period, one each of the aerobic and anaerobic lysimeters was stopped and emptied. The remaining lysimeters were left operational (data are not reported here). Values of all leachate parameters analyzed for this research are presented in Table C- I





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through C-i 1 in appendix C. These include the raw data and graphs of the leachate parameters.

2.3.1 pH

Figure 2-3 depicts the change in pH over the course of the experiment. Both the

aerobic and anaerobic lysimeters remained in acidic condition during the beginning of the experiment. The period of time required to stabilize the pH for the aerobic and anaerobic lysimeters was 200 and 600 days, respectively. Average pH measurements of approximately 8.9 (aerobic) and 7.1 (anaerobic) were observed at the end of the experiment.

Two phases (acidic and alkaline or methane phase) of the pH of the aerobic and

anaerobic lysimeters were observed during a test period. The low pH occurring during the initial phase of the research was attributed to a build up of organic acid concentrations and the related microbial activities. Once the organic waste decomposition process began, the biodegradable fraction of waste was converted into organic acids by various biological reactions, and the accumulation of the organic acids lowered the pH. For the aerobic lysimeters, air was injected through four front ports of the lysimeter using manifolds. The pH was low (< 6) for the first 150 days, and high VFA and alkalinity concentrations indicated that anaerobic conditions were predominant, suggesting that air was not evenly distributed through the manifolds. An increase in the pH of the aerobic lysimeters was observed after air was injected into the only bottom port. Typically, the pH of the system increases to neutral conditions as the organic acids are consumed by methanogenic bacteria. A large amount Of CO2 production in an unbalanced ecosystem may also contribute to lowering the pH as well. High concentrations of VFA and alkalinity were measured in the anaerobic lysimeter leachate during the initial acid phase.





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The pH of the aerobic lysimeters measured in the latter half of the experiment (9.0) was more alkaline than the pH measured from the anaerobic lysimeters (7.2). According to other lysimeter studies, higher pH was observed from the aerobic lysimeters in comparison with that of the anaerobic lysimeter. The range of pH of aerobic lysimeters has been reported as 7 9 (Stessel and Murphy, 1992; O'Keefe and Chynoweth, 2000; Agdag and Sponza, 2004). Summerfelt et al. (2003) also observed an increase of pH when air was injected into their aquaculture system. They reported that this increase was because of CO2 stripping by air; a decrease in CO2 leads to a decrease of carbonic acid (H2CO3) and bicarbonate concentrations (HCO3) consuming H ions. These relationships can be described by carbonate systems as follows: CO2 gas <- H2CO3 (1)

H2CO3 --* HCO3 + H+ (2)

HCO3" E--) CO3 + H+ (3)

They additionally concluded that, because the dehydration of carbon acid is rate-limiting, pH may not increase instantaneously.

2.3.2 Organic Carbon Concentration

Figure 2-4 depicts the change of COD concentrations for the lysimeters versus

time. The initial average COD concentrations in the leachate of the aerobic and anaerobic lysimeters were 36,000 mg/L and 66,000 mg/L, respectively. The COD values for the aerobic lysimeters increased up to greater than 84,000 mg/L and decreased rapidly after pH was stabilized. Although one of the aerobic lysimeters showed high COD (70,000 mg/L) at day 50, the overall COD concentrations of the aerobic lysimeters were lower than values in the anaerobic lysimeters. Similar trends occurred for COD as were observed for BOD5 (Figure 2-5). The BOD values of the aerobic lysimeters decreased





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rapidly down to below 100 mg/L from day 200 while BOD values of the anaerobic lysimeters decreased relatively slowly.

The primary contributor to high COD or BOD concentrations in landfill leachate is volatile fatty acids (McBean et al., 1995). Figure 2-6 (a) depicts the changes in acetic acid, one of the major volatile fatty acids (VFA), as a function of mass loss. Acetic acid is used as a substrate by methanogenic bacteria and contributes to the formation of an acidic environment when they are unbalanced with the growth of methanogenic bacteria. For these reasons, VFA concentrations are used as an indicator to assess landfill conditions (USEPA, 2004). For example, the decrease in acetic acid concentration in lysimeter 3 and

4 corresponds to the point when the pH began to rise. In the aerobic lysimeters, high concentrations of acetic acid were noted during the first phase of the experiment due to improper air distribution as discussed earlier. Acetic acid in leachate from the aerobic lysimeters was degraded to less 1 mg/L by day 200, which corresponds with the time required to deplete COD and BOD.

Among different types of short carbon chain fatty acids, acetic, propionic and

butyric acids are known as major VFAs that are involved in biodegradation processes in anaerobic conditions. Production and degradation of these major VFAs in selected aerobic and anaerobic lysimeters are presented in Figure 2-6 (b). In both aerobic and anaerobic lysimeters, the concentration of VFAs was mainly: acetic acids > butyric acids > propionic acids. These results are similar to those found by Parawira et al (2004). They also explained that high butyric acids were mainly attributed to high carbohydrates in waste. Under the same condition, the degradation of VFAs in anaerobic condition was found to be in the following order: butyric acids > acetic acids > propionic acids. Wang et





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al. (1999) explained that various enzymatic reactions in microorganisms dictate a greater decreasing rate of butyric acids than that of other VFAs. However, more biosynthetic processes are involved in butyric acid production than acetic acid due to longer carbon chains. In aerobic lysimeters, all three major VFAs were depleted together like other bulk organic carbon.

The ratio of BOD5 to COD is often used to assess the biodegradability of the organic matter in leachate, and thus to assess the degree of landfill stabilization. In old stabilized landfills, the BOD5/COD ratio is below 0.10 (Kjeldsen et al, 2002). A low BOD5/COD suggests that a leachate is low in biodegradable organic carbon and relatively high in hard-to-biodegrade organic compounds such as humic compounds. In this research, low BOD5/COD ratios were observed with the aerobic lysimeters after day 200 (0.04 on average) (Figure 2-7). Relatively high BOD5/COD ratios were exhibited from the anaerobic lysimeters (0.36 on average). These values fall into the range of average BOD5/COD ratios proposed by Kjeldsen et al. (2002) for the acid phase (0.58) and the methanogenic phase (0.06).

2.3.3 Nitrogen

Figure 2-8 shows the changes in ammonia-nitrogen in the lysimeter during the course of experiment. Ammonia concentrations from the aerobic lysimeters remained relatively constant, showing a general increase during the course of the experiment. In a different fashion, ammonia concentrations in the anaerobic leachate increased dramatically at a point corresponding to an increase in system pH. Ammonia concentrations in the anaerobic lysimeters increased to concentration in the range of 1000-1600 mg/L. These values then dropped to 800-1000 mg/L and stabilized. Small increases of ammonia concentrations in leachate of aerobic column were observed after





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day 180, but they were still approximately 4 times lower than values of anaerobic lysimeters. The trend of increases in ammonia concentrations also can be found in operating bioreactor landfills (Reinhart and AlYousfi, 1996).

Since ammonia is generally produced from the deamination process of amino

acids (a monomer of proteins), elevated ammonia concentrations may be associated with protein decomposition. Cali et al. (2005) reported that an active methanogenic bacteria community increased the ammonia concentration. Several researchers have proposed that the enhancement of waste decomposition and leachate recirculation in anaerobic bioreactor landfills results in increased ammonia concentration (Reinhart and Al-Yousfi, 1996; Berge et al., 2005).

2.3.4 Dissolved Solids Content

Figure 2-9 shows the change in total dissolved solids (TDS) in the aerobic and anaerobic lysimeters through the course of the experiment. For both aerobic and anaerobic lysimeters, like the change in other organic matter, TDS concentrations were lower as the pH was stabilized. TDS of the aerobic lysimeters were rapidly stabilized approximately 8 to 10 g/L after day 200. TDS of the anaerobic lysimeters were still greater than that of the aerobic lysimeters but TDS of lysimeter 4 fell below 20 g/L on day 700. Typical TDS concentration in landfills is within the range of 2 to 60 g/L (Kjeldsen, 2002).

Figure 2-10 depicts the change in alkalinity in the aerobic and anaerobic

lysimeters through the course of the experiment. For the aerobic lysimeters, the alkalinity increased to 16,000 mg/L as CaCO3 in the lysimeter 1, but another lysimeter showed low alkalinity which was below 2,000 mg/L as CaCO3 but it increased 8,000 mg/L as CaCO3 again. The alkalinity was lowered below 2000 mg/L as CaCO3 for both aerobic





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lysimeters after the pH was stabilized at alkaline condition. For the anaerobic lysimeters, relatively high alkalinity was maintained over the entire test period. Generally, alkalinity could be generated by CO2 accumulation and ammonification (Fannin, 1987). It is also consumed by nitrification (Gujer and Jenkins, 1974).

2.3.5 Oxidation Reduction Conditions

Figure 2-11 depicts the change of sulfide concentrations in the aerobic and

anaerobic lysimeters over a period of time. Little changes of sulfide were observed during the acid phase of both aerobic and anaerobic lysimeters. However, rapid increases in sulfides along with an increase of pH were exhibited from the aerobic lysimeters and lysimeter 4. The sulfide level of lysimeter 1 was lowered on day 2 10 again, but increased up to 2,600 itg/l, during the alkaline phase. The trends of change in sulfide concentration of lysimeter 2 exhibited are similar to that of lysimeter 1. The highest sulfide level found in lysimeter 2 was 1,200 p~g/L. In lysimeter 4, sulfide concentrations increased as the pH increased.

It is notable that great concentrations of sulfide were found in the system where air injection had been taking place. It is hard to understand how sulfide could be presented in an aerobic environment. In comparison with sulfate concentrations, sulfide was formed by sulfate reduction (Figure 2-12). However, Figure 2- 10 shows that high dissolved oxygen concentration was also found in the same condition. Snoeyink and Jenkins (1980) reported that sulfide could be detected under aerobic conditions. They explained that this phenomenon was caused by a non-equilibrium situation for the reaction between oxygen and sulfide. Therefore, it can be concluded that sulfide can be found before it oxidizes for a second time by dissolved oxygen. This result indicates that anaerobic zones were





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presented in the aerobic lysimeters producing sulfide. Relatively high ammonia concentrations found from the aerobic lysimeters (Figure 2-9) also indicated the presence of anaerobic zones in the aerobic lysimeters.

2.3.7 Gas Quality

Biogas emitted from the aerobic and anaerobic lysimeters was measured for 02, CO2 and CH4. Figures 2-13 and 2-14 depict the changes in gas concentrations of aerobic and anaerobic lysimeters over a period of time. The initial air injection rate of the aerobic lysimeters was 70 mL/min. The air injection rate was regulated by changes of oxygen levels within the range of 70 to 120 mL/min. High CO2 concentrations were observed from aerobic lysimeters during the first 50 days, but decreased to lower than 20%. The concentrations of CH4 and CO2 of the anaerobic lysimeters changed during the acidic phase, but stabilized to approximately 60% CH4 and 40% CO2 during the methane phase.

Overall, a total of 40,100 liters (1,400 ft3) of gas was injected into each aerobic lysimeters for a test period, and 45% and 43% of the oxygen included in the air added was converted into CO2 in lysimeter 1 and 2, respectively. In the anaerobic lysimeters, 500 and 1,600 liters of biogas (CO2 and CH4) were produced from lysimeters 3 and 4. Most of the gas generated was mainly concentrated on the methanogenic phase in anaerobic lysimeters while a relatively steady gas generation was exhibited over time in aerobic lysimeters as summarized in Figure 2-15.

Lab air was added to recover the lysimeter 3, which had remained in acidic

condition (pH <6) for 600 days. Figure 2-16 depicts the change in gas concentrations, gas generation rate and pH during air injection. The pH was adjusted to 7.1 at day 4, and air injection was stopped on day 5. After oxygen was depleted in the lysimeter, methane concentrations substantially increased along with biogas generation rate and reached





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above 50% at the 9th day. The biogas generation rate of lysimeter 3, after air injection, was 2.3 L/day on average for 10 days. During the rest of test period, the pH of the lysimeter 3 went down to 6.5, but further decrease was not observed. In comparison with the conditions of lysimeter 3 before air addition, the amount of biogas produced was substantially increased and high percentage of methane (> 55%) was maintained (Figure 2-16).

2.4 Discussion

2.4.1 Differences between Aerobic and Anaerobic Lysimeters

The largest differences between the aerobic and anaerobic lysimeters can be found from the enhancement of waste biodegradation. Based on the leachate quality results of this study, a period required for the aerobic lysimeters to decompose 90% of BOD was 160 days in the aerobic lysimeters while more than 700 days were required for the lysimeter 4.

Other differences between the two systems were the methane concentrations

contained in exit gas. Air addition to the aerobic lysimeters lowered CH4 concentrations in the exit gas dramatically. Though a small amount of methane was found in the exit gas of the aerobic lysimeter, it was less than 1% of the CO2 gas generated.

It is noted that pH had a relatively low impact on waste decomposition in the aerobic lysimeter. Though the pH of the aerobic lysimeters was acidic, settlement consistently occurred (see Chapter 5). It was probable that the acidic condition was localized only on the bottom part, where air was not supplied properly.

Tables 2-3 and 2-4 present the initial and final characteristics of the aerobic and anaerobic lysimeters. The data presented in Table 2-4 for the anaerobic lysimeters indicate that these systems were not stabilized yet. Water loss from the aerobic lysimeters





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was calculated using the water carrying capacity of the exit gas assuming that the gas was 100% saturated with water vapor. The overall performance of the aerobic lysimeters with respect to waste decomposition and leachate quality was substantially greater than those of the anaerobic lysimeters. However, the concentrations of sodium in the aerobic lysimeters were still too high to meet drinking water standards. The final concentration of ammonia in the aerobic lysimeters was also substantially higher than the criteria value of ambient water (0.897 at 30'C and pH 8.0) (USEPA, 1999). Although large quantities of waste were decomposed, leachate of the anaerobic lysimeters still contained high concentrations of organics, ammonia and anions (Table 2-4). Leachate generated would be used for recirculation, but excessive volume of leachate must be treated at an on-site or off-site wastewater treatment plant.

2.4.2 The Comparison of Leachate Parameters with Other Studies

Tables 2-5 and 2-6 summarize the comparison of leachate constituent

concentrations from this study with those from other studies. For aerobic landfill studies, the maximum concentrations of COD, BOD and ammonia in this study appeared to be greater than those of other studies, but they were in a similar range overall. The pH of the aerobic lysimeters of this study was, however, greater than other studies. As previously discussed, this would be because of the relocation of carbonic acids, bicarbonate, and carbonate due to CO2 removal by air stripping (Summerfelt, 2003). The high pH of the aerobic lysimeters implies that great concentrations of carbonate ions were dissolved due to high partial pressure Of C02, and these carbonate ions might consume more H' ions when CO2 was removed. If alkalinity data of other studies were available, it would be





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clear to describe the differences between this and other studies by comparing the concentrations of carbonate ions.

2.4.3 Implications for Full-scale Application

Unlike the lab-scale simulated landfill, it is extremely difficult to aerate an entire large-scale landfill. Highly compacted wastes make it difficult for an air stream to penetrate into the recesses of a landfill. Moreover, leachate characteristics resulted from air addition may be variable. As the analytical results have shown, leachate characteristics of lysimeters I and 2 were different, despite starting with the same waste stream and the same operational condition. The leachate characteristics of lysimeter 1 were similar to those of the anaerobic lysimeters during the first 180 days showing great concentration of organic matter despite air addition. This was because of the large anaerobic zones formed at the bottom of the lysimeter by improper air addition to the bottom.

Aerobic zones can be formed around air injection wells but anaerobic zones may still be present in the same landfill. However, coexistence of the aerobic and anaerobic zones can be used for recovery of acid-stuck 'sour' landfills. In this research, the air addition was conducted under the hypothesis that environments formed by aeration for a short period can be favorable to anaerobic microorganisms. A great amount of VFAs, which caused acidic conditions, may be rapidly consumed by aerobes living in a relatively wide pH range. Conversion of carbonic acid (H2CO3I) to C02 caused by air stripping may increase the pH. With air addition with low flow rate, the anaerobic zones may be protected from oxygen intrusion because oxygen may be depleted by the respiration of aerobes. An additional technical strategy would be to add buffer such as lime along with air addition. Buffer added may increase the alkalinity concentration.





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Without high alkalinity, the pH of the landfill may decrease again when air addition is stopped. This could happen when methanogenic bacterial population was not enough to adapt to the new condition.

As Reinhart et al. (2002) pointed out, the reduction of leachate volume due to air stripping could be one of the advantages of the aerobic landfills. In this research, a total of 31m m3 of air was added during a test period (1 year). Comparing the volume of water initially added with final leachate volume, approximately 2 1% of leachate volume was reduced. Reduced volume of leachate implies that the operation of aerobic landfills can be economical in terms of saving the cost for the leachate treatment.

2.4.4. Limitations

Since CI-L gas is one of the gases causing global warming, CH4 reduction can be one of the advantages of the aerobic lysimeters. However, landfill gas released without a' flare system could be adverse to the environment. Berge et al. (2005) and Reinhart et al. (2002) pointed out that various kinds of volatile organic compounds (VOC) and nitrous oxide, a more potent greenhouse gas than methane, can be emitted without the flare system. Future research is required to identify the gas constituents and develop the filtering system as an alternative.

As previously mentioned, an extra monitoring job may be required to check

moisture content and gas contents around the gas injection well. Certain ratios of methane and oxygen can be flammable according to Coward and Jones (1952) and Liao et al. (2005). While air was added, CH4 concentrations were low, but the unpredictable changes in 02 and CH4 concentrations were observed from the aerobic landfill (Read et al, 2001) and high concentrations of CH4 and 02 could coexist when air addition starts (Lee et al, 2002).





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2.5 Conclusions

In this research, the gas and leachate quality from aerobic and anaerobic

simulated bioreactor landfills were compared. Waste streams referenced from EPA and FDEP were loaded into 4 stainless steel lysimeters with a density of 3500 kPa. All lysimeters were prepared with the same conditions, and two of them were assigned for aerobic and two for anaerobic bioreactor landfill simulations. Leachate and gas generated from the lysimeters were analyzed for chosen parameters to make comparisons between aerobic and anaerobic landfills.

Leachate analysis results indicated that organic compounds as measured by COD, TOC, BOD and VFAs in the aerobic lysimeters were degraded more rapidly than those in anaerobic lysimeters. Except for the acidic phase, the pH of the aerobic lysimeters rapidly increased and was stabilized around pH 9.0 while anaerobic lysimeters had remained in acidic phase for more than 400 days, and stabilized exhibiting pH 7.3. The concentrations of ammonia in anaerobic lysimeters increased along with an increase of pH. Ammonia concentrations in aerobic lysimeters varied little over time, but ammonia levels were significantly lower than those of anaerobic lysimeters. Sulfide results imply that both aerobic and anaerobic zones were coexisting in aerobic lysimeters. This may be caused by the limit of oxygen distribution in the lysimeters because of high density and low hydraulic conductivity of wastes under overburden pressure.






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Table 2- 1. MSW components
Waste components Sources Processing for size reduction
Office paper Mixed scrap paper purchased at Grind with a paper shredder
office supply store
Cardboard Mixed corrugated boxes Scissors and razor blade
Newspaper Local newspaper Grind with a paper shredder
Plastics PET bottles collected from a Scissors
plastics recycler
Foodwase Comerial og oodGrind with a coffee grinder (less
Foodwase Comerial og oodthan 1/32") Southern yellow pine (SYP) Home improvement store Cut with band-saw (2" x 2")
CCA-treated wood Home improvement store Gather saw dust after drilling

Galvanized steel Home improvement store Cut with metal cutter (1/2" x
__________________________ __________________________1/2")
Aluminum Home improvement store Cut with metal cutter (1/2" x
1/2")
Cathode-Ray Tube(CRT) glass CRT monitors Crush with a hammer (1/4-1/8")
Mixed cullet Mixed container glass Crush with a hammer (1 /4 1/8")






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Table 2-2. Parameters and methods for analy sis.
Parameters Method
Alkalinity Standard Method 2320B
Ammonia Standard Method 4500-D
BOD Standard Method 5210B
COD HACH 2720
Conductivity Standard Method 2510
pH Standard Method 4500-H'
TOC EPA SW846, Method 9060
Sulfide HACH 8131
Sulfate, Floride and Chloride EPA SW846, Method 9056
Sodium EPA SW846, Method 9060A
VFA VFA analysis method using GC
(Innocente et al., 2000)






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Table 2-3. Comparison of initial and final characteristics of the aerobic lysimeters Initial Final (1 year)
Lys 1 Lys 2 Lys 1 Lys 2
Water quantity (mL) 19,000 19,000 15,137* 15,582 (15,056*
Dry waste quantity (g) 12,784 12,784 8,389* 8,740 (8,715*)
pH 5.7 5.7 8.5 8.5
COD (mg/L) 20,000 28,000 3,400 4,700
BOD 13,000 16,000 200 30
TOC 6,000 7,000 2,600 2,200
Ammonia 70 40 500 250
Fluoride 80 30 0 0
Chloride 200 130 1,200 1,700
Sodium 80 140 800 900
* predicted






28


Table 2-4. Comparison of initial and final characteristics of the anaerobic lysimeters Initial Final (2 years)
Lys 3 Lys 4 Lys 3 Lys 4
Water quantity (mL) 19,000 19,000 18,844* 18,833 (18,704*)
Dry waste quantity (g) 12,784 12,784 11,290 9,258 (8,997*)
pH 4.5 4.9 6.5 7.4
COD (mg/L) 65,000 67,000 42,000 24,000
BOD 48,000 62,000 14,000 6,500
TOC 26,000 27,000 12,000 5,600
Ammonia 120 100 1,000 800
Fluoride 1,500 1,400 460 200
Chloride 1,450 1,400 670 500
Sodium 2,000 2,000 4,800 3,800
* predicted






29


Table 2-5. Comparison of leachate parameters with other aerobic landfill studies Parameters
(mg/L Compost Lysimeter Lysimeter Lysimeter Lysimeter This tud
except for study study I b study 2c study 3d study 4e This study
pH)
Air flow 20L/min 38L/min for
20L/mmn
rate 30min. at 8.4 70(once per a 12-h 1300L/min 20mL/min 120mL/min
week) intervals
COD 2434-31812 861-22026 130-23000 500-5000 2-1000 3400-47000
BOD5 8-11571 100-10000 10-2000 30-45000
Ammonia 98-558 260-630 7-400 2-100 40-700
TDS 3300-11400 700-7700
pH 7.1-8.2 5.17-7.98 5.24-7.5 7-9 7.8 4.5-9.1
aKrogmann and Woyczechowski, 2000; bAgada and Sponza, 2004; CWarith and Takata, 2004; dStessel and Murphy, 1992; eBorglin et al., 2004





Table 2-6. Comparison of leachate parameters with other anaerobic landfill studies Parameters Bioreactor
(mg/L Conventional Conventional Bioreactor
landfill lab This study
except for landfill 1a landfill 2b landfills scaled
pH) scale
pH)
COD 140-152000 1000-40000 20-17000 100-88000 2000-80000
BOD5 20-57000 50-25000 0-10000 6600-60000
TOC 30-29000 7000-19000
Ammonia 50-2200 50-1500 76-1850 100-1600
TDS 2000-60000 2000-25000 18000-50000
pH 4.5-9 3-7.5 5.4-8.6 4-7.5 4.5-7.5
Ca 10-7200 300-4000 20-4000
a Kjeldsen et al., 2002; bPokhrel, 2004; CReinhart and AI-Yousfi, 1996; dPohland and Kim, 1999






30


Back Front




Leachate Carriage
I e[ injection system


To the 4-gas collection system






Main
body











Air injection Leachate collection Figure 2-1 Schematic of the lysimeter






31




SYP CRT glass Mixed cullet

CCA treated wood5% 1 % 6%
I %
Alumium Paper office paper
4% 27%
Galvanized steel
4%




15%
paper newsprint
6%



Food waste Paper cardboard
15% 16%




Figure 2-2. The composition of fabricated municipal solid waste for this research.






32









10
AEROBIC
9 8 7

6
6 O-0 lys 1
5 000 O 0 0 0-- lys 2

4 I I
0 100 200 300 400
8
ANAEROBIC
7


6


5 lys 3
-0- lys 4


0 100 200 300 400 500 600 700 800

Days Figure 2-3. Comparison of pH between aerobic and anaerobic lysimeters versus time






33





100000
AEROBIC
-0- lysimeter 1
-0-- lysimeter 2 80000




60000



0
S40000




200000
0 100 200 300 400

100000
ANAEROBIC +-- lysimeter 3
-0- lysimeter 4 80000




60000



0
) 40000




20000




0
0 200 400 600 800

Days Figure 2-4. Changes in COD of aerobic and anaerobic lysimeters versus time






34





80000
AEROBIC
----- lys 1
-0- lys2


60000





40000
O
0



20000
O0
600
o
.0

0 a
0 100 200 300 400

80000
ANAEROBIC
-0- lys 3 O-0 lys4





0
60000








20000

O OO


0.
Cbo


0 100 200 300 400 500 600 700 800

Days Figure 2-5. Changes in BOD of aerobic and anaerobic lysimeters versus time






35







30000
AEROBIC
lys 1

25000 -- lys 2



20000 15000 10000




5000



0'
0 100 200 300 400
30000

-0- lys 3 ANAEROBIC
-0- lys 4
25000



S20000



15000



S10000



5000




0 200 400 600 800

Days
(A)






36





30000
lys 1 (aerobic) Acetic acids
0-- Propionic acids 25000 -- V- Butyric acids


20000

-I
15000


S10000
O/
> //

5000

o- ...... 00o..... ...... oo ....... = : ==..
0 1 o o
0 50 100 150 200 250 300 350 400
30000
-0- Acetic acids lys 4 (anaerobic)
0... Propionic acids
25000 -vT- Butyric acids


20000
-o

15000


10000
so o o 0 o .


0



0 200 400 600 800

Days
(B)
Figure 2-6. Changes in VFAs of aerobic and anaerobic lysimeters versus time (A) acetic
acid only and (B) acetic acid, propionic acid and butyric acid





37





1.0
Lys 1
S 0 Lys 2
8 A A v Lys 3
0.8 A Lys 4



0.6 O
0.6 Oiw A m A
0 0
U O


OA
0 V
0.4 V
020 0 0

0
0.2



0.0 8
0 200 400 600 800

Days
Figure 2-7. Changes in the ratio BOD/COD of the aerobic and anaerobic lysimeters over
time






38









600

lys 1 AEROBIC
--0- lys 2
500



400



300



200



100



0
0 100 200 300 400

1800
-- lys 3 ANAEROBIC
1600 lys4

1400 1200

1000

o 800 < 600

400 200

0
0 200 400 600 800
Days Figure 2-8. Changes in ammonia concentrations versus time






39





60
AEROBIC
-0- lys 1 0-- lys2 50



40



30



20



10
0.

0 0 0
0..... d%
0
0 100 200 300 400

ANAEROBIC
60 lys 3
lys4 50


40


30
HA

20


10


0 I I I
0 100 200 300 400 500 600 700 800

Days Figure 2-9. Changes in TDS of the aerobic and anaerobic lysimeters versus time





40




20000
AEROBIC Lys
0 Lys 2

S15000
O
0



b 10000

0


5000


o . 0 0 0 o o o 0 .0 0 0
0 0
.0 -- -0 0 ~ 0~cxc00 OO 00

0 100 200 300 400
20000
ANAEROBIC


S15000 A



-~AA
-" 10000




< 5000

-A- Lys 3
--A Lys4

0
0 200 400 600 800
Days

Figure 2-10. Changes in alkalinity of the aerobic and anaerobic lysimeters versus time






41





4000
Lys 1 AEROBIC
0-O Lys2

3000




a 2000
-U



1000 0 0
1000 0 0 0.000

0O
0

0 0
0 100 200 300 400

4000

-- Lys 3 ANAEROBIC
OLys 4 0.- Lys4

3000


E o

S2000



9
1000ooo


~ ~..........6 --0
0 200 400 600 800

Days Figure 2-11. Changes in sulfide and pH versus time







42






100 4--- 000 10
Lysimeter 1


80 8
3000 x


T 6b
I 60 x 6 E
X2

-2000
EXXX 0001
40 xx / -4

20 I i 2
I I1 1000


I \
000



0. 60 -0
0 100 200 300 400

Days
180 1400 10

160 Lysimeter 2 1200
160 1200

140 -x 8
x 1000
120 x x
x 6





o :" _-'_ 80
100 800




2 4006-*
60 y6
\ -400 t
40 -2 / 200
20 1

0 -0 0
0 100 200 300 400

Days


Figure 2-12. The changes in sulfate and sulfide versus time in the presence of dissolved
oxygen






43





40 140
Lys 1 CH4
C02 120
i 02
0I Air injection
30 ..
"-' 7 I r2 --I100 '
,_. I I Iill "'1i I80

-'.~ ,,
L2 -h - -
60

[..... ... 40

10 .

\ .. '.. .20
k11 1:1I[;/ I I"2
I J /F I I I

0 I 0 I /
0 100 200 300 400

Days



Figure 2-13. The changes in air injection rate and gas concentrations of aerobic lysimeter






44





100
Lys 4
CH4 ........ C02
80 ---02


0


.4.



Ci2
20

20


I\
0

0 200 400 600 800

Days Figure 2-14. Changes in gas concentrations of anaerobic lysimeter 4






45






5000
lys 1 .................. ly s 2
lys 3 S 000 - lys 4

O


3000




0
2000




1000

.. .. . .

0 I I I
0*
0 200 400 600 800

Days Figure 2-15. Cumulative biogas vs. days in aerobic and anaerobic lysimeters






46






60

Air injection started Air injection ended
50



40 1

.6
.30 0
o 0

.0
20 '0.CH

0
10 1
I 0"v
090

3.0




*7.0
2.0

0
0 '0 1.5.o 6.5

00
1.0


6.0 0.5
0 5 10 15 20 25 30

Days
Figure 2-16. Changes in gas concentrations, pH and gas generation rate after air injection
into lysimeter 3














CHAPTER 3
THE FATE OF HEAVY METALS IN SIMULATED LANDFILL BIOREACTORS
UNDER AEROBIC AND ANAEROBIC CONDITIONS

3.1 Introduction

An issue of current debate in the solid waste community is the fate of heavy metals disposed in MSW landfills (SWANA, 2003). Heavy metals may be present as a result of industrial residuals, but more importantly for MSW landfills, they result from manufactured products. Examples include lead from electronic devices and copper, chromate and arsenic from treated wood. This debate has taken on more immediate concern as several US states have banned certain wastes containing heavy metals (e.g., leaded cathode ray tubes) from disposal in landfills (SWANA, 2003). These bans are in part a result of fears regarding the fate of the disposal of metals in landfills.

For the most part, heavy metals have been thought to be relatively well contained in typical anaerobic landfills. According to Kjeldsen et al. (2002), the amount of heavy metals dissolved and contained in leachate is very low relative to those present in the waste. Most metals are thought to be released during the initial stage of landfill decomposition as a result of the lower pH. Once a landfill enters the methanogenic phase, heavy metal concentrations in leachate dramatically decrease, and in many cases, their levels decrease to lower than drinking water standards (Kjeldsen et al., 2002). Bioreactor landfills are becoming a more common method of managing MSW, and the impact of these systems on metal leachability should be evaluated. Since bioreactor landfills involve exposing a much larger percentage of waste to moisture, the total mass



47





48


of metals released might be expected to be high relative to dry landfills. On the other hand, given that bioreactor landfills recirculate leachate to the landfill and that traditional bioreactors promote anaerobic waste decomposition (and thus enhance metal removal by the mechanisms described previously), the impact to the environment may be limited.

An alternative bioreactor landfill technique is to add air in addition to moisture.

Taking into consideration that the leaching behavior of heavy metals is mainly controlled by redox, pH and the presence of ligands (Benjamin, 2002), it may be that the fate of heavy metals in aerobic systems will differ from anaerobic systems. The long term fate of heavy metals in aerated landfills is a question yet to be satisfactorily addressed.

In this research, the fate of heavy metals in simulated aerobic and anaerobic

bioreactor landfills was studied. Four stainless-steel lysimeters were used, two each for aerobic and anaerobic conditions. Heavy metal-containing wastes were included in the waste stream added to each lysimeter. Leachate collected from each lysimeter was analyzed for heavy metals over time. In order to evaluate the heavy metals adsorbed in solid wastes, waste samples were removed from two of the lysimeters at the end of the experimental period and analyzed for heavy metal content.

3.2 Materials and Methods

A detailed description of the lysimeters was presented in chapter 2. The methods as described here focus on the metal-containing components in the fabricated waste and on metal analysis in the leachate and waste samples. 3.2.1 Heavy Metal Sources in Synthetic Waste

Several MSW components were chosen as sources of heavy metals. Each

component, its corresponding heavy metals and the percentage of each component are presented in Table 3 -1. Aluminum and galvanized steel sheets were purchased from a





49


local hardware store and cut into 1.5 cm x 1.5 cm square. Galvanized steel served as a source of both Fe and Zn. CCA-treated wood was used as a source of Cr, Cu and As. Crushed cathode ray tube (CRT) monitor glass was used as a Pb source. Total Cu, Cr, and As concentrations were 2350 50, 2890 56, and 1330 10 mg/kg, respectively. Crushed CRT monitor glass used in this research was a mixture of the funnel sections of 30 CRT color monitors. Jang and Townsend (2003) reported that 413 mg/L of Pb leached from CRT funnel glass using the toxicity characteristics leaching procedure (TCLP).

3.2.2 Sampling Methods

Leachate samples were collected weekly via a sampling port located at the bottom of each lysimeter. A portion of the leachate collected was used for analysis of general water quality parameters and the remainder was injected back into the lysimeters. A 50 mL aliquot was preserved with concentrated nitric acid and used for heavy metal analysis.

Lysimeter studies were conducted for 379 and 741 days for aerobic and anaerobic lysimeters, respectively. After the lysimeter studies were completed, solid wastes were removed from single aerobic and anaerobic lysimeter and analyzed for heavy metals. The samples were divided by depth into 4 fractions. Details about these fractions are summarized in Table A-2 in appendix A. Each fraction was then separated into 5 categories which include office paper, cardboard, newspaper, wood blocks and plastics. The separated samples were ground using an Urschell mill (Fritsch, German).

3.2.3 Analytical Methods

Leachate samples were digested with nitric and hydrochloric acids following EPA method 3050B and 30 1 OA for solid and liquid digestion, respectively (USEPA, 2003). Approximately 2 g of the ground samples were digested using nitric acid and 30%





50


hydrogen peroxide and then analyzed for heavy metals using ICP-AES following US EPA, SW-846 Method 60 1OB (USEPA, 2003). Digested samples were filtered using ashfree cellulose filters and analyzed for heavy metals and cations using Inductively Coupled Plasma-Atomic Emission Spectrometry (ICP-AES) (Thermo Electronics, USA). Leachate samples preserved with concentrated nitric acid were analyzed for a total of 8 metals (As, Cu, Cr, Mn, Zn, Pb, Fe and Al).

3.3 Results and Discussions

3.3.1 Changes in Metal Concentrations versus Time and the Percentage of Mass
Loss

The following section presents the results (for each metal) of the aerobic and

anaerobic lysimeters as separated plots. The experimental time period for the aerobic and anaerobic lysimeters differed. Because of their time scale difference, the cumulatitive mass of metal leaching was plotted for all lysimeters as a function of waste mass loss. The estimation of mass loss is described in appendix A. The total amounts of leachate produced and used for the analysis are summarized in Table 3-3.

3.3.1.1 Aluminum

The changes in Al concentration in aerobic and anaerobic conditions over a period of time are depicted in Figure 3-1. High Al concentrations were observed from both aerobic and anaerobic lysimeters (18 and 20mg/L) for the first 10 to 20 days. Whereas the Al concentrations of anaerobic lysimeters dramatically decreased to below 0.5mg/L within 100 days, great changes in Al concentrations were not observed from the aerobic lysimeters. The changes in Al concentrations in aerobic lysimeters were mainly controlled by pH; high concentrations of Al were observed from both lysimeters 1 and 2 at pH < 6 and pH > 8, and lowest Al concentrations (< 1 mg/L) were observed at 6 < pH





51


< 8. Average Al concentrtaions (7.9 mg/L) of the aerobic lysimeters were significantly higher than those of the anaerobic lysimeters (0.28 mg/L).

Generally, Al leaching is not greatly affected by redox conditions but mainly

controlled by pH. Meima and Comans (1997) reported that Al solubility was low in the pH range 6 to 7, which corresponds to the Al results in the aerobic condition presented in Figure 3-1. Among many ligands forming Al compelxation, hydroxide ion (OH) is known as a major ion to control the solubility of Al in aquatic systems. The equilibrium of Al with gibbsite (AI(OH)3)), an Al-OH complex, is characterized by a U-shaped pHleaching curve (Eary, 1999). However, Al concentrations in anaerobic lysimeters did not appear to follow with the solubility of gibbsite. Besides pH and redox conditions, large differences in leachate characterisitcs between aerobic and anaerobic lysimeters included the high organic content and anions such as floride and sulfate in the leachate of the anaerobic lysimeters. The most likely explanation of low Al solubility in the anaerobic conditions is complexation of Al and organic matter. Tipping (2005) reported that Al solubility is strongly associated with organic content. Skyllberg (2001) also reported that high dissolved organic content made Al solubility significantly decrease. Therefore, low solubilitity of Al during the first phase of the anaerobic lysimeters is the result of complexation of Al with high concentrations of organic matter.

3.3.1.2 Arsenic

Figure 3-2 depicts the change in As over time. High As concentrations were

observed from the anaerobic lysimeter before day 220. The greatest As concentration was

3.2 mg/L on the day 89. The As concentration then lowered below 1.5 mg/L after day 400. For lysimeter 4, As concentrations were continuously low, showing the lowest value,

0.27 mg/L on the last sampling day (day 741). In contrast to the anaerobic lysimeters, As





52


concentrations in the aerobic lysimeters decreased initially and increased with increasing pH. The As leaching pattern of the aerobic lysimeters appears similar to Al leachate trends. The lowest As concentration of lysimeter 1 was 0.12 mg/L on the day 163. For lysimeter 2, extremely low As concentrations were observed, with several samples below the detection limit (0.011 mg/L) despite a pH < 6. Overall, As dissolved in the leachate of the aerobic lysimeters was significantly lower than that of the anaerobic lysimeters (p <

0.05).

Figure 3-9 depicts the distribution of As concentrations observed from the aerobic and anaerobic lysimeters at various pH conditions. It has been reported that As solubility changes with pH and is characterized by a U-shaped curve in oxidizing conditions (Drever, 1988). However, Carbonell-Barrachina et al (1999) reported that in the presence of sulfide, Fe, and Mn, the solubility of As was dramatically lower and did not follow a U-shaped solubility curve. However, As concentrations were not impacted by these constituents in the anaerobic lysimeters. The most likely explantion is that the anaerobic lysimeters had poor-anoxic conditions during the first phase. The low sulfide concentrations at a pH < 6 confirm that the redox potential was not low enough for sulfide to become involved in As precipitation. For the aerobic lysimeters, As concentrations were low under acidic conditions and increased up to I mg/L at a pH of 9.

Masscheleyn et al (1991) found that As solubility decreased substantially as the redox potential increased. The changes in As solubility are also associated with the oxidation state of iron; Fe (III) has a strong affinity for arsenate. Therefore, low arsenic concentrations are likely dictated by the low solubility of arsenate. Under oxidizing conditions, As solubility may increase or decrease by pH changes. This is because of the





53


effect of pH on the total oxyaionic arsenate concentrations by pH. This changes in solubility results in the relatively higher concentrations of As at alkaline conditions observed in the aerobic lysimeters (Figure 3-2).

3.3.1.3 Chromium

The initial Cr concentrations of the anaerobic lysimeters were higher than those of the aerobic lysimeters (Figure 3-3). However, Cr concentrations in the anaerobic lysimeters gradually decreased to below 0.05mg/L by the day 453. As the pH of lysimeter

4 changed to moderately alkali (pH > 7.4) after day 464, minor increases in Cr concentrations were observed. In contrast, clear Cr leaching trends were not exhibited by the aerobic lysimeters before day 100, but an increase in Cr concentration did occur following day 150. This increase in Cr concentration corresponds to an increase in pH. Overall, the average Cr concentrations of the aerobic lysimeters were significantly greater than those of the anaerobic lysimeters. This may be because thermodynamically Cr can be present as an ionic form at alkaline pH under oxidized condition.

The toxicity of Cr is determined by its oxidation state. Among the various Cr oxidation states, only trivalent and hexavalent forms are taken into consideration in natural aquatic systems. Hexavalent Cr is considered more toxic than trivalent Cr due to its high mobility and solubility. Cr (VI) may be reduced to Cr (III) at low ORP potential. Cr (VI) becomes unstable and is reduced to Cr (III) at low pH. In order to maintain the oxidation state of Cr as Cr (VI) at a low pH, it is necessary to keep highly oxidizing conditions (Richard and Bourg, 1991). In contrast to other metals such as As and Cu, Cr (III) is not likely to precipitate with sulfide. Chromium solubility is mainly controlled by Cr(OH)3(s). Generally, Cr(OH)3(s) is formed in a pH range of 6.5 to 7 under moderately oxidizing or reducing conditions.





54


According to the potential-pH diagram of Cr (Figure 3-10), total Cr obtained from both aerobic and anaerobic lysimeters in an acidic environments is likely to be Cr (III) as Cr(OH)2+. In contrast, dissolved Cr from the aerobic lysimeter at a pH 9 could be Cr (VI) as CrO42-. Since all Cr species presented on the potential-pH diagram are based upon assuming thermodynamic equilibrium, all Cr obtained from the aerobic lysimeters at high pH may not be Cr (VI).

It is noted that an increase in Cr was observed from lysimeter 4 around a neutral pH ("A" in Figure 3-3). The most likely explanation for this is the lower Fe concentrations of lysimeter 4 than of those of lysimeter 3 (Figure 3-6). In the presence of Fe, Cr may be precipitated with Fe rather than OH due to rapid kinetics. The complexation of Fe and Cr decreases Cr solubility lower than the complexation of OH and Cr (Eary and Rai, 1987). Therefore, an increase of Cr concentration at the end of lysimeter 4 would be the result of a decrease of Fe concentrations.

3.3.1.4 Copper

Overall copper concentrations of the aerobic lysimeters were one to three orders of magnitude higher than those of the anaerobic lysimeters (Figure 3-4). For the aerobic lysimeters, clear Cu leaching patterns over time were not observed, but relatively large changes in Cu concentrations were observed at lysimeter 1 from the day 140 to 190. This period of time corresponded to a pH change from 5.5 to 9. For anaerobic lysimeters, Cu concentrations gradually decreased for the first 450 days. The initial Cu concentrations of lysimeter 3 and 4 were 0.082 and 0.234 mg/L, respectively. Although the concentrations slightly increased after day 450, final concentrations of Cu remained lower than the initial values.





55


Copper solubility is controlled by several Cu-containing minerals forming

complexation with Fe and sulfide. In addition, Cu sulfides may coexist with the sulfides of other metals such as Zn, Pb and As (Faure, 1991). Representative mineral deposits formed with OH-, Fe and/or sulfide include chalcocite (Cu2S), chalcopyrite (CuFeS2), cuprite (Cu20) and malachite (Cu2(OH)2CO3). These minerals are widely distributed over a pe-pH diagram. Ionic Cu is present only at a pH less than neutral and under highly oxidizing conditions (pe > 2.5). For these reasons, high concentrations of Cu may not be found under landfill conditions where low ORP and neutral pH are predominant. These concepts can be applied to the Cu distribution patterns displayed at a low pH in the pHCu concentration chart shown in Figure 3-11.

However, there is a disparity in the Cu concentrations observed and those

thermodynamically predicted for the alkali conditions of the aerobic lysimeters; most of the Cu precipitated by complexation with various Cu-containing mineral deposits at alkali and oxidizing conditions. Edwards et al (2000) reported high Cu concentrations in drinking water at alkali conditions, calling it 'the blue color phenomenon' since water color changed to blue with high concentrations of Cu. Critchley et al (2004) explained this 'blue water' was caused by microorganism-intermediated-Cu leaching from a part of the water delivery system. In this research, 'blue water' was also observed from condensate passing through a copper tube which connected to a gas collection system of an aerobic lysimeter. Further development of Cu corrosion caused a small hole on the same copper tube and called for a replacement of the copper tube with plastic materials (Figure B-7).





56


Another possibility of copper leaching from the aerobic lysimeters is the binding of Cu with ammonium (NH3+). Since both Cu and ammonium are cations, their complexations are present as an ionic form and can be dissolved in aquatic systems. Arzutug et al (2004) reported that Cu leaching increased with ammonia concentrations. However, complexation of Cu and ammonia may occur in a relatively narrow range of ORP and pH (Hoar and Rothwell, 1970). When plotting Cu and ammonia data obtained from the aerobic lysimeters, no clear evidence to prove the relationship between Cu concentrations and ammonia was found (r2 = 0.021). Furthermore, since Cu complexes with sulfide rather than ammonia in the presence of sulfide (Alymore, 2001), it may be difficult to leach high concentrations of Cu under landfill conditions.

3.3.1.5 Lead

Figure 3-5 depicts the changes in lead concentrations over time. For aerobic

lysimeters, Pb concentrations dramatically increased to 1.7 2 mg/L within 30 days and then gradually decreased. After the pH of the aerobic lysimeters stabilized, Pb concentrations decreased to levels similar to the anaerobic lysimeters. In contrast, little change in Pb concentrations was observed from the anaerobic lysimeters and low concentrations of Pb were maintained over the test period. Generally, Pb concentrations in landfill leachate have been reported to be very low (Charlatchka and Cambier, 2000; Jang and Townsend, 2003). This is because lead may precipitate with various ligands such as carbonate ions (C03-), sulfide, and volatile fatty acids (VFA).

Lead solubility is generally controlled by carbonate, or other Pb hydroxides and phosphate in noncalcareous soils (Bradle, 2005). Charlatchka and Cambier (2000) concluded that Pb solubility increased under oxidizing conditions at a pH of 6.2 However, Pb may precipitate as a form of PbS under reducing conditions in the presence





57


of sulfur. Lead is generally present in an ionic form at a pH < 6 under oxidizing conditions (Drever, 1988).

As shown in Figure 3-5, Pb leached from the aerobic lysimeters significantly

greater than from the anaerobic lysimeters. Most samples with high concentrations were distributed in the acidic phases. This leaching pattern corresponds to the characteristics of Pb previously discussed. For both aerobic and anaerobic lysimeters, Pb concentrations decreased with an increase in pH. Most Pb in alkali conditions may be precipitated as forms of PbCO3 or PbS depending upon redox potentials.

3.3.1.6 Iron

As shown on Figure 3-6, initial Fe concentrations of the aerobic lysimeters

(110mg/L for both aerobic lysimeters) were higher than those of the anaerobic lysimeters (20-22mg/L). Iron concentrations of the aerobic lysimeters increased to 250mg/L on the 30th day and then gradually decreased. During changes in pH of the aerobic lysimeters to alkali conditions, Fe concentrations lowered substantially to below 10mg/L. In contrast to the aerobic lysimeters, Fe concentrations of the anaerobic lysimeters increased from 20 to 600mg/L for the first 450 days and then decreased with increasing pH. Although lysimeters 3 and 4 are both anaerobic lysimeters, the final Fe concentrations were substantially different (1 65mg/L and 5.6 mg/L for lysimeters 3 and 4, respectively).

Generally, free Fe concentration is strongly associated with the redox condition of the system. Iron is present in aquatic systems in two oxidation states; Fe (III) and Fe (II). Ferric (Fe3+) and ferrous (Fe 2) irons can be transformed to each other depending upon the redox conditions. Iron (III) is precipitated as a mineral deposit such as Fe203 or Fe(OH)3 at a pH > 5. Iron (III) is also involved in complexation with metals. Under moderately oxidizing and reducing condition, Fe (II) ions are dominant in the pH range





58


of 5 to 9. In the presence of sulfur, Fe (1I) is likely to be precipitated as pyrite (FeS2) at pH > 5 under reducing conditions (Drever, 1988). Despite the many other reactions Fe is involved with, Fe solubility is strongly controlled by sulfide concentrations. Thus, in order to lower Fe concentration, the redox potential of the system needs to be low enough to reduce sulfate to sulfide. Sulfide concentrations of lysimeter 3 were lower than those of the other lysimeters, resulting in greater concentrations of Fe observed from lysimeter

3.

3.3.1.7 Manganese and Zinc

Leaching patterns of Mn and Zn look very similar for the aerobic and anaerobic lysimeters (Figures 3-7 and 3-8) and are thus discussed together. Relatively high concentrations of Mn and Zn were exhibited from the aerobic lysimeters for the first 150 days. The highest concentrations of Mn and Zn were 11 mg/L and 270 mg/L, respectively. Manganese and zinc concentrations then substantially decreased to below

0.2 mg/L and 10 mg/L, respectively. In contrast, little change in Mn and Zn concentrations was observed from the anaerobic lysimeters for 450 days, but decreased following that period. This corresponds to when the pH of the anaerobic lysimeters increased.

Since Mn and Zn are mainly precipitated by sulfide, differences of Mn and Zn concentrations between lysimeters 3 and 4 are strongly associated with the sulfide concentrations present in each lysimeter. Additionally, the solubility of Zn and Mn is associated with organic matter. A decrease in Zn and Mn solubility accompanies an increase of pH and could be accounted for by the generation of pH-dependent charge sites on organic matter (McBride and Blasiak, 1979; Miyazawa et al, 1993).





59


3.3.2 Organic Wastes as Absorbents of Heavy Metals

Analytical results of the 8 metals absorbed on office paper (OP), cardboard (CB),

newspaper (NP), and wood blocks (WD) are shown on Figure 3-12. Concentrations of Al, As, and Cu adsorbed on the lignocellulosic materials of the aerobic lysimeter appeared greater than those of the anaerobic lysimeter.

Heavy metals adsorbed on solid wastes from the aerobic and anaerobic lysimeter were statistically analyzed using the ANOVA test. Test results are presented in appendix C. All metals, except for Pb, absorbed on CB in the aerobic lysimeter and were significantly higher than those of the anaerobic lysimeter. Other significant differences between the aerobic and anaerobic lysimeters were found with Al, As, Mn and Cu adsorbed on NP, OP and WD.

Figure 3-13 depicts the differences of total mass of metals adsorbed on

lignocellulosic materials between the aerobic and anaerobic lysimeters. These values were calculated by multiplying the metal concentrations by the mass of each waste obtained from the garbage separation. Interestingly, the observed trends of adsorption of some metals did not to correspond with their leaching trends. These trends can be found from adsorption trends of As, Mn and Pb. The amounts of metals leached and adsorbed between aerobic and anaerobic lysimeters are compared in Table 3-5. These results indicate that metal adsorption may be influenced by environmental conditions such as pH, redox, and the presence of other ligands. Ravat et al (2000) reported that the binding of selected metals (Zn, Cu, and Pb) on lignocellulosic materials is strongly pH-dependent in the absence of interference from other ligands. Adsorption of Fe on organic matter is mainly controlled by the oxidation states of Fe; Fe (III) has greater affinity for organic matter than Fe (II) does (Jansen et al., 2003), corresponding to the large differences of Fe





60


adsorbed between the aerobic and anaerobic lysimeters (Table 3-6). Adsorption of As also mainly occurs when As is oxidized. As the pH rises, ionic forms of As changes progressively (H2AsO4, HAsO42 and AsO4 3) with each species showing different adsorption properties (Drever, 1988).

It is noted that there were large differences between metals released through

leachate and metals that remained in the lysimeters by adsorption. These differences can be numerically expressed using the ratios between metal leached (LC) and adsorbed

(AD). The smallest LC/AD ratio can be found from Al; only 0.06% of the amount of Al adsorbed was released from the lysimeters. The LC/AD ratios of most metals fell into around or below 2%. Relatively high LC/AD ratios were exhibited from a few metals in the anaerobic lysimeter; LC/AD ratios of As, Mn and Zn were 13.8%, 16.5% and 8.6%, respectively. However, if the amounts of metals precipitated as particulate forms without adsorption are taken into consideration, the ratio of metals between leached and remained would be much smaller than LC/AD ratios.

Lignocellulosic materials such as paper and wood products occupy as much as 45% of MSW landfills (USEPA, 2003). Cellulose and lignin are reported as the major heavy metal adsorbents (Basso et al, 2004). Lignin especially provides many chemical functional groups such as carboxyl and phenolic groups. Babel and Kurniawan (2003) concluded that lignin was considered as the best low-cost adsorbent for Pb and Zn. Basso et al. (2002) also reported that maximum sorption capacity increases due to lignin contents during Cd sorption research. Cellulose has also been heavily demonstrated to remove heavy metals such as Cd, Cu, Ni, Zn and Pb (Sublet, 2003; Okieimen et al., 2005 and Shukla and Pai, 2005).





61


Figure 3-14 depicts metal concentrations adsorbed on selected lignocellulosic

materials (newspaper and cardboard) and plastic waste. Greater concentrations of most metals except for Pb and Zn adsorbed on lignocellulosic materials were observed in the aerobic lysimeter. Bradle (2005) explained that many metals tend to adsorb the organic matter as the pH increases under oxidizing condition. Zhang and Itoh (2003) reported that carbonized mixture of polyethylene terephthalate (PTE) and waste ash could be used as a metal sorbent. However in this research, metal concentrations adsorbed on plastic waste were substantially low in comparison with metal concentrations adsorbed on the lignocellulosic materials.

3.4 Discussion

Among the various factors affecting heavy metal leaching under landfill conditions, the redox and pH may play the most critical role. Under the given redox condition and pH, the metal oxidation state, ligands, adsorption behavior can be determined. In many cases, metal precipitation can be controlled by Fe (II) and sulfide concentrations in both aerobic and anaerobic condition. Cr, Cu, Pb, Zn and As are reported to adsorb on hydrous ferric oxide at pH > 6, and the precipitation of Cu, Fe, Pb, Mn and Zn is controlled by sulfide in anaerobic condition. In addition to those ligands, hydroxide ion (OH) also can play an important role to precipitate Al and Cr (Drever, 1988). The various chemical interactions are depicted in Figure 3-15.

3.4.1 Overall Comparison of Metal Behavior

Figure 3-16 describes the overall trends of metal leaching in aerobic and anaerobic lysimeters. Among 8 metals under consideration, (Al, As, Cr, Cu, Pb, Mn, Fe and Zn) greater concentrations of Al, Cr, Cu and Pb were observed in the leachate of the aerobic lysimeters, and As, Mn, Fe and Zn were observed in the leachate of the anaerobic





62


lysimeters. For the anaerobic lysimeters, the metal leaching occurred in the acid phase, while occurrence of the metal leaching was relatively well distributed over the pH (5 < pH < 9) for the aerobic lysimeters except for Pb.

Cumulative masses of metals were calculated by the multiplication of the amount of leachate used for analysis by the concentration of the metals in the leachate sample (Figure 3-17). The total amounts of leachate produced and used for the analysis are summarized in Table 3-3.

Among the 8 metals under consideration, As, Fe, Pb, Mn and Zn increased

substantially for the first 10 to 15% of mass loss and reached a plateau. These leaching patterns indicate that metal leaching mainly occurred at the initial phase, an acidic environment, in both aerobic and anaerobic conditions. In contrast to these metals, only minor change in cumulative masses of Al and Cu was observed from the anaerobic lysimeters whereas a consistent increase in these metals was exhibited from the aerobic lysimeters. Relatively high concentrations of Al were exhibited during the initial stage of the anaerobic lysimeters, but no further changes were observed. It is notable that cumulative concentrations of Al and Cr increased more rapidly after 25% mass loss occurred. An increase in the rate of accumulation of these metals corresponds to an increase in pH to 9.

Overall metal leaching behavior is strongly associated with pH and redox

conditions. Since the anaerobic lysimeters remained in the acidic condition (pH < 6) for more than 400 days, great amounts of metals such as As, Mn, Fe, Cr and Zn were released through the leachate. In contrast to the anaerobic lysimeters, greater cumulative concentrations of Al, Cu and Pb were observed from the aerobic lysimeters. Leaching of





63


these metals might be influenced by the different environment of the aerobic lysimeters such as an alkaline pH and oxidizing conditions.

3.4.2 Comparison to Other Studies

Generally, heavy metal concentrations found in anaerobic landfills are reported low (Kjeldson et al., 2002). The presence of sulfide and the low solubility of metals at neutral pH may reduce metal concentrations in leachate. Metal concentrations of the aerobic and anaerobic lysimeters along with MSW leachate summarized from the literature are presented in Table 3-6. Metal concentrations of the aerobic lysimeters listed in Table 3-6 are the average value of the lysimeters 1 and 2. For the anaerobic lysimeters, since lysimeter 3 remained in acidic condition for most of a test period, metal results of lysimeter 3 during the methanogenic phase were not included in Table 3-6. For the anaerobic lysimeters, As and Zn concentrations were substantially higher than those of MSW leachate during acidic phase. They were reduced then during the methanogenic phase and similar to those of MSW leachate despite the presence of CCA-treated wood. Cu concentrations of the anaerobic lysimeters were extremely low, and they were lower than even drinking water standards. Although most metal concentrations of the anaerobic lysimeters were greater than drinking water standards, they were similar or lower than those of general MSW leachate during the methanogenic phase.

For the aerobic lysimeters, As, Cr, Fe, Mn and Zn concentrations were lower than those of MSW leachate and the anaerobic lysimeters during the first acidic phase. Aluminum and copper concentrations were greater than those of the anaerobic lysimeters. Only Pb concentration was greater than that of MSW landfill leachate and the anaerobic lysimeters. However, different aspects of metal leaching were observed from the aerobic





64


lysimeters during the alkaline phase; the concentrations of most metals except for Fe were greater than those of the anaerobic lysimeters.

The leaching behavior of CCA-treated wood of the aerobic and anaerobic

lysimeters was compared to a similar study (Jambeck, 2004). Jambeck (2004) researched the leaching behavior of CCA-treated wood mixed with MSW through the 6.7-m high PVC column tests. A total 2% of CCA-treated wood was included in the column and rain water was used for Jambeck's study while 1% of CCA-treated wood and DI water were used for the present study (Table 3-7). In comparison metal leaching results from Jambeck's study proved similar to the anaerobic lysimeter here (Figure 3-18). Extremely low Cu concentrations were observed in both studies. The range of Cr concentrations of Jambeck's study was higher than that of the anaerobic lysimeters during acid phase, but the median of Cr concentrations of Jambeck's study was bottom of the range. In contrast to anaerobic condition, significantly different leaching trends of As and Cu were exhibited from the aerobic lysimeters; overall As concentrations of the aerobic lysimeter were lower than those of the anaerobic column studies during the acid phase. The 95th percentile of As concentrations of the aerobic lysimeter was in the range of the anaerobic system, but they were detected at the very beginning of the aerobic lysimeter operation. After pH stabilized, the median As concentrations of both the aerobic and anaerobic systems became identical. However, Cu concentrations were two orders of magnitude higher than those of anaerobic lysimeters for both the acid and methane (alkaline) phase.

In comparison with other column study (Jambeck, 2004), it can be concluded that leached As, Cu and Cr concentrations might not always be followed by the initial mass of CCA-treated wood and metal concentrations contained. The differences between As and





65


Cu concentrations between the anaerobic and aerobic columns proved too high to be comparable. Within the same anaerobic systems, As and Cr concentrations between Jambeck and this study were not much different despite different initial As and Cr masses. Similar leaching trends could be observed in the same anaerobic or aerobic system, but overall concentrations of As and Cr per initial mass may depend more on the chemistry of the system.

3.4.3 Implication for Disposal of Heavy Metals

In this research, CCA-treated wood and CRT monitor glass were used as metal

sources to represent treated wood and electronic waste. Though the land-disposal of these wastes was banned in several states in the U.S. (SWANA, 2003), they can still be landdisposed as a form of home appliances and ash. Since the greatest amount of As may be leached during the first acid phase of anaerobic landfills, landfill owners need to monitor the leachate quality during landfill construction. Though thermodynamically As may precipitate with sulfide, a maximum of 70% of As dissolved in solution may combine with sulfide at pH 8 (Carbonell-Barrachina, 1999). This adsorption ratio decreased as the pH decreased. In addition to the sulfide, Fe, organic matters, and carbonate are also known as As adsorbent. However, the adsorption efficiency of those ligands were low relative to sulfide (Carbonell-Barrachina, 1999). Thus, in the presence of an As source, As can be found in landfill leachate for all operation periods. For aerobic landfills, As concentrations may slightly increase during the alkaline phase. Overall As concentrations found in aerobic landfills may be lower than those of anaerobic landfills.

Thermodynamically, Cr concentration in landfill condition during all phases may be low. During the first acid phase, Cr may be combined with high concentrations of Fe

(11), and Cr may precipitate with sulfide during the methane phase. The lower





66


concentration of Fe (11) may increase the Cr solubility, but the change is very small. Cu concentrations in anaerobic landfills are typically extremely low. However, high concentrations of Cu may be found in both acid and alkaline phase of aerobic landfills. Since the microorganisms which contribute Cu leaching may corrode all Cu-made equipment connected to the landfill, it is important to avoid using any Cu-containing equipment for gas and leachate collection systems. Pb has high solubility under oxidizing conditions at pH < 6. Thus it would be recommendable to monitor leachate quality for the first acid phase of aerobic landfills. However, air addition facilities are generally installed after landfill closure, Pb concentrations may not be high over operation period. Pb concentration in anaerobic landfill conditions is generally low.

As previously discussed, air addition into a current anaerobic bioreactor landfill may enhance waste decomposition substantially. However, unlike anaerobic landfills, concerns about high Al and Cu concentration and a risk of Cr (VI) may arise. In order to avoid these potential risks, it would be recommendable to inject air into the shallow well rather than the deep well. Since Al, Cu and Cr are redox-sensitive, great amount of these metals could be reduced after passing through the anaerobic zone. However, since air injection into the shallow wells may reduce the air diffusion efficiency into a landfill, further economical and efficiency of air distribution analysis are needed.

3.4.4 The Imp act of Air on Metal Mobility

Although total amounts of metals leached were not considerably high, it is

necessary to pay great attention to certain metals due to changes of their toxicity by different pH and redox conditions. For example, among Cr species dissolved in leachate, Cr (111) can be dominant in current anaerobic sanitary landfills, however, thermodynamically Cr (VI) becomes a major Cr component in the environment formed





67


upon air intrusion. Hexavalent chromium is a highly toxic metal causing decreased pulmonary function and pneumonia (Bradle, 2005). On the contrary, toxicity of As can be reduced by air injection. In oxidizing conditions, Arsenate can be oxidized to As (VI), which is less toxic. Since As (VI) is less mobile and has a low solubility, the total amount of As leached can be reduced. The influence of air injection on the amount of metals can be more clearly understood by comparing the percentage of As, Cr and Cu (Table 3-4). Under the scenario of high As content in leachate caused by co-disposal of fly ash or ground CCA-treated wood, it would be recommended to inject air in order to reduce As toxicity and the amount leached. However, Cr (VI) in an aerobic landfill can be present as an ionic form under oxidizing conditions, and the cumulative mass may increase over a period of time. For these reasons, it is necessary that landfill personnel develop a strategy to optimize the influence of air injection.

Based on the metal results, the metal leaching behavior of old landfill metals can be predictable. Kjeldsen et al. (2002) proposed that the condition of old landfills would be aerobic due to air intrusion. Once air intruded in anaerobic landfills, metal leaching may occur by the dissociation of the metals that adsorbed on decomposed organic matters. Among 8 metals under consideration, Pb would be the most leachable metal. As Figure 313 shows a large amount of Pb was adsorbed on organic matter. Bozkurt et al. (1999) explained that the pH of the landfill would change to acid again after the long-term process. This was due to the oxidation of sulfate and organic matter. Therefore, the moderate oxidizing condition, and low pH, would accelerate Pb leaching.

3.5 Conclusions

Research on the fate of metals in simulated aerobic and anaerobic landfills was

conducted. The leaching behavior of selected metals was significantly different between





68

aerobic and anaerobic lysimeters. Among the 8 metals evaluated (Al, As, Cr, Cu, Pb, Mn, Fe and Zn), the concentrations of Al, Cu, Cr and Pb in leachate of the aerobic lysimeters were significantly greater than those of the anaerobic lysimeters, and the average concentrations of As, Fe, Mn, and Zn in the anaerobic lysimeters were significantly greater in concentration than observed in the anaerobic lysimeters.

After a test period, one each of the aerobic and anaerobic lysimeters was

dismantled and the metals adsorbed on decomposed lignocellulosic waste were analyzed. Greater concentrations (mg metal/kg waste) of Fe, Mn, As, Al and Cu were found from the aerobic lysimeters. All metals, except for Pb, adsorbed on the cardboard in the aerobic lysimeter and were significantly higher than those of the anaerobic lysimeter. Through this research, aerobic landfills were found to have a greater potential to release several metals through leachate than that of anaerobic landfills; aerobic landfills have greater Al, Cr, Cu and Pb leaching potential due to oxidizing condition. Experimental results confirmed this notion. In the presence of CCA-treated wood and electronic waste (CRT monitor glass), high As and Pb concentrations were observed from the anaerobic lysimeter and the aerobic lysimeter during the acid phase, respectively.





69


Table 3-1 .Heavy metal sources in fabricated waste stream.
Waste components Contained heavy metals % of component in fabricated
____________________waste
CCA treated wood Copper, Chromium and 1%
Arsenic
Cathode-ray Tube (CRT) glass Lead 1%
Aluminum sheet Aluminum 4%
Galvanized steel sheet Zinc, Manganese and Iron 4%



Table 3-2. Results of statistical analysis of metal leached between aerobic and anaerobic Average concentrations (mg/L) F P-value F-crit
Aerobic Anaerobic
Al 7.89 1.28 206.89 8.3E-33 3.89
As 0.40 1.28 66.15 4.1E-14 3.89
Cr 0.19 0.10 40.67 1.19E-09 3.89
Cu 2.87 0.02 81.10 1.58E-16 3.89
Fe 35.06 167.68 53.09 6.92E-12 3.89
Mn 1.91 4.57 35.80 9.75E-09 3.89
Pb 0.22 0.03 31.32 7.07E-08 3.89
Zn 54.36 201.12 66.30 3.87E-14 3.89



Table 3-3. The amount of leachate produced and used for analysis lys 1 lys 2 lys 3 lys 4
leachate produced (mL) 8,717 9,747 18,571 15,979
leachate released (mL) 4,024 4,081 6,135 6,111
(used for analysis) I





70


Table 3-4. Leachability of As, Cr, and Cu

Initial conc. (mg/lys) mg released % released
aerobic anaerobic aerobic anaerobic

As 1279.1 Lys 1 Lys 3 1.53 8.21 0.12% 0.64%
mg Lys 2 Lys 4 1.17 9.30 0.09% 0.73%

Cr 1573.1 Lys 1 Lys 3 0.72 0.50 0.05% 0.03%
mg Lys 2 Lys 4 0.56 1.24 0.04% 0.08%
723.9 Lys 1 Lys 3 9.82 0.10 1.36% 0.01%
CU mg Lys 2 Lys 4 6.37 0.34 0.88% 0.05%



Table 3-5. Comparison of cumulative mass of metal dissolved in leachate and adsorbed
on li nocellulosic materials (unit: mg)
Aerobic Anaerobic LC/AD LC/AD
lys 1 lyis 2 lys 3 lys 4 (lys 2) (lys 4) of lys 2 of lys 4
Al 27.3 25.9 7.8 8.7 41900 15700 0.06% 0.06%
As 1.5 1.2 8.2 9.3 110 67.6 1.06% 13.76%
Cr 0.7 0.6 0.5 1.2 222.2 243 0.25% 0.51%
Cu 9.8 6.4 0.1 0.3 238.1 151.2 2.68% 0.22%
Fe 161.3 64.2 1200 506.4 31100 19200 0.21% 2.63%
Mn 10.5 2 31 29.3 494.2 177.3 0.40% 16.51%
Pb 0.6 0.5 0.2 0.1 47.8 156.7 1.13% 0.08%
Zn 292.9 64.2 1,100 1,100 9,400 12,500 0.69% 8.60%






71


Table 3-6. Comparison of average metal concentrations of the aerobic and anaerobic
lysimeters with MSW leachate and regulatory levels (SAIC, 2000; USEPA,
1996 and 2003; Kjeldsen et al., 2002) (unit: mg/L)

Aerobic lysimeters Anaerobic lysimeters Drinking
MSW TLP TC water
leachate Alkali Methane limits str
Acid phase phase Acid phase phase

Al 15.05 4.56 9.07 0.31 0.17 0.2*
As 0.44 0.19 0.55 1.5 0.43 5 0.01
Cr 0.24 0.09 0.26 0.1 0.14 5 0.1
Cu 0.14 4.01 1.75 0.02 0.07 1.3
Fe 3.00 61.66 3.89 188.26 10.32 0.3*
Pb 0.13 0.37 0.03 0.03 0.01 5 0.015
Mn 6.08 3.35 0.09 5.63 0.07
Zn 5.1 96.32 8.26 250.52 4.58 5*
* secondary drinking water standards



Table 3-7. Comparison of characteristics of CCA-treated wood used for Jambeck (2004)
and this study
Jambeck(2004) This study
The % of CCA-treated included in waste stream 1% 1% 1%
As 1390 20.0 1960 27. 1330 10
Cu 814 52.4 1340 54.0 2350 50
Cr 1450 68.3 2550 48.0 2890 56






72



100
AEROBIC Lys
Lysi1 O Lys 2

o *

10 0*00 0 0 00
00 0 0 00 00 0
0 *0 0
0 0 0


0 0
0







0.1
0 100 200 300 400
100
ANAEROBIC
A Lys 3 Lys4 10 O




-A

A
1&


A AA AA
0.1 A IL



0.01



0.001
0 200 400 600
Days

Figure 3-1. Changes of Al concentrations over time





73





1.0
AEROBIC Lys 1
Lysi1
0.8 00 0 Lys 2
0
0.8 0 0



9 0.6 0
O O O
0.4 O-l 0
0O
40o *o

0.
<0.4 * 0 0



0 0 go
0
0 0
0.2 g
o0

0.0 o

0 100 200 300 400
3.5
A ANAEROBIC A Lys 3
3.0 4~ Lys4
2 5 -,- A
.. 2,0 A A x
2.5 A A
A
4 A A AA
2.0 A
8b A

< 1.5 -A A

A AA AA 4
1.0 A A
Al A A A tt

0.5 A AA
A k AA A 4Q4Z,

0.0
0 200 400 600 800
Days

Figure 3-2. Changes of As concentrations over time





74





0.5

Lys 1 AEROBIC O
0 Lys 2 So
0.4


00 0 0 0
p S

0.3 0
0 0
E o* 00
0
0.2 0 0 O o 0
0 S 0
*o0
o d" o

o 0 0
o 0o
0
0.0 11 1 1 1 1
0 50 100 150 200 250 300 350
0.8

ANAEROBIC A Lys 3
A Lys 4

0.6

A


0.4 A
SAA

A A
0.2 AA A
A AA
,tAAA
A PAA
SAA A

0.0,
0 200 400 600
Days

Figure 3-3 Changes of Cr concentrations over time






75





100
AEROBIC 0 Lys 1 O Lys 2



10 -0 0

e o
0 0 0 0
00


of* 0
U0 0 0
0 0000
0 0 00

0 0

0 0
0 O0

0.1 I I I
0 50 100 150 200 250 300 350
1
ANAEROBIC








0.1 A Ls
;A



-6 A
A A AL A A A



AAA Lys 4 AdUL A A A i&AA A A
UkAA A AAh A j AA
0.01 A&A A A AA,&A A A

Below Detection Limit A A A A


A Lys 3
A Lys 4
0.001
0 200 Days 400 600

DaysFigure 3-4. Changes of Cu concentrations over time Figure 3-4. Changes of Cu concentrations over time






76





10
AEROBIC 0 Lys 1
O Lys 2


1- 000
e*
0 00 0
0 0

0 o
0 0 0
0.1 0
0 0

0 0 0 0 QD
0 00 0


0.01 0





0.001 I I I I I
0 50 100 150 200 250 300 350
1

A Lys 3 ANAEROBIC
A Lys 4



0.1 A A AA A


A AA A
A AA A 4 A
A01 AA A
A:*r6A AA AAAAA AAA A A
A A
0.1 AA AAA A A
0.01 A

A A A
A A

Below detection limit of ICP for Pb

0.001 I I ,
100 200 300 400 500 600 700
Days


Figure 3-5. Changes of Pb concentrations over time





77





1000
AEROBIC Lys
Lys 1
0 0 Lys 2

100 0O O s
00 0 0
10O 0 0 0

8 o4
O0 0
E 1o
6 10

0 0 0
0 *
0.

01

0 .1 I I I I
0 50 100 150 200 250 300 350
1000
A Lys 3 ANAEROBIC AAA A A A 1%
A Lys 4 A AA A


A AA A
100 -A A A A


6A A


10 AA A







0 200 400 600
Days

Figure 3-6. Changes of Fe concentrations over time






78





100
AEROBIC O Lys 1

0 Lys 2


10 @
.0 00 0
0 0 0
0 00 0




11 00 0 0
0 0


0O 0
0.1 eo o 0.0O


.009 o 0 0 0


0.01 ,
0 50 100 150 200 250 300 350
100

ANAEROBIC A Lys 3 A Lys 4


10 -A
AAA


A 1 AA A'

A


A
0.1 A A
A A A



0.01 ,
100 200 300 400 500 600 700

Days


Figure 3-7. Changes of Mn concentrations over time






79





300
AEROBIC Lys 1

250 O Lys 2
250

0 0 200



E 150
N
0 0
100
o *

050
o g0 co 0 0 0 BDL


0 50 100 150 200 250 300 350

Days


500

A ANAEROBIC A Lys 3
A A Lys 4 400
A




200
A A
AA A
P300 A A A

AA
N 200 A A

AA
A A A A
A
100

AA


100 200 300 400 500 600 700

Days


Figure 3-8. Changes of Zn concentrations over time





80






3.5
As 0 lys 1

3.0 A O lys 2
A v lys 3
A A lys 4
2.5 A

V
o 2.0 VV

1.5 A
o 4

1.0 A*
A CkiA

0.5 AA A
.0 ()
0.0
4 5 6 7 8 9 10

pH


Figure 3-9. Distribution of As over a C-pH diagram





81


20 1.2

16- 1.0
CrOT4 -0.8


3+ 0.6

4rO

4-002.

0 r( J ( O ) s I 0 .
I4/ Cr (OH)OV


-82


"0 2 46 8012 14
pH

Figure 3 -10. Potential- pH diagram of Cr (Richard and Bourg, 199 1)






82






100
Cu



10 -0 00 @0
00 *t 000

0 g 0
0 A S 0.1
0V




0.0 0 Iys 2
VS 6 VV Iys 3
V ~A lys 4

0.001 III
4 5 6 7 8 9 10

pH


Figrue 3-11. Distribution of Cu over a C-pH diagram







83






18 0.035

16 Al CB As CB
16__NP 0.030 NP
-OP
'~14 OP WD
Sb WD 0.025
12
0.020
S10

8 0.015
6
0.010


0.005


0 IWLLU 0.000
2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw 2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw


0.08 0.08

Cr CB Pb CB
NP NP
OP OP
S0.06 WD 0.06 WD



F0.0
0.04 0.04

0
0.0


0.02 0.02
0.00 0.00

2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw 2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw



Figure 3-12. Adsorption of metal on solid wastes






84





0.16 10
Mn CB Fe CB

0.14 NP NP
OP 8 OP
0.12 WD WD

0.10 6

,F 0.08

0.06 4
0
,, 0.04
2
0.02

0.00 0
2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw 2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw


0.08 3.5
Cu CB Zn CB
NP 3.0 NP
OP 1OP
( 0.06 WD WD
E 2.5
'o 2.0

0.04
1.5
10
0.02

0.5


0.00 0.0 -2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw 2-1 2-2 2-3 2-4 4-1 4-2 4-3 4-4 Raw



Figure 3-12(continued)






85






50000 120
Al As

100
40000

80
30000 60
060
20000 ':



10000 20


0 0
aerobic anaerobic aerobic anaerobic
0
"0 300 300
Cr Cu
0
250 C 250 C

Ca
0 200 200
0

150 -150


100 -100


50 50


0 0
aerobic anaerobic aerobic anaerobic



Figure 3-13. The comparison of aerobic and anaerobic lysimeters in respect of total mass
of metals adsorbed on lignocellulosic materials




Full Text
9
Townsend et al., 1996), 35C was used as the starting point because the anaerobic seed
used was from a mesophilic digester.
2.2.3 Fabricated Waste Stream
The waste stream fabricated for this research was based on typical MSW
composition estimates previously reported for the U. S. and Florida (see Figure B-4 and
B-5). For simplification purposes, several minor components, such as textiles and tires,
were excluded from the fabricated waste stream. A greater portion of commingled paper
was allotted as a substitute for those excluded materials. The relative amount of office
paper, cardboard and newsprint in commingled paper (4.6 : 2.6 : 1) was again estimated
from previous published data (FDEP, 2003 and USEPA, 2005). Figure 2-2 presents the
fabricated waste stream composition used. Table 2-1 presents a description of each
component, the source, and the method of sample preparation. Commercial grade dog
food (Pedigree, USA) was used as the food waste portion of the fabricated waste stream.
To support complementary research on the fate of certain heavy metals in aerobic and
anaerobic landfill environment (chapter 3), a part of the wood waste fraction was
comprised of CCA and a part of the glass fraction was comprised of leaded cathode ray
tube (CRT) glass. Detailed waste components and their sources are presented in Table 2-
1 and Figure 2-2.
Mixed fabricated waste samples were created and loaded into the columns as four
distinct fractions to prevent waste component stratification in a particular place in the
column, (composition of the fabricated waste fractions and their weight are summarized
in appendix C). Prior to loading, 6 inches (15.3 cm) of river rock was placed at the
bottom of each lysimeter, and a geotextile was placed between the rock and waste. Each
waste fraction was then loaded and compacted until it occupied 25 % of the depth of the


Cr concentrations (mg/L)
180
Figure C-12. Cr concentration versus pH in leachate from the lysimeters


60
adsorbed between the aerobic and anaerobic lysimeters (Table 3-6). Adsorption of As
also mainly occurs when As is oxidized. As the pH rises, ionic forms of As changes
progressively (H2ASO4', HAs042' and As043') with each species showing different
adsorption properties (Drever, 1988).
It is noted that there were large differences between metals released through
leachate and metals that remained in the lysimeters by adsorption. These differences can
be numerically expressed using the ratios between metal leached (LC) and adsorbed
(AD). The smallest LC/AD ratio can be found from Al; only 0.06% of the amount of A1
adsorbed was released from the lysimeters. The LC/AD ratios of most metals fell into
around or below 2%. Relatively high LC/AD ratios were exhibited from a few metals in
the anaerobic lysimeter; LC/AD ratios of As, Mn and Zn were 13.8%, 16.5% and 8.6%,
respectively. However, if the amounts of metals precipitated as particulate forms without
adsorption are taken into consideration, the ratio of metals between leached and remained
would be much smaller than LC/AD ratios.
Lignocellulosic materials such as paper and wood products occupy as much as 45%
of MSW landfills (USEPA, 2003). Cellulose and lignin are reported as the major heavy
metal adsorbents (Basso et al, 2004). Lignin especially provides many chemical
functional groups such as carboxyl and phenolic groups. Babel and Kumiawan (2003)
concluded that lignin was considered as the best low-cost adsorbent for Pb and Zn. Basso
et al. (2002) also reported that maximum sorption capacity increases due to lignin
contents during Cd sorption research. Cellulose has also been heavily demonstrated to
remove heavy metals such as Cd, Cu, Ni, Zn and Pb (Sublet, 2003; Okieimen et al., 2005
and Shukla and Pai, 2005).


123
where Ae is the change in void volume; AH is the change in thickness of the waste layer;
H0 is the original thickness of the waste layer; ti is the starting time of secondary
settlement; and t2 is the ending time of secondary settlement.
Bjamgard and Edgers (1990) proposed the phase separate method to describe the
major causes of landfill settlement. They separated landfill settlement curve into two
phases, (Ca)min and (Ca)max by slope. The first phase, (Ca)m¡n, indicates that settlement
occurs by mechanical interactions such as delayed compression of the refuse. Landfill
settlement that occurs in the second phase, (Ca)max, is caused by both mechanical
interactions and waste decomposition (Figure 5-3)
5.2.4 Estimation of Mass Loss
The mass loss in each lysimeter was estimated by gas and leachate quality. The gas
volume generated was coupled with the biochemical reaction models for cellulose
biodegradation were used to calculate the mass of waste required for the moles of gas
obtained. Total organic carbon (TOC) concentrations were used to calculate the amount
of mass degraded, but remained in leachate before gas conversion. Details are described
in appendix A.
The settlement, gas and leachate data were collected for 1 and 2 years from the
aerobic and anaerobic lysimeters, respectively. At the end of the test period, lysimeters 2
and 4 were dismantled and partially decomposed waste was excavated. The waste loss
was then compared with the mearsured mass loss. The mass losses predicted of
lysimeters 2 and 4 were 4,069 g and 3,787 g, respectively, while the actual mass losses
were 4,044 g and 3,525 g, respectively.


TABLE OF CONTENTS
page
ACKNOWLEDGMENTS iv
LIST OF TABLES viii
LIST OF FIGURES x
CHAPTERS
1. INTRODUCTION 1
1.1 Problem Statement 1
1.2 Objectives 2
1.3 Research Approach 3
1.4 Outline of Dissertation 5
2. COMPARATIVE STUDIES OF LEACHATE AND GAS QUALITY OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILL BIOREACTORS 6
2.1 Introduction 6
2.2 Material and Methods 7
2.2.1 General Description of the Lysimeter 7
2.2.2 Temperature Control 8
2.2.3 Fabricated Waste Stream 9
2.2.4 Air Injection 10
2.2.5 Leachate and Gas Analysis 10
2.2.6 Recovery of the Anaerobic Lysimeters 11
2.2.7 Prediction of Waste Mass Loss 12
2.3 Results and Discussion 12
2.3.1 pH 13
2.3.2 Organic Carbon Concentration 14
2.3.3 Nitrogen 16
2.3.4 Dissolved Solids Content 17
2.3.5 Oxidation Reduction Conditions 18
2.3.7 Gas Quality 19
2.4 Discussion 20
2.4.1 Differences between Aerobic and Anaerobic Lysimeters 20
2.4.2 The Comparison of Leachate Parameters with Other Studies 21
2.4.3 Implications for Full-scale Application 22
v


4-6. Summary of cellulose and lignin content of the wood samples 112
5-1. (Ca)min and (Ca)max values of lys 1 through 4 132
5-2. k values of aerobic and anaerobic lysimeters 132
5-3. Comparison of compress indices between current study and other studies 133
A-l. Actual mass loss and predicted values of the aerobic and anaerobic lysimeter 155
A-2. Mass and density of wastes excavated by depth 155
C-l. pH of the aerobic and anaerobic lysimeters 189
C-2. Conductivity of the aerobic and anaerobic lysimeters 192
C-3. Alkalinity of aerobic and anaerobic lysimeters 194
C-4. Total dissolved solids (TDS) of aerobic and anaerobic lysimeters 196
C-5. Total organic contents (TOC) of aerobic and anaerobic lysimeter 197
C-6. Chemical oxygen demand (COD) of aerobic and anaerobic lysimeter 199
C-l. NH3+-N concentrations of aerobic and anaerobic lysimeters 201
C-8. Sulfide concentrations of aerobic and anaerobic lysimeters 203
C-9. Volatile fatty acids (VFA) of lysimeter 1 205
C-10. Volatile fatty acids (VFA) of lysimeter 2 206
C-l 1. Volatile fatty acids (VFA) of lysimeter 3 207
C-l2. Volatile fatty acids (VFA) of lysimeter 4 209
C-l3. Fabricated waste in lysimeters 211
C-14. Metal concentrations of lysimeter 1 212
C-l5. Metal concentrations of lysimeter 2 213
C-l 6. Metal concentrations of lysimeter 3 214
C-17. Metal concentrations of lysimeter 4 216
C-l 8. ANOVA results of metals and organic absorbence 218
IX


58
of 5 to 9. In the presence of sulfur, Fe (II) is likely to be precipitated as pyrite (FeS2) at
pH > 5 under reducing conditions (Drever, 1988). Despite the many other reactions Fe is
involved with, Fe solubility is strongly controlled by sulfide concentrations. Thus, in
order to lower Fe concentration, the redox potential of the system needs to be low enough
to reduce sulfate to sulfide. Sulfide concentrations of lysimeter 3 were lower than those
of the other lysimeters, resulting in greater concentrations of Fe observed from lysimeter
3.
3.3.1.7 Manganese and Zinc
Leaching patterns of Mn and Zn look very similar for the aerobic and anaerobic
lysimeters (Figures 3-7 and 3-8) and are thus discussed together. Relatively high
concentrations of Mn and Zn were exhibited from the aerobic lysimeters for the first 150
days. The highest concentrations of Mn and Zn were 11 mg/L and 270 mg/L,
respectively. Manganese and zinc concentrations then substantially decreased to below
0.2 mg/L and 10 mg/L, respectively. In contrast, little change in Mn and Zn
concentrations was observed from the anaerobic lysimeters for 450 days, but decreased
following that period. This corresponds to when the pH of the anaerobic lysimeters
increased.
Since Mn and Zn are mainly precipitated by sulfide, differences of Mn and Zn
concentrations between lysimeters 3 and 4 are strongly associated with the sulfide
concentrations present in each lysimeter. Additionally, the solubility of Zn and Mn is
associated with organic matter. A decrease in Zn and Mn solubility accompanies an
increase of pH and could be accounted for by the generation of pH-dependent charge
sites on organic matter (McBride and Blasiak, 1979; Miyazawa et al, 1993).


24
2.5 Conclusions
In this research, the gas and leachate quality from aerobic and anaerobic
simulated bioreactor landfills were compared. Waste streams referenced from EPA and
FDEP were loaded into 4 stainless steel lysimeters with a density of 3500 kPa. All
lysimeters were prepared with the same conditions, and two of them were assigned for
aerobic and two for anaerobic bioreactor landfill simulations. Leachate and gas generated
from the lysimeters were analyzed for chosen parameters to make comparisons between
aerobic and anaerobic landfills.
Leachate analysis results indicated that organic compounds as measured by COD,
TOC, BOD and VFAs in the aerobic lysimeters were degraded more rapidly than those in
anaerobic lysimeters. Except for the acidic phase, the pH of the aerobic lysimeters rapidly
increased and was stabilized around pH 9.0 while anaerobic lysimeters had remained in
acidic phase for more than 400 days, and stabilized exhibiting pH 7.3. The concentrations
of ammonia in anaerobic lysimeters increased along with an increase of pH. Ammonia
concentrations in aerobic lysimeters varied little over time, but ammonia levels were
significantly lower than those of anaerobic lysimeters. Sulfide results imply that both
aerobic and anaerobic zones were coexisting in aerobic lysimeters. This may be caused
by the limit of oxygen distribution in the lysimeters because of high density and low
hydraulic conductivity of wastes under overburden pressure.


130
El-Fadel and Khoury (2000) pointed out that the determination of the kinetic
coefficient (k) (or hydrolysis coefficient) would be at best a difficult task in landfills.
Hoeks (1983) and Ham (1988) also pointed out that this coefficient (k) would be changed
by landfill phase. The k values reported were 0.046, 0.028 0.139 and 0.462 1.386 yr'1
by slow, moderate, and rapid waste decomposition, respectively (Hoeks, 1983; Ham,
1988). From this point of view, settlement loss mass relationship obtained from this
research may be meaningful. The mass loss calculated from gas and leachate qualities
may provide the important tool to determine this kinetics coefficient. Using the
settlement data and mass loss, the kinetics coefficient can be determined by:
= \-e~h
M o
In
1-
M
M
= -kt
o y
(8)
Figure 5-6 shows the correlations of the equation (8) using the settlement data and
mass loss of the aerobic lysimeters. However, Hoeks (1983) and Ham (1988) mentioned,
for the anaerobic lysimeters, k values were variable by the different phases (Figure 5-7).
The results of the k value are summarized in Table 5-2. Figure 5-8 shows the example of
the application of the k values and settlement waste loss relationship. Park and Lee
(1997) pointed out that the bioconsolidation model might be useful to correlate landfill
settlement in terms of waste decomposition behavior, but landfill settlement behavior
would be return to the mechanical stage after biodegradable fraction is depleted.


109
Table 4-1. Methane yields, VS and mass fraction of the lignocellulosic materials in raw
waste
Office paper
Cardboard
Newspaper
Dog food
SYP
VS, %
86.8%
98.5%
95.3%
90.0%
77.0%
Mass fraction,
%
27%
16%
6%
5%
15%
Methane
yield
(L/gvs)
0.401
0.265
0.094
0.027
0.540
Table 4-2. Comparison of methane yields of MSW with other studies
Sorting method
Fermenter
CH4 yield
(L/g VS added)
References
Mechanical -sorted
Lab plant
0.260
Baere, 1984
Mechanical -sorted
Pilot plant
0.187
Baere and Verstraete, 1984
Mechanical -sorted
Pilot plant
0.230
Valorga, 1985
Hand-sorted
CSTR
0.390
Pauss et al., 1984
Hand-sorted
BMP assay
0.205
Owens et al., 1993
Source-sorted
CSTR
0.399
Mata-Alvarez et al., 1990
Mechanical -sorted
(pre-composted)
CSTR
0.145
Mata-Alvarez et al., 1990
-
Landfill
0.17*
CAA
-
Landfill
0.10*
AP-42
Hand-sorted
BMP assay
0.337
This study
* L/g of refuse
Table 4-3. Biodegradable volatile solid (BVS) of organic fraction of t
ie raw waste
VS
Biodegradable
fraction of VS
(BVS)
Biodegradable
fraction of dry waste
office paper
87%
74%
86%
cardboard
99%
65%
66%
newspaper
95%
21%
23%
dog food
77%
77%
77%
wood
90%
6%
7%


225
Myles, T. G. http://www.utoronto.ca/forest/termite/lig-mat.htm
Innocente, N., Moret, S., Corradini, C. and Conte, L. S., 2000, A rapid method for the
quantitative determination of short-chain free volatile fatty acids from cheese
Journal of Agricultural and Food Chemistry, 48(8), 3321-3323.
O'Keefe, D. M. and Chynoweth, D. P. 2000, Influence of phase separation, leachate
recycle and aeration on treatment of municipal solid waste in simulated landfill
cells, Bioresource Technology, 72, 55-66.
Okieimen, F. E., Sogbaike, C. E. and Ebhoaye, J. E., 2005, Removal of cadmium and
copper ions from aqueous solution with cellulose graft copolymers, Separation and
Purification Technology 44(1), 85-89.
Oweis, I. S. and Khera, R. P., 1998, Geotechnology of waste management, Boston, PWS
Publishing.
Owen W, Stuckey D, Healy J, Young L, McCarty P, 1979, Bioassay for monitoring
biochemical methane potential and anaerobic toxicity, Water Research, 13, 485-492.
Owens, J. M. and Chynoweth, D. P., 1993, Biochemical methane potential of municipal
solid-waste (MSW) components, Water Science and Technology, 27(2), 1-14.
Pagarwal. U, 2005, Raman imaging of lignin and cellulose distribution in black spruce
wood Picea mariana cell walls Proceedings of the 59th APPITA Annual
Conference and Exhibition incorporating the 13th ISWFPC, Carlton, Victoria,
Australia
Parawira, W., M. Murto, J. S. Read and B. Mattiasson, 2004, Volatile fatty acid
production during anaerobic mesophilic digestion of solid potato waste. Journal of
Chemical Technology and Biotechnology, 79(7), 673-677.
Park H. I. And Lee, S. R., 1997, Long-term settlement behavior of landfills with refuse
decomposition, Journal of Solid Waste Technology and Management, 24(4), 159-
165
Pauss, A., Nyns, E. J. and Naveau, FL, 1984, Production of methane by anaerobic
digesting of domestic refuse, EEC Conference on Anaerobic and Carbohydrate
Hydrolysis of Waste, Luxembourg, 8-10
Pohland, F. G., 1980, Leachate Recycle as Landfill Management Option, Journal of the
Environmental Engineering Division-ASCE, 106(6), 1057-1069.
Pohland, F. G. and Kim, J. C., 1999, In situ anaerobic treatment of leachate in landfill
bioreactors, Water Science and Technology, 40(8), 200-210.
Pokhrel, D. and Viraraghavan, T., 2004, Leachate generation and treatment A review,
Fresenius Environmental Bulletin, 13(3B), 223-232.


144
aerobic lysimeters, BOD concentration of lysimeter 4 was still approximately 35,000
mg/L. For the aerobic lysimeters, organic carbon dissolved in the leachate degraded
rapidly, resulting in treatment of the leachate with respect to organic carbon. The
settlement data of the aerobic lysimeters (Chapter 5) showed that constant settlement was
observed even after the BOD concentration was below 50 mg/L. This observation
indicated that waste could be continuously decomposed maintaining low BOD
concentration. Consequently, air addition may contribute to lower leachate treatment cost
as well as more waste decomposition.
Another important finding of this research lies is that alkaline pH (pH > 9.0) and
more oxidizing redox conditions can be formed by air addition. Under the oxidizing and
alkaline conditions in the aerobic lysimeters, there were greater concentrations of Al, Cr,
Cu and Pb in leachate of the aerobic lysimeters relative to those in the anaerobic
lysimeters among the eight metals under considerations (Chapter 3). A greater impact of
high pH on an increase in the solubilities of Al and Cr was observed. The high pH was
not reported from all research studying waste decomposition under aerobic conditions.
However, air addition may increase the potential to raise the pH by CO2 stripping as
discussed in Chapter 2. Among the leaching behaviors of the metals under the relatively
high pH, it is notable that the oxidation state of Cr can be converted into the hexavalent
form in aerobic landfill condition in the presence of CCA-treated wood. Song et al.
(2005) reported that Cr leached from CCA-treated wood could be oxidized to hexavalent
form under alkaline conditions. It was reasonable for Cr species to oxidize to hexavalent
form under the oxidizing condition. However, although the manufacturing of CCA-
treated wood is banned currently for most uses, it has been recognized as a safe material


228
Tipping, E., 2005, Modelling A1 competition for heavy metal binding by dissolved
organic matter in soil and surface waters of acid and neutral pH, Geoderma, 127(3-
4), 293-304.
Townsend, T. G., Miller, W. L., Lee, H. J. and Earle, J. F. K., 1996, Acceleration of
landfill stabilization using leachate recycle, Journal of Environmental Engineering-
ASCE, 122(4), 263-268.
Townsend, T., Tolaymat, T., Solo-Gabriele, H., Dubey, B., Stook, K. and Wadanambi, L.,
2004, Leaching of CCA-treated wood: implications for waste disposal, Journal of
Hazardous Materials, 114(1-3), 75-91.
Ugwuanyi, J. O., Harvey, L. M. and McNeil, B., 1999, Effect of process temperature, pH
and suspended solids content upon pasteurization of a model agricultural waste
during thermophilic aerobic digestion, Journal of Applied Microbiology, 87(3),
387-395.
United States Department of Agriculture (USDA), 1999, Wood handbook, FPL-GTR-
113, Madison, WI, USA.
United States Environmental Protection Agency (USEPA), 1996, Test Methods for
Evaluating Solid Waste, SW-846, 3r ed.; Office of Solid Waste: Washington, D. C.,
USA.
United States Environmental Protection Agency (USEPA), 1997, Emission factor
documentation for AP-42 section 2.4 municipal solid waste landfills, Office of
Air Quality Planning and Standards and Office of Air and Radiation, North
Carolina, USA
United States Environmental Protection Agency (USEPA), 1999, Water quality criteria;
notice of availability; 1999 update of ambient water quality criteria for ammonia,
Federal Register vol. 64 no.245, Washington D.C., USA.
United States Environmental Protection Agency (USEPA), 2002, Municipal solid waste
in the United States: 2000 facts and s, EPA530-R-02-001, Office of solid waste and
emergency response, Wahington D. C., USA.
United States Environmental Protection Agency (USEPA), 2003, Municipal solid waste
in the United States: 2003 facts and s, EPA 530-R-03-011, Office of solid waste
and emergency response, Wahington D. C., USA.
United States Environmental Protection Agency (USEPA), 2003, Municipal Solid Waste
Generation, Recycling, and Disposal in the United States Facts and s for 2003,
U.S.EPA, Washington D.C., USA.
United States Environmental Protection Agency (USEPA), 2003, National Primary
Drinking Water Standards, EPA epa816-f-03-016, Office of Water, Washington D. C.,
USA.


T/§UI IV q/Sui [Y
72
too
10 -
AEROBIC
Lys 1
O Lys 2
CD
O
O O
o
o
o
o 8* cy.
8 0o oo*o
o
0.1
100
200
300
400
Figure 3-1. Changes of A1 concentrations over time


Sulfate (mg/L) Sulfate (mg/L)
42
a
-o
3
C/3
10
- 8
- 6
L 0
Days
}
u
T3
3
cn
r 10
- 8
- 6
- 4
- 2
L 0
Days
Figure 2-12. The changes in sulfate and sulfide versus time in the presence of dissolved
oxygen
Dissolved oxygen (mg/L) X Dissolved oxygen (mg/L) -


Fe, mg/L Fe, mg/L
77
1000
0 50 100 150 200 250 300 350
Figure 3-6. Changes of Fe concentrations over time


115
wood NP
5% 6%
(A)
OP
6% CB
(B)
Figure 4-3. The changes in the percentage of waste components after decomposition; (A)
raw waste components and (B) decomposed waste (aerobic)


11
was added to make up for the amount of leachate used for analysis. Table 2-2 summarizes
the parameters and methods used for each analysis.
Biogas samples generated from both the aerobic and anaerobic lysimeters were
collected and analyzed for methane, carbon dioxide, and oxygen. For the aerobic
columns, the gas volume was measured using a wet-tip gas meter. For the anaerobic
lysimeters, the gas was gathered in 5-L and 10-L air-sampling bags, and the volume
contained in the bags was measured using the water-gas replacement method (see Figure
B-4). A LANTEC GEM 500 (SAIC, San Diego, CA) gas meter was used for gas analysis
for both the aerobic and anaerobic lysimeters. Additionally, gas samples collected from
the anaerobic lysimeters were analyzed for CH4 and CO2 using a gas chromatograph
equipped with a GS-Carbon plot column (Agilent Technology, Palo Alto, CA) to confirm
the measurements analyzed by LANTEC GEM500 gas meter.
2.2.6 Recovery of the Anaerobic Lysimeters
Since both of the anaerobic lysimeters (lys 3 and 4) remained in an acidic condition
(pH < 6) for 500 days, 100 g of sodium bicarbonate was added as a buffer to the top of
each lysimeter on day 300. The pH of the top part of the lysimeters changed to neutral,
but the pH of leachate collected from the bottom port remained low (5 to 5.5). The pH of
the leachate from lysimeter 4 increased to pH 7 from day 400. Since only minimum
changes in leachate pH of the lysimeter 3 were observed after buffer addition, additional
sodium bicarbonate was added to the bottom port rather than to the top of the lysimeters;
a total of lOOg of sodium bicarbonate was added (20 g each were added on days 420,
453, 469, 532, and 555 again). Only a temporary increase in pH was observed after this
addition. As a next step in increasing pH, lab air was injected into lysimeter 3 on day 627.
Before air injection, the methane concentration of the lysimeter 3 was 35%, and the pH of


Cumulative metals released out of the lysimeters (mg)
93
Mass loss, %
Figure 3-17 (continued). Changes in cumulative mass of metal released over a mass loss,
%


Methane yields (L Q\{^lg VS)
117
0.5
0.4
0.3 -
0.2 -
0.1
0.0
*
i
I I Raw waste
V////A Anaerobic
K888881 Aerobic
ii
ni
idfe
officepaper cardboard newspaper wood
(A)
0.035
0.030
> 0.025
SC
0.020
2
13
0.015
C
C3
-g
3a 0.010
0.005
0.000
Figure 4-5. Methane yields and weight differences of lignocellulosic materials among
raw and two lysimeters (A) all lignocellulosic materials; (B) wood only


Mass loss (g) Mass loss (g)
158
Days
(C)
Days
(D)
Figure A-2. Waste mass loss by TOC and gas generation
(j
'Si)
y
o
H


22
clear to describe the differences between this and other studies by comparing the
concentrations of carbonate ions.
2.4.3 Implications for Full-scale Application
Unlike the lab-scale simulated landfill, it is extremely difficult to aerate an entire
large-scale landfill. Highly compacted wastes make it difficult for an air stream to
penetrate into the recesses of a landfill. Moreover, leachate characteristics resulted from
air addition may be variable. As the analytical results have shown, leachate
characteristics of lysimeters 1 and 2 were different, despite starting with the same waste
stream and the same operational condition. The leachate characteristics of lysimeter 1
were similar to those of the anaerobic lysimeters during the first 180 days showing great
concentration of organic matter despite air addition. This was because of the large
anaerobic zones formed at the bottom of the lysimeter by improper air addition to the
bottom.
Aerobic zones can be formed around air injection wells but anaerobic zones may
still be present in the same landfill. However, coexistence of the aerobic and anaerobic
zones can be used for recovery of acid-stuck sour landfills. In this research, the air
addition was conducted under the hypothesis that environments formed by aeration for a
short period can be favorable to anaerobic microorganisms. A great amount of VFAs,
which caused acidic conditions, may be rapidly consumed by aerobes living in a
relatively wide pH range. Conversion of carbonic acid (HaCCV) to CO2 caused by air
stripping may increase the pH. With air addition with low flow rate, the anaerobic zones
may be protected from oxygen intrusion because oxygen may be depleted by the
respiration of aerobes. An additional technical strategy would be to add buffer such as
lime along with air addition. Buffer added may increase the alkalinity concentration.


223
Holt, D. M. and E. B. G. Jones, 1983, Bacterial Degradation of Lignified wood cell wall
in anaerobic aquatic habitats, Applied and Environmental Microbiology, 46(3),
722-727.
Holtz, R. D. and Kovacs, W. D., 1981, An introduction to geotechnical engineering.
Englewood Cliffs, N.J., Prentice-Hall.
Jain P., Kim H., and Towsend, T. G., 2005, Heavy metal content in soil reclaimed from a
municipal solid waste landfill, Waste Management 25(1), 25-35.
Jambeck, J., 2004, The disposal of CCA-treated wood in simulated landfills: potential
impacts, Doctoral dissertation, University of Florida
Jang, Y. C. and Townsend, T. G., 2003, Leaching of lead from computer printed wire
boards and cathode ray tubes by municipal solid waste landfill leachates,
Environmental Science & Technology, 37(20), 4778-4784.
Jang, Y. C. and Townsend, T. G., 2003, Effect of waste depth on leachate quality from
laboratory construction and demolition debris landfills, Environmental Engineering
Science, 20(3), 183-196.
Jansen, B., Nierop, K. G. J. and Verstraten, J. M., 2003, Mobility of Fe(II), Fe(III) and A1
in acidic forest soils mediated by dissolved organic matter: influence of solution pH
and metal/organic carbon ratios, Geoderma 113(3-4), 323-340.
Jerger. D., and Tsao, T. G, 1987, Feed composition in anaerobic digestion of biomass,
Chynoweth, D. P. and R. Isaacson edited. London, New York, Elsevier Applied
Science.
Kayhanian, M., 1995, Biodegradability of the organic fraction of municipal solid-waste
in a high-solids anaerobic digester, Waste Management & Research, 13(2), 123-
136.
Kjeldsen, P.,Barlaz, M. A., Rooker, A. P., Baun, A., Ledin, A. and Christensen, T. H,
2002, Present and long-term composition of MSW landfill leachate: A review.
Critical Reviews in Environmental Science and Technology 32(4), 297-336.
Komilis, D. P. and Ham, R. K., 2003, The effect of lignin and sugars to the aerobic
decomposition of solid wastes, Waste Management, 23(5), 419-423.
Komilis, D. P., Ham, R. K. and Stegmann, R., 1999, The effect of municipal solid waste
pretreatment on landfill behavior: a literature review, Waste Management Research,
17, 10-19.
Krogmann, U. and Woyczechowski H., 2000, Selected characteristics of leachate,
condensate and runoff released during composting of biogenic waste, Waste
Management & Research 18(3), 235-248.


108
For these reasons, it is proposed to conduct research on decomposition of woody
materials under air addition into the C&D debris landfill as a future research.
4.5 Conclusions
Decomposed waste was excavated from lysimeters 2 (aerobic) and 4 (anaerobic)
after a test period (1 and 2 years for aerobic and anaerobic lysimeters, respectively). The
mass of waste lost by waste decomposition of the aerobic lysimeter was greater than from
the anaerobic lysimeter. The mass losses measured and predicted by gas generation were
very similar. Due to the overburden pressure applied to the top layer of the lysimeters, the
greatest moisture content was observed from the top layer of both aerobic and anaerobic
lysimeters.
The overall methane yield (a measure of the degree of waste decomposition
occuring) of the aerobic lysimeter was lower than that of the anaerobic lysimeter despite
a shorter test period. The greatest difference of methane yields between raw and
decomposed waste was observed from office paper. This difference decreased following
the order: office paper > cardboard > newspaper > wood. Though cellulose
concentrations of decomposed SYP between the aerobic and anaerobic lysimeters were
statistically the same, the methane yields of SYP blocks excavated from the aerobic
lysimeter were significantly lower than those of the anaerobic lysimeter. The lower
methane yields of newspaper were also observed from the aerobic lysimeters. However,
lignin degradation was not observed from either aerobic or anaerobic lysimeters.


5. LANDFILL SETTLEMENT BEHAVIOR WITH WASTE DECOMPOSITION 119
5.1 Introduction 119
5.2 Materials and Methods 120
5.2.1 Lysimeters 120
5.2.2 Application of Overburden Pressure 121
5.2.3 Compression Index and Phase Separate Method 122
5.2.4 Estimation of Mass Loss 123
5.2.5 Volume Loss versus Mass Loss 124
5.3 Results 125
5.3.1 Settlement Behavior over Time 125
5.3.2 The Relationship between The Settlement and Mass Loss 126
5.3.3 Ultimate Settlement 127
5.4 Discussion 127
5.4.1 Compression Index 127
5.4.2 Correlation of Mass Loss and Volume Loss 128
5.4.3 Application 129
5.5 Conclusions 131
6. SUMMARY AND CONCLUSIONS 141
6.1 Summary 141
6.2 The Implication of This Research 143
6.3 Conclusions 145
6.4 Future Work 147
APPENDIX
A. ADDITIONAL PROCEDURES AND CONCEPTS 149
A. 1 Prediction of Mass Loss by Gas and Leachate 149
A.2 Estimation of Biodegradable Volatile Solids (BVS) 152
A.3 Lysimeter Dismantlement 153
B. SUPPLEMENTAL FIGURES 159
C. LYSIMETER EXPERIMENT RAW DATA AND GRAPHS 169
C.l Graphs 169
C.2 Raw Data 189
LIST OF REFERENCES 219
BIOGRAPHICAL SKETCH 231
vii


C-2. The change in BOD5 of the aerobic and anaerobic lysimeters over the percentage
of mass loss 170
C-3. The change in ammonia of the aerobic and anaerobic lysimeters over time 171
C-4. The change in fluoride of the aerobic and anaerobic lysimeters over time 172
C-5. The change in chloride (Cl) of the aerobic and anaerobic lysimeters over time 173
C-6. The change in sulfate of the aerobic and anaerobic lysimeters over time 174
C-7. The change in calcium (Ca) of the aerobic and anaerobic lysimeters over time 175
C-8. The change in sodium (Na) of the aerobic and anaerobic lysimeters over time 176
C-9. The change in biogas produced from the aerobic lysimeters 177
C-10. The change in biogas produced from the anaerobic lysimeters 178
C-l 1. A1 concentration versus pH in leachate from the lysimeters 179
C-12. Cr concentration versus pH in leachate from the lysimeters 180
C-l3. Cu concentration versus pH in leachate from the lysimeters 181
C-14. Mn concentration versus pH in leachate from the lysimeters 182
C-l5. Pb concentration versus pH in leachate from the lysimeters 183
C-l6. Zn concentration versus pH in leachate from the lysimeters 184
C-l 7. Change in methane yields of the waste layer 2-1 and 2-2 185
C-l 8. Change in methane yields of the waste layer 2-3 and 2-4 186
C-l 9. Change in methane yields of the waste layer 4-1 and 4-2 187
C-20. Change in methane yields of the waste layer 4-3 and 4-4 188
xiii


195
Table C-3 (continued)
date
lys 1
lys 2
date
lys 3
lys 4
4/3/2005
16500
16500
4/10/2005
15500
15150
4/17/2005
15750
16050
4/24/2005
13650
15750
4/30/2005
14550
14850
5/7/2005
14700
15000
5/16/2005
13200
14250
5/23/2005
13050
13500
5/31/2005
12000
12750
6/6/2005
13200
13350
6/14/2005
13800
14250
6/28/2005
13800
12900
7/12/2005
13500
12750
7/27/2005
12600
12300
8/10/2005
13050
12300


148
benefits of aerobic landfills. However, in order to apply the air injection technique to the
field, much work still needs to be done in order to solve the problems that may
potentially occur.
For landfill gas, closure of the flare system due to air addition may allow landfill
gas to emit into the atmosphere without filtration. Thus it is important to identify these
gas constituents and to evaluate the potential impact of these gases on the environment.
Moreover, it is required to explore the potential risk of explosion of landfill gas by
mixing certain concentrations of oxygen and methane (Coward and Jones, 1952).
All landfills may not have the problem with clogging of leachate collection system
(LCS), but the failure of the LCS was reported due to biological and chemical reactions
(Fleming et al., 1999). There is also potential risk of biological clogging on the geotextile
and LCS of the aerobic landfills (Reinhart and Chopra, 2000). Further research is needed
to assess the chemical and biological clogging possibly occurred in aerobic landfills in
comparison with anaerobic landfills.
It may be also required to assess the possible impact of rapid settlement of aerobic
landfills on the stability of an entire landfill. The examples of the potential risks may
include 1) the collapse or breakage of gas collection system and geomembrane on the top
of a landfill, and 2) the failure of landfill slope due to unbalanced settlement which may
occur in air injection areas.


209
Table C-12 Volatile fatty acids (VFA) of lysimeter 4 (unit: mg/L)
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/7/2003
2694.5
243.4
0.0
97.1
8/13/2003
5911.9
216.3
1323.3
2605.4
8/15/2003
5931.9
206.6
1346.7
2585.0
8/19/2003
4163.9
220.4
1317.0
2776.7
8/22/2003
1668.6
193.0
1476.3
3338.6
8/26/2003
1793.1
187.8
1473.3
3285.1
9/9/2003
2880.9
171.0
1686.9
3271.3
10/8/2003
5258.6
759.0
3006.2
4820.7
10/15/2003
6366.8
743.4
3018.9
4803.3
10/22/2003
7897.2
761.1
3205.4
4977.5
11/5/2003
7383.7
3775.5
3558.7
5437.6
11/15/2003
8051.3
3653.4
3452.6
5241.6
11/19/2003
8961.9
3837.7
3580.4
5374.4
11/28/2003
9574.0
3973.5
3703.3
5476.4
12/10/2003
10036.6
4004.2
3754.1
5496.6
12/19/2003
6580.8
3043.4
3907.1
10782.7
1/5/2003
8193.8
3405.5
4473.5
11310.0
1/8/2003
7685.9
3100.9
3937.7
10264.8
1/14/2003
7559.2
2617.0
3356.1
8451.9
2/13/2003
8839.9
2995.2
4048.9
9055.6
3/16/2004
9964.5
3469.7
4734.5
10038.3
7/17/2004
15855.6
3891.7
3845.4
4636.0
8/18/2004
22080.7
3290.6
3601.4
5822.7
9/22/2004
12112.4
2824.2
4058.2
6519.4
9/30/2004
11624.9
2706.9
3989.4
6413.2
10/21/2004
14603.2
3118.2
4725.8
7205.4
11/3/2004
12304.7
2711.0
3912.8
6141.4
12/8/2004
13600.4
3576.6
3977.6
5854.4
12/12/2004
11090.6
3256.8
3250.8
5184.9
12/20/2004
12577.5
3733.5
3017.6
5102.5
1/22/2005
13159.3
3351.2
2018.2
3570.5
1/30/2005
16013.2
4858.1
2757.5
4242.8
2/8/2005
16170.3
5019.1
2499.3
4130.4
2/16/2005
15610.6
5419.8
2553.6
4078.9
3/3/2005
11124.0
5202.6
1615.2
2950.1
3/12/2005
10597.2
4919.0
1298.3
2249.0
3/25/2005
8711.5
4910.3
1086.6
1672.3
4/3/2005
8461.1
5422.0
1171.0
1735.4
4/17/2005
9927.2
5086.0
1039.4
1864.3
4/24/2005
8648.7
4514.9
856.9
1490.3
4/30/2005
8194.4
4767.2
937.3
1453.1
5/7/2005
6624.6
4145.4
776.0
1144.7
5/16/2005
8062.8
5347.5
942.1
1201.3
5/23/2005
6288.2
4771.3
839.7
894.0
5/31/2005
4314.8
3642.3
630.1
498.2
6/6/2005
5017.1
4667.2
773.4
546.1


212
Table C-14. Metal concentrations of lysimeter 1 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
7/28/2004
2.27
0.08
0.02
0.32
110.94
3.64
0.03
28.58
8/29/2004
2.94
0.24
0.12
3.31
90.13
4.03
0.75
97.71
9/2/2004
1.28
0.45
0.18
0.11
7.47
2.30
0.03
95.27
9/3/2004
2.86
0.39
0.10
2.05
74.80
4.25
0.27
132.47
9/10/2004
4.45
0.59
0.16
8.81
115.63
6.24
0.79
209.81
9/15/2004
3.22
0.42
0.10
8.91
79.81
4.33
0.63
172.08
10/7/2004
3.96
0.43
0.10
82.61
5.15
0.55
214.32
10/14/2004
6.73
0.39
0.13
0.99
47.40
6.65
0.12
244.32
10/19/2004
4.84
0.25
0.09
6.50
29.69
5.34
0.58
214.90
11/6/2004
4.33
0.28
0.11
7.23
65.77
6.99
0.46
203.98
11/24/2004
2.93
0.19
0.07
6.45
50.69
5.45
0.50
137.60
12/19/2004
2.05
0.16
0.04
1.49
89.08
2.84
0.08
82.72
1/6/2005
1.17
0.11
0.04
0.24
45.10
4.12
0.03
69.49
1/21/2005
0.80
0.12
0.05
0.51
2.31
1.10
0.00
5.52
1/30/2005
3.45
0.26
0.11
1.32
1.68
0.08
0.02
2.55
2/5/2005
7.68
0.30
0.17
4.48
2.74
0.09
0.03
6.46
2/16/2005
5.47
0.32
0.11
4.79
1.08
0.04
0.01
4.09
2/24/2005
6.76
0.37
0.12
5.56
1.29
0.03
0.04
3.83
3/1/2005
6.91
0.31
0.10
3.42
0.90
0.03
0.01
3.52
3/12/2005
6.44
0.35
0.11
2.90
1.13
0.03
0.01
4.09
3/20/2005
7.49
0.37
0.13
2.14
1.71
0.04
0.01
5.00
3/26/2005
10.28
0.40
0.20
3.64
2.21
0.08
0.06
4.96
4/3/2005
11.50
0.50
0.22
1.73
2.09
0.06
0.02
6.83
4/17/2005
11.80
0.71
0.33
1.79
4.85
0.12
0.06
10.83
4/24/2005
11.77
0.59
0.27
1.65
3.58
0.08
0.05
9.43
4/30/2005
12.19
0.74
0.43
1.79
6.03
0.14
0.06
14.88
5/7/2005
14.05
0.80
0.43
2.17
5.31
0.11
0.05
15.03
5/16/2005
13.82
0.77
0.45
1.85
6.56
0.14
0.05
16.79
5/31/2005
10.74
0.72
0.45
0.91
8.14
0.19
0.03
13.70
6/6/2005
10.79
0.72
0.44
0.60
7.12
0.18
0.02
14.27
6/14/2005
10.81
0.55
0.37
0.61
6.94
0.19
0.02
13.13
6/28/2005
9.52
0.38
0.32
1.13
7.67
0.17
0.03
11.85
7/5/2005
8.40
0.30
0.26
1.31
7.42
0.18
0.04
11.31
7/11/2005
9.02
0.30
0.23
0.95
6.62
0.18
0.02
10.70
7/27/2005
8.12
0.25
0.26
0.60
6.17
0.14
0.02
10.31
8/10/2005
9.37
0.29
0.28
0.79
6.82
0.14
0.02
10.56


59
3.3.2 Organic Wastes as Absorbents of Heavy Metals
Analytical results of the 8 metals absorbed on office paper (OP), cardboard (CB),
newspaper (NP), and wood blocks (WD) are shown on Figure 3-12. Concentrations of Al,
As, and Cu adsorbed on the lignocellulosic materials of the aerobic lysimeter appeared
greater than those of the anaerobic lysimeter.
Heavy metals adsorbed on solid wastes from the aerobic and anaerobic lysimeter
were statistically analyzed using the ANOVA test. Test results are presented in appendix
C. All metals, except for Pb, absorbed on CB in the aerobic lysimeter and were
significantly higher than those of the anaerobic lysimeter. Other significant differences
between the aerobic and anaerobic lysimeters were found with Al, As, Mn and Cu
adsorbed on NP, OP and WD.
Figure 3-13 depicts the differences of total mass of metals adsorbed on
lignocellulosic materials between the aerobic and anaerobic lysimeters. These values
were calculated by multiplying the metal concentrations by the mass of each waste
obtained from the garbage separation. Interestingly, the observed trends of adsorption of
some metals did not to correspond with their leaching trends. These trends can be found
from adsorption trends of As, Mn and Pb. The amounts of metals leached and adsorbed
between aerobic and anaerobic lysimeters are compared in Table 3-5. These results
indicate that metal adsorption may be influenced by environmental conditions such as
pH, redox, and the presence of other ligands. Ravat et al (2000) reported that the binding
of selected metals (Zn, Cu, and Pb) on lignocellulosic materials is strongly pH-dependent
in the absence of interference from other ligands. Adsorption of Fe on organic matter is
mainly controlled by the oxidation states of Fe; Fe (III) has greater affinity for organic
matter than Fe (II) does (Jansen et al., 2003), corresponding to the large differences of Fe


Gas concentrations (%)
44
Days
Figure 2-14. Changes in gas concentrations of anaerobic lysimeter 4


68
aerobic and anaerobic lysimeters. Among the 8 metals evaluated (Al, As, Cr, Cu, Pb, Mn,
Fe and Zn), the concentrations of Al, Cu, Cr and Pb in leachate of the aerobic lysimeters
were significantly greater than those of the anaerobic lysimeters, and the average
concentrations of As, Fe, Mn, and Zn in the anaerobic lysimeters were significantly
greater in concentration than observed in the anaerobic lysimeters.
After a test period, one each of the aerobic and anaerobic lysimeters was
dismantled and the metals adsorbed on decomposed lignocellulosic waste were analyzed.
Greater concentrations (mg metal/kg waste) of Fe, Mn, As, Al and Cu were found from
the aerobic lysimeters. All metals, except for Pb, adsorbed on the cardboard in the
aerobic lysimeter and were significantly higher than those of the anaerobic lysimeter.
Through this research, aerobic landfills were found to have a greater potential to release
several metals through leachate than that of anaerobic landfills; aerobic landfills have
greater Al, Cr, Cu and Pb leaching potential due to oxidizing condition. Experimental
results confirmed this notion. In the presence of CCA-treated wood and electronic waste
(CRT monitor glass), high As and Pb concentrations were observed from the anaerobic
lysimeter and the aerobic lysimeter during the acid phase, respectively.


51
< 8. Average Al concentrtaions (7.9 mg/L) of the aerobic lysimeters were significantly
higher than those of the anaerobic lysimeters (0.28 mg/L).
Generally, A1 leaching is not greatly affected by redox conditions but mainly
controlled by pH. Meima and Comans (1997) reported that A1 solubility was low in the
pH range 6 to 7, which corresponds to the A1 results in the aerobic condition presented in
Figure 3-1. Among many ligands forming A1 compelxation, hydroxide ion (OH ) is
known as a major ion to control the solubility of A1 in aquatic systems. The equilibrium
of A1 with gibbsite (Al(OH)3)), an Al-OH complex, is characterized by a U-shaped pH-
leaching curve (Eary, 1999). However, A1 concentrations in anaerobic lysimeters did not
appear to follow with the solubility of gibbsite. Besides pH and redox conditions, large
differences in leachate characterisitcs between aerobic and anaerobic lysimeters included
the high organic content and anions such as floride and sulfate in the leachate of the
anaerobic lysimeters. The most likely explanation of low A1 solubility in the anaerobic
conditions is complexation of A1 and organic matter. Tipping (2005) reported that A1
solubility is strongly associated with organic content. Skyllberg (2001) also reported that
high dissolved organic content made A1 solubility significantly decrease. Therefore, low
solubilitity of A1 during the first phase of the anaerobic lysimeters is the result of
complexation of A1 with high concentrations of organic matter.
3.3.1.2 Arsenic
Figure 3-2 depicts the change in As over time. High As concentrations were
observed from the anaerobic lysimeter before day 220. The greatest As concentration was
3.2 mg/L on the day 89. The As concentration then lowered below 1.5 mg/L after day
400. For lysimeter 4, As concentrations were continuously low, showing the lowest value,
0.27 mg/L on the last sampling day (day 741). In contrast to the anaerobic lysimeters, As


Total mass of metals adsorbed on lignocellulosic materials (mg)
86
600
500
400
300
200
100
0
14000
12000
10000
8000
6000
4000
2000
0
Figure 3-13. (continued)


227
Science Applications International Corporation (SAIC), 2000, Characterization and
evaluation of landfill leachate.,Draft, ,EPA Contract 68-W6-0068,, Arlington, VA,
USA.
Sheridan, S. (2003), Modeling solid waste settlement as a function of mass loss, Masters
thesis, University of Florida, Gainesville, FL, USA.
Shukla, S. R. and Pai, R. S., 2005, Adsorption of Cu(II), Ni(II) and Zn(II) on dye loaded
groundnut shells and sawdust, Separation and Purification Technology, 43(1), 1-8.
Skyllberg, U., Raulund-Rasmussen, K. and Borggaard, O. K., 2001, pH buffering in
acidic soils developed under Picea abies and Quercus robur effects of soil organic
matter, adsorbed cations and soil solution ionic strength, Biogeochemistry, 56(1),
51-74.
Snoeyink, V. L. and Jenkins, D., 1980, Water chemistry. New York, Wiley.
Solid Waste Association of North America (SWANA), 2003, The environmental
consequences of disposing of products containing heavy metals in municipal solid
waste landfills, SWANA, Silver Springs, MD, USA.
Sowers, G. F., 1973, Settlement of waste disposal fills. Soil mechanics and foundation
engineering conference, Minneapolis, MN., USA.
Stegmann, R., 1983, New Aspects on Enhancing Biological Processes in Sanitary
Landfill, Waste Management & Research, 1, 201-211.
Stessel, R. I. and Murphy R. J., 1992, A lysimeter study of the aerobic landfill concept
Waste Management Research, 10,485-503.
Stinson, J. A. and Ham, R. K., 1995, Effect of lignin on the anaerobic decomposition of
cellulose as determined through the use of a biochemical methane potential method,
Environmental Science & Technology, 29(9), 2305-2310.
Sublet, R., Simonnot, M. O., Boireau, A. and Sardin, M., 2003, Selection of an adsorbent
for lead removal from drinking water by a point-of-use treatment device, Water
Research 37(20), 4904-4912.
Summerfelt, S. T., Davidson, J. and Waldrop, T., 2003, Evaluation of full-scale carbon
dioxide stripping columns in a coldwater recirculating system, Aquacultural
Engineering, 28(3-4), 155-169.
Tchobanoglous, G., Theisen, H. and Vigil, S. A., 1993, Integrated solid waste
management: engineering principles and management issues, New York, McGraw-
Hill.


LIST OF FIGURES
Figure Page
2-1. Schematic of the lysimeter 30
2-2. The composition of fabricated municipal solid waste for this research 31
2-3. Comparison of pH between aerobic and anaerobic lysimeters versus time 32
2-4. Changes in COD of aerobic and anaerobic lysimeters versus time 33
2-5. Changes in BOD of aerobic and anaerobic lysimeters versus time 34
2-6. Changes in VFAs of aerobic and anaerobic lysimeters versus time (A) acetic acid
only and (B) acetic acid, propionic acid and butyric acid 36
2-7 Changes in the ratio BOD/COD of the aerobic and anaerobic lysimeters over time ..37
2-8. Changes in ammonia concentrations versus time 38
2-9. Changes in TDS of the aerobic and anaerobic lysimeters versus time 39
2-11. Changes in sulfide and pH versus time 41
2-12. The changes in sulfate and sulfide versus time in the presence of dissolved
oxygen 42
2-13. The changes in air injection rate and gas concentrations of aerobic lysimeter 43
2-14. Changes in gas concentrations of anaerobic lysimeter 4 44
2-15. Cumulative biogas vs. days in aerobic and anaerobic lysimeters 45
2-16. Changes in gas concentrations, pH and gas generation rate after air injection into
lysimeter 3 46
3-1. Changes of A1 concentrations over time 72
3-2. Changes of As concentrations over time 73
3-3. Changes of Cr concentrations over time 74
x


Settlement, %
136
Days
Figure 5-3. Settlement behaviors and compression coefficients of aerobic and anaerobic
lysimeter over a period of time


Cumulative methane volume (mL) Cumulative methane volume (mL)
188
Figure C-20. Change in methane yields of the waste layer 4-3 and 4-4


15
rapidly down to below 100 mg/L from day 200 while BOD values of the anaerobic
lysimeters decreased relatively slowly.
The primary contributor to high COD or BOD concentrations in landfill leachate
is volatile fatty acids (McBean et al., 1995). Figure 2-6 (a) depicts the changes in acetic
acid, one of the major volatile fatty acids (VFA), as a function of mass loss. Acetic acid is
used as a substrate by methanogenic bacteria and contributes to the formation of an acidic
environment when they are unbalanced with the growth of methanogenic bacteria. For
these reasons, VFA concentrations are used as an indicator to assess landfill conditions
(USEPA, 2004). For example, the decrease in acetic acid concentration in lysimeter 3 and
4 corresponds to the point when the pH began to rise. In the aerobic lysimeters, high
concentrations of acetic acid were noted during the first phase of the experiment due to
improper air distribution as discussed earlier. Acetic acid in leachate from the aerobic
lysimeters was degraded to less 1 mg/L by day 200, which corresponds with the time
required to deplete COD and BOD.
Among different types of short carbon chain fatty acids, acetic, propionic and
butyric acids are known as major VFAs that are involved in biodegradation processes in
anaerobic conditions. Production and degradation of these major VFAs in selected
aerobic and anaerobic lysimeters are presented in Figure 2-6 (b). In both aerobic and
anaerobic lysimeters, the concentration of VFAs was mainly: acetic acids > butyric acids
> propionic acids. These results are similar to those found by Parawira et al (2004). They
also explained that high butyric acids were mainly attributed to high carbohydrates in
waste. Under the same condition, the degradation of VFAs in anaerobic condition was
found to be in the following order: butyric acids > acetic acids > propionic acids. Wang et


126
waste decomposition ((Ca)max phase). The occurrence of biological activity during the
settlement can be also confirmed by an increase in cumulative gas over time (Figure 5-3).
For lysimeter 4, a rapid settlement curve was observed after a long lag period. It is noted
that the depth difference (AH) of lysimeter 4 during the (Ca)max phase is close to AH of
lysimeter 1 during the same phase. This result indicates that the aerobic landfill may
show the better settlement rate as opposed to the anaerobic landfill within a very short
period of time. However, great settlement performance was shown in both aerobic and
anaerobic lysimeters during the (Ca)max phase rather than (Ca)min phase. Therefore, it is
concluded that waste decomposition plays an important role for landfill settlement under
both aerobic and anaerobic conditions.
5.3.2 The Relationship between The Settlement and Mass Loss
Figure 5-4 depicts the relationship between the percentage of overall settlement and
mass loss. Since lysimeter 3 remained in the acid phase for 600 days, the percentage of
mass loss and its settlement at the end of a test period were substantially lower than other
three lysimeters. It is noted that the relationship between settlement and the overall mass
loss of each lysimeter correlated well (r = 0.88) despite their different overall mass loss
percentage. This can be confirmed by plotting all mass loss and settlement data of the
four lysimeters.
Figure 5-5 depicts the changes in the lysimeter depth over percent of mass loss
correlated (a) linearly and (b) semi-logarithmically. Comparing the relation coefficient
(r ) between two plots, semi-logarithmic correlation showed better fit between mass loss
and settlement (r2 = 0.89). This relation can be mathematically expressed as follows:
[settlement, %] = (AH, %) = 16.90 log [mass loss, %] 6.24 (4)


67
upon air intrusion. Hexavalent chromium is a highly toxic metal causing decreased
pulmonary function and pneumonia (Bradle, 2005). On the contrary, toxicity of As can be
reduced by air injection. In oxidizing conditions, Arsenate can be oxidized to As (VI),
which is less toxic. Since As (VI) is less mobile and has a low solubility, the total amount
of As leached can be reduced. The influence of air injection on the amount of metals can
be more clearly understood by comparing the percentage of As, Cr and Cu (Table 3-4).
Under the scenario of high As content in leachate caused by co-disposal of fly ash or
ground CCA-treated wood, it would be recommended to inject air in order to reduce As
toxicity and the amount leached. However, Cr (VI) in an aerobic landfill can be present
as an ionic form under oxidizing conditions, and the cumulative mass may increase over
a period of time. For these reasons, it is necessary that landfill personnel develop a
strategy to optimize the influence of air injection.
Based on the metal results, the metal leaching behavior of old landfill metals can be
predictable. Kjeldsen et al. (2002) proposed that the condition of old landfills would be
aerobic due to air intrusion. Once air intruded in anaerobic landfills, metal leaching may
occur by the dissociation of the metals that adsorbed on decomposed organic matters.
Among 8 metals under consideration, Pb would be the most leachable metal. As Figure 3-
13 shows a large amount of Pb was adsorbed on organic matter. Bozkurt et al. (1999)
explained that the pH of the landfill would change to acid again after the long-term
process. This was due to the oxidation of sulfate and organic matter. Therefore, the
moderate oxidizing condition, and low pH, would accelerate Pb leaching.
3.5 Conclusions
Research on the fate of metals in simulated aerobic and anaerobic landfills was
conducted. The leaching behavior of selected metals was significantly different between


199
Table C-6 Chemical oxygen demand (COD) of aerobic and anaerobic lysimeter (unit:
r mg/L) ,
Date
Lys 1
Lys 2
Date
Lys 3
Lys 4
7/28/2004
6950
11250
8/8/2003
64745
66941
8/12/2004
21900
27850
8/13/2003
77801
62846
8/17/2004
34050
26600
8/15/2003
59820
58277
8/24/2004
22700
21200
8/19/2003
78988
82964
9/3/2004
64950
24400
8/26/2003
69671
77267
9/10/2004
69250
19600
9/2/2003
72876
74656
9/15/2004
66750
18900
9/12/2003
72876
77030
9/23/2004
66850
19250
9/19/2003
73593
83079
10/5/2004
66249
13464
10/3/2003
67983
81600
10/26/2004
61302
5814
10/10/2003
65943
79560
11/3/2004
70788
5406
10/17/2003
65892
84150
11/11/2004
66453
15402
10/26/2003
62934
81498
11/19/2004
70635
7293
10/31/2003
61710
80631
11/24/2004
50337
16371
11/12/2003
57936
75939
12/1/2004
42228
29631
11/19/2003
58956
74460
12/8/2004
34884
24939
11/26/2003
58497
74970
12/12/2004
30498
23766
12/3/2003
56559
73899
1/6/2005
38900
38850
12/10/2003
56253
73287
1/13/2005
18600
32700
12/17/2003
55233
70890
1/21/2005
13650
12/23/2003
58191
69870
2/5/2005
4100
1450
1/2/2004
49929
68085
2/8/2005
4000
3700
1/14/2004
56661
60996
2/24/2005
3400
6850
1/22/2004
51561
64821
3/12/2005
3650
4300
1/29/2004
54621
63648
3/20/2005
4600
4500
2/22/2004
58089
62577
3/26/2005
4600
4500
2/26/2004
52785
61302
4/3/2005
5050
5250
7/28/2004
61450
67850
4/17/2005
6700
3900
8/12/2004
50400
56600
4/26/2005
5890
4440
8/17/2004
52650
55850
4/30/2005
6700
4100
8/24/2004
54100
57650
5/7/2005
8550
5000
9/3/2004
53150
57850
5/16/2005
6800
4350
9/10/2004
53150
58150
5/23/2005
7350
4350
9/15/2004
53350
58050
5/31/2005
7290
5320
9/23/2004
51250
56450
6/6/2005
6730
5250
10/5/2004
50694
58701
6/14/2005
7230
5160
10/26/2004
52122
53652
6/28/2005
5880
4610
11/3/2004
41922
56610
11/11/2004
50337
52734
11/19/2004
52632
52326
11/24/2004
52632
52836
12/1/2004
52479
54519
12/8/2004
51408
51000
12/12/2004
48246
49572
1/6/2005
52000
39900
1/13/2005
52350
38850


CHAPTER 4
THE EVALUATION OF LIGNOCELLULOSIC WASTE DECOMPOSITION OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILLS
4.1 Introduction
According to an EPA report, 153 million tons of lingocellulosic materials were
generated as part of the MSW stream in 2003 (USEPA, 2005). In addition to MSW
components such as paper and wood products, lignocellulosic wastes also include forest
industry, pulp and paper industry, agricultural and food processing wastes (Source:
http://www.utoronto.ca/forest/termite/lig-mat.htm, Last accessed November 17th, 2005).
Since lignocellulosic materials occupy approximately 65% of the U.S. MSW stream by
weight (USEPA, 2005), their biological decomposition in landfills results in a large
volume of gas production and leads to the settlement of the landfill surface. An
understanding of this decomposition process is thus helpful to a landfill lysimeter.
The term lignocellulosic implies, materials consisting primarily of lignin and
cellulose. In addition to macro-cellulose (glucan), many different sugar residues
(hemicellulose) may be present. Cellulose consists of numerous glucoses linked by P-1-4
linkages, while lignin is composed of benzene-ring-containing monomers with net
structures. Lignin is present extensively with cellulose within and between distinctive
morphological regions protecting the cellulose and other cell structures (Pagarwal, 2005;
Brown, 1985). Chandler et al. (1980) showed the biodegradability of lignocellulosic
materials to be strongly associated with lignin content. Stinson (1995) showed that
methanogenic activity was highly affected by the degree of delignification of wood
95


168
Figure B-12. Separated newspaper and office paper


196
Table C-4 Total dissolved solids (TPS) of aerobic and anaerobic lysimeters (unit: mg/L)
date
lys 1
lys 2
date
lys 3
lys 4
7/28/2004
4.00
4.00
8/13/2003
37.62
48.26
8/4/2004
16.44
12.72
8/15/2003
44.41
37.62
8/17/2004
15.32
6.04
8/22/2003
48.16
53.78
9/9/2004
43.84
2.98
8/26/2003
48.31
50.70
10/8/2004
39.92
0.86
9/9/2003
49.63
49.06
10/19/2004
46.02
0.70
9/16/2003
50.12
52.31
10/26/2004
41.12
11.02
9/24/2003
48.42
53.94
11/6/2004
49.26
7.88
10/21/2003
43.04
54.84
11/13/2004
39.42
6.28
10/28/2003
27.54
43.64
1/25/2005
15.54
6.20
7/28/2004
32.00
30.00
1/30/2005
2.26
8/4/2004
28.72
30.66
5-Feb
6.12
1.48
8/17/2004
30.64
29.06
16-Feb
7.16
4.64
9/9/2004
32.60
37.72
24-Feb
5.50
8.48
10/8/2004
27.46
35.42
1-Mar
4.38
6.84
10/19/2004
33.92
35.00
12-Mar
5.74
6.18
10/26/2004
41.08
36.26
20-Mar
6.42
6.10
11/6/2004
32.22
33.70
26-Mar
5.68
6.32
11/13/2004
38.10
36.28
3-Apr
6.46
6.14
1/25/2005
34.08
27.66
10-Apr
6.86
6.52
1/30/2005
25.74
17-Apr
8.32
5.90
5-Feb
33.04
25.48
24-Apr
8.14
4.88
16-Feb
33.34
27.14
30-Apr
8.1
6.26
24-Feb
35.30
28.54
7-May
8.88
6.70
1-Mar
31.46
23.34
16-May
8.14
6.48
12-Mar
34.04
29.14
23-May
8.18
6.42
20-Mar
35.82
26.30
31-May
7.9
6.28
26-Mar
33.78
29.32
6-Jun
8.82
7.08
3-Apr
35.20
25.88
14-Jun
7.76
6.82
10-Apr
32.42
25.8
28-Jun
7.34
6.32
17-Apr
33.98
27.98
5-Jul
6.82
7.28
24-Apr
30.86
24.72
12-Jul
5.275
5.8
30-Apr
30.84
25.7
10-Aug
6.5
7.68
7-May
33.44
26.12
16-May
32.92
24.48
23-May
30.66
23.72
31-May
30.42
22.52
6-Jun
31.62
23.48
14-Jun
29.84
22.42
28-Jun
28.22
21.94
5-Jul
29.54
20.96
12-Jul
28.60
19.78
10-Aug
27.84
17.28


10
lysimeter. Two liters of DI water were added along with the compaction of each waste
fraction. After loading, 11 L of additional water was added from the top of each lysimeter.
The goal of adding water was to bring the waste in each lysimeter, at the beginning of the
experiment, to field capacity. A capacity of 58% was targeted as this was the field
capacity measured for this waste under the initial compaction conditions of the lysimeter.
The waste was compacted to a density of 30 lb/ft dry (480.6 kg/m dry)-
2.2.4 Air Injection
Two computer-controlled pump drives (Model No. 7550-10, Cole-Parmer) were
used for air addition. Air was saturated and warmed prior to injection to keep moisture in
the waste from evaporating. Air was injected on the ports located at the side of the
aerobic lysimeters using a manifold from day 1 to day 164 and changed to the most
bottom port from day 164 to the end of a test period. A flow rate of 70 mL/min was found
to be suitable for control purposes and to maintain low exit gas oxygen concentrations.
The flow rate was adjusted several times during the experiment when oxygen
concentrations in the exit gas became less than 1% to maintain aerobic conditions.
2.2.5 Leachate and Gas Analysis
Leachate samples were collected on a weekly basis. Leachate was analyzed for
sulfide and dissolved oxygen immediately; analysis for pH, alkalinity, and conductivity
was carried out within one hour after collection. After this initial analysis, 15 mL of
leachate was preserved with sulfuric acid and placed in acid-rinsed high-density
polyethylene (HDPE) bottles for later analysis of chemical oxygen demand (COD), total
organic carbon (TOC), volatile fatty acids (VFA) and ammonia. For metal analysis, 50
mL of leachate was preserved with concentrated nitric acid and stored at 4C. The
remaining leachate was recirculated back to the top of the lysimeters. Deionized water


206
Table C-10.
Volatile fatty acids
rVFA) of lysimeter 2
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/18/2004
11166.0
123.6
553.1
1801.7
9/22/2004
2935.9
92.1
506.9
1095.2
9/30/2004
2335.6
84.2
470.3
1024.8
10/21/2004
1432.7
74.0
432.5
1002.9
11/3/2004
1033.7
66.8
415.4
899.2
12/12/2004
3467.0
767.1
1257.8
6229.2
12/20/2004
2726.1
618.7
1120.2
5420.8
1/22/2005
693.0
99.1
1057.2
3258.2
1/30/2005
124.8
19.3
206.7
611.9
2/8/2005
2.7
0.0
3.3
10.8
2/16/2005
7.5
0.0
3.8
7.1
3/3/2005
42.5
0.0
19.2
29.0
3/12/2005
44.0
0.0
19.5
30.3
3/25/2005
36.5
43.2
0.0
27.6
4/3/2005
33.6
42.9
18.8
28.2
4/17/2005
7.9
0.0
2.1
3.7
4/24/2005
8.3
2.1
1.5
0.0
4/30/2005
13.7
2.5
2.8
6.6
5/7/2005
5.3
0.0
1.6
3.7
5/16/2005
8.4
2.7
1.8
4.6
5/23/2005
5.7
1.8
1.6
3.7
5/31/2005
2.7
2.8
1.7
3.7
6/6/2005
3.1
0.0
1.6
4.8
6/14/2005
2.6
2.3
1.5
3.7
6/24/2005
1.9
1.8
0.0
4.0
7/5/2005
0.7
0.0
1.7
4.6
7/11/2005
0.6
3.5
1.8
4.7
7/19/2005
0.9
0.0
2.9
5.8


224
Eary, L. E and Rai, D., 1987, Kinetics of Chromium(IlI) oxidation to chromium(VI) by
reaction with manganese-dioxide, Environmental Science & Technology, 21(12),
1187-1193.
Lee, G., Bigham, J. M. and Faure, G., 2002, Removal of trace metals by coprecipitation
with Fe, A1 and Mn from natural waters contaminated with acid mine drainage in
the Ducktown Mining District, Tennessee, Applied Geochemistry, 17(5), 569-581.
Lee, FL, 1996, Waste composition and characteristics as predictors of landfill
stabilization, doctoral dissertation, University of Florida
Lee, J. Y., Lee, C. H. and Lee, K. K., 2002, Evaluation of air injection and extraction
tests in a landfill site in Korea: implications for landfill management,
Environmental Geology, 42(8), 945-954.
Liao, S. Y., Cheng, Q., Jiang, D. M., and Gao, J., 2005, Experimental study of
flammability limits of natural gas-air mixture, Journal of Hazardous Materials,
119(1-3), 81-84.
Ling, H. L, Leshchinsky, D., Mohri, Y., and Kawabata, T., 1998, Estimation of municipal
solid waste landfill settlement, Journal of Geotechnical and Geoenvironmental
Engineering, 124(1), 21-28.
Marques, A. C. M., Filz, G. M. and Vilar, O. M., 2003, Composite compressibility model
for municipal solid waste, Journal of Geotechnical and Geoenvironmental
Engineering, 129(4), 372-378.
Masscheleyn, P. H., Delaune, R. D. and Patrick, W. H., 1991, Effect of redox potential
and pH on arsenic speciation and solubility in a contaminated soil, Environmental
Science & Technology, 25(8), 1414-1419.
Mata-Alvarez, J., Cecchi, F., Pavan, P. and Llabres, P., 1990, The performance of
digesters treating the organic fraction of municipal solid wastes differently sorted,
Biological Wastes, 33, 181-199
McBean, E. A., Rovers, F. A. and Farquhar, G. J., 1995, Solid waste landfill engineering
and design, Englewood Cliffs, N.J., Prentice Hall PTR.
McBride, M. B. and Blasiak, J. J., 1979, Zinc and copper solubility as a function of pH in
an acid soil, Soil Science Society of America Journal, 43(5), 866-870.
Meima, J. A. and Comans, R. N. J., 1997, Geochemical modeling of weathering reactions
in municipal solid waste incinerator bottom ash, Environmental Science &
Technology, 31(5), 1269-1276.
Miyazawa, M., Pavan, M. A. and Neto, L. M., 1993, A possible mechanism for
manganese release from acid soil, Pesquisa Agropecuaria Brasileira, 28(6), 725-
731.


107
processes (Tchobanoglous et al., 1993). However, the effect of lignin on biodegradation
of lignocellulosic decomposition is reported to be less in aerobic conditions (Komilis and
Ham, 2003). The biodegradable fraction of lignocellulosic materials in terms of lignin
content can be written as:
Anaerobic: B = 0.83 (0.028) X (r2 = 0.94) (Chandler et al., 1980)
Aerobic: B = 0.85 (0.01) X (r2 = 0.50) (Komilis and Ham, 2003)
where X is the initial lignin contents (as %VS). The effect of lignin on lignocellulosic
decomposition in aerobic condition appeared to be more variable than that in the
anaerobic condition.
The BMP assay and lignin analysis results found no evidence that lignin could be
decomposed in either aerobic and anaerobic landfill conditions. If lignin components
were decomposed or depolymerized, methane yields of SYP blocks could be higher than
those of raw SYP. The most likely explanation may be because of particular ecological
demands of lignin scavengers such as White rot fungi and actinomycetes (Akhtar et al.,
1997); oxygen demands of these microorganisms are reported as 15% in atmosphere
(Reid and Seifert, 1982) while average oxygen content in the aerobic lysimeter was
below 6.5%. Furthermore, overburden pressure applied may limit oxygen transfer.
This research suggests that aerobic landfills may have some advantage for
decomposing high lignin-containing waste in comparison with anaerobic landfills.
However, the percentage of woody waste in MSW stream is relatively small because
Class III MSW landfills do not accept construction and demolition (C&D) debris.
Moreover, according to the United States Environmental Protection Agency (USEPA)
report, the recycling ratio of newspaper substantially increased in 2003 (USEPA, 2005).


Metal concentrations (mg/L)
184
pH
Figure C-16. Zn concentration versus pH in leachate from the lysimeters


207
Table C-ll.
Volatile fatty acids
rVFA) of lysimeter 3
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/7/2003
1554.1
245.3
0.0
25.2
8/13/2003
5827.3
250.0
1552.1
2356.0
8/15/2003
6130.6
217.2
1633.6
2420.7
8/19/2003
6053.6
215.0
1725.8
2642.0
8/22/2003
4858.8
190.3
1871.4
3027.2
8/26/2003
3132.3
189.0
1807.9
3108.4
9/9/2003
3662.6
185.9
1775.5
2794.8
10/8/2003
7937.9
766.0
3099.2
4090.7
10/15/2003
9727.6
814.7
3484.5
4468.6
10/22/2003
9846.3
770.7
3252.5
4111.8
11/5/2003
10064.4
3984.7
3676.7
4501.6
11/15/2003
11155.3
3995.7
3716.8
4409.8
11/19/2003
10719.7
3849.9
3630.4
4283.9
11/28/2003
11690.5
4238.9
3970.9
4558.8
12/10/2003
10418.4
3681.0
3603.5
4082.4
12/19/2003
8760.7
3471.0
4544.1
6219.2
1/5/2004
7523.3
2792.3
3649.2
4777.0
1/14/2043
8602.9
3084.2
4010.1
5133.5
1/8/2004
10141.2
3784.2
4975.7
6531.1
2/13/2004
9903.0
3553.1
4898.4
6098.9
3/16/2004
10625.4
3616.4
5025.0
6182.5
7/17/2004
13562.0
3304.9
3533.5
3583.1
8/18/2004
27827.8
4940.0
4545.7
6098.8
9/22/2004
17197.7
3640.1
5805.2
6827.0
9/30/2004
12572.9
2739.3
4210.0
5343.0
10/21/2004
11580.7
2639.4
3930.8
5135.3
11/3/2004
12671.8
2809.7
4037.1
5397.7
12/12/2004
8997.0
3211.7
4022.8
5578.8
12/20/2004
14545.6
3897.8
4411.4
6253.5
1/22/2005
15118.1
3327.8
3976.9
5180.8
1/30/2005
16654.3
4341.5
4405.8
5347.6
2/8/2005
18671.3
4486.5
4509.3
5667.4
2/16/2005
16358.8
3847.6
4212.3
5383.6
3/3/2005
14300.5
3576.5
3598.3
6033.2
3/12/2005
14220.9
3209.0
3108.9
5614.0
3/25/2005
14513.6
3534.5
3495.5
5950.6
4/3/2005
14191.5
3391.0
3240.1
5696.3
4/17/2005
13386.3
2435.1
2837.4
5459.3
4/24/2005
12372.5
2138.5
2408.6
4692.6
4/30/2005
11690.6
2100.7
2295.6
4397.6
5/7/2005
13138.9
2411.2
2569.5
4926.4
5/16/2005
13708.8
2693.4
2825.1
5358.1
5/23/2005
31195.0
5051.0
5536.2
9693.8
5/31/2005
12812.7
2619.3
2634.6
4941.7
6/6/2005
12600.7
2630.5
2579.5
4771.7
6/14/2005
14119.6
2921.4
2853.6
5442.2


19
presented in the aerobic lysimeters producing sulfide. Relatively high ammonia
concentrations found from the aerobic lysimeters (Figure 2-9) also indicated the presence
of anaerobic zones in the aerobic lysimeters.
2.3.7 Gas Quality
Biogas emitted from the aerobic and anaerobic lysimeters was measured for O2,
CO2 and CH4. Figures 2-13 and 2-14 depict the changes in gas concentrations of aerobic
and anaerobic lysimeters over a period of time. The initial air injection rate of the aerobic
lysimeters was 70 mL/min. The air injection rate was regulated by changes of oxygen
levels within the range of 70 to 120 mL/min. High CO2 concentrations were observed
from aerobic lysimeters during the first 50 days, but decreased to lower than 20%. The
concentrations of CH4 and CO2 of the anaerobic lysimeters changed during the acidic
phase, but stabilized to approximately 60% CH4 and 40% CO2 during the methane phase.
Overall, a total of 40,100 liters (1,400 ft3) of gas was injected into each aerobic
lysimeters for a test period, and 45% and 43% of the oxygen included in the air added
was converted into CO2 in lysimeter 1 and 2, respectively. In the anaerobic lysimeters,
500 and 1,600 liters of biogas (CO2 and CH4) were produced from lysimeters 3 and 4.
Most of the gas generated was mainly concentrated on the methanogenic phase in
anaerobic lysimeters while a relatively steady gas generation was exhibited over time in
aerobic lysimeters as summarized in Figure 2-15.
Lab air was added to recover the lysimeter 3, which had remained in acidic
condition (pH <6) for 600 days. Figure 2-16 depicts the change in gas concentrations, gas
generation rate and pH during air injection. The pH was adjusted to 7.1 at day 4, and air
injection was stopped on day 5. After oxygen was depleted in the lysimeter, methane
concentrations substantially increased along with biogas generation rate and reached


5-1. The changes in settlement, cumulative gas (CO2) and pH over time 134
5-2. The changes in settlement, cumulative gas (CO2 and CH4) and pH over time 135
5-3. Settlement behaviors and compression coefficients of aerobic and anaerobic
lysimeter over a period of time 136
5-4. Relationship between settlement and overall mass loss of the aerobic and
anaerobic lysimeters 137
5-5. Relationship between percentage of settlement and mass loss 138
5-6. Correlation of logarithm of mass loss of the aerobic lysimeters over time 139
5-7. Different k values of anaerobic lysimeters at lag and log phases 139
5-8. Settlement prediction of the aerobic lysimeters 140
A-l. Schematic of mass loss by waste decomposition 156
A-2. Waste mass loss by TOC and gas generation 158
B-l. Schematics of aerobic and anaerobic lysimeters used for this research 159
B-2. The carriage system 160
B-3. A schematic of the temperature control system 161
B-4. Schematic of gas volume measuring tool; before gas measurement, fill tap-water
up to the top scale 162
B-5. The nation-wide composition of discarded municipal solid waste in 2003 163
B-6. The composition of municipal solid waste in Florida in 2000 163
B-7. (A) Blue water phenomenon observed from gas collection system of aerobic
lysimeters; (B) a hole on copper tube caused by corrosion of Cu 164
B-8. Solid samples excavated from one of the aerobic lysimeter 165
B-9. Decomposed papers were commingled together (aerobic lysimeter) 166
B-10. Not well degraded office paper (aerobic lysimeter) 167
B-l 1 Wood blocks excavated from aerobic lysimeter 167
C-l. The change in COD of the lysimeters over the percentage of mass loss 169
xii


100
80
60
40
20
0 <
100
80
60
40
20
0 i
re C
mim \ hhhi
185
2-2 CB
10 20 30 40
50
Days
. Change in methane yields of the waste layer 2-1 and 2-2


163
yard waste
8%
food waste
17%
paper
26%
office paper
15%
newspaper
1%
cardboard
10%
plastics
15%
glass
metals 6/o
7%
Figure B-5. The nation-wide composition of discarded municipal solid waste in 2003
(EPA, 2005)
miscellaneous
12%
textile
ferrous metal
13%
food waste
9%
plastics
8%
paper
43%
office paper
6%
newspaper
9%
cardboard
14%
other paper
14%
Figure B-6. The composition of municipal solid waste in Florida in 2000 (FDEP, 2002)


2
Air addition has been suggested as another means, in concert with leachate
recirculation, to achieve rapid landfill stabilization. It has been reported that waste
decomposes more rapidly in aerobic systems relative to anaerobic systems (Read et al.,
2001). Additional reports suggest that air injection may stop the production of methane
(one of the most serious greenhouse gases), change the leachate quality for the better,
reduce the amount of volatile organic compounds (VOCs), and improve the degradability
of anaerobically recalcitrant materials (Grima et al., 2000; Read et al., 2001; Lee et al.,
2002; Reinhart et al., 2002). Some of these potential benefits have been investigated at
the lab scale (Stessel and Murphy, 1992), and some positive outcomes have been reported
from field studies (Read et al., 2001; Lee et al, 2002). However, in order to apply this
new technique successfully to full-scale operating landfills, further investigation is
necessary. While anaerobic bioreactors have been heavily simulated in previous studies,
there are few cases involving the simulation of aerobic landfills. It is also rare to find
side-by-side simulations on the same waste stream under the same field conditions
comparing aerobic and anaerobic systems.
1.2 Objectives
The main objective of this research was to compare aerobic and anaerobic landfills
using simulated landfill lysimeters. In the early development of anaerobic bioreactors,
several fundamental simulated landfill experiments were performed that have provided
much of our understanding of such processes to date (Pohland, 1980). This research
presents the results of parallel aerobic and anaerobic simulated bioreactors. Several
different parameters of concern were investigated: leachate and gas quality, settlement,
heavy metal fate, and decomposition of lignocellulosic materials. The following were
specific objectives of this research:


121
of each lysimeter and a geotextile was placed between rock and waste. Each waste
fraction was then loaded and compacted until it occupied 25 % of the depth of the
lysimeter. Two liters of DI water was added along with the compaction of each waste
fraction. After loading, 11L of additional water was added from the top of each lysimeter.
The waste was compacted to a density of 30 lb/ft dry (480.6 kg/m dry)-
5.2.2 Application of Overburden Pressure
Overburden pressure was applied to the fabricated waste in order to characterize the
decomposition of waste disposed in a deep landfill. An Enerpac hydraulic cylinder and
hand pump was utilized to generate overburden pressure. The hydraulic cylinder was
mounted on the top of the carriage system. Overburden pressure was transferred through
a shaft attached to the cylinder to the waste in the lysimeter. As settlement proceeded, the
length of the shaft was extended by exchanging and connecting other pieces of shaft of
different lengths.
A load of 2040 lb/ft2 (98 kPa) was used as the overburden pressure to simulate a lift
of waste overlain by 40 ft (12 m) of waste. This was for the simulation of the pressure
applied under 4 ft (1.2 m) cover soil and 36ft (11 m)-depth of waste. It was assumed that
the cover soil occupied 10 % of the volume of the landfill, the density of the cover soil
was 110 lb/ft3 (1760 kg/m3), and the compacted waste density was 44.4 lb/ft3 (710
kg/m3). In the field, waste can be compacted by modem technology to 60 lb/ft3 (960
kg/m3). The density of waste used for this research falls into the good compacted range
for MSW (Oweis and Khera, 1998)
The measurement of the initial depth of each lysimeter was conducted after the first
application of overburden pressure at day 1. Thus the percentage of settlement addressed
here is the ratio between the depth of waste deformed and the initial depth of waste. The


193
Table C-2 (continued)
date
lys 1
lys 2
date
lys 3
lys 4
12/13/2004
> 20000
15500
12/20/2004
18200
1/5/2005
15700
13300
1/17/2005
14500
12600
1/25/2005
19200
18000
1/30/2005
17200
16100
2/5/2005
18400
14500
2/8/2005
14900
2/16/2005
18900
19000
2/24/2005
> 20000
18000
3/1/2005
19200
> 20000
3/12/2005
18800
> 20000
3/20/2005
19900
19900
3/26/2005
19400
> 20000
4/3/2005
> 20000
> 20000
4/10/2005
> 20000
> 20000
4/17/2005
> 20000
> 20000
4/24/2005
19700
> 20000
4/30/2005
> 20000
> 20000
5/7/2005
17200
> 20000
5/16/2005
23200
23200
5/23/2005
22400
22600
5/31/2005
21800
21700
6/6/2005
22000
21000
6/14/2005
24600
23800
6/28/2005
25900
22900
7/12/2005
25600
23600
7/19/2005
24800
22700


80
60
40
20
0
100
80
60
40
20
0
eC
IHHHf \ HHHI
186
2-3 CB
2-3 NP
2-3 OP
2-3 wood
cellulose
Blank
Newspaper
Jk 1
'i
10
- -rTU
0 ; $=&
20 30 40 50
2-4 CB
2-4 NP
2-4 OP
2-4 wood
cellulose
Blank
Newspaper
Col 1 vs Blank
^ 1
=?=g~T ^
10 20 30 40 50
Days
. Change in methane yields of the waste layer 2-3 and 2-4


127
It is noted that this relationship is applicable to both aerobic and anaerobic
lysimeters. It also shows that the degree of waste decomposition is associated with
volume loss irrespective of the decomposition rate.
5.3.3 Ultimate Settlement
Prior to applying the settlement model as a function of mass loss, it is important to
consider the limitation of the waste decomposition. As previously discussed in chapter 4,
waste can be degraded up to the point BVS. Biodegradable volatile solid of the raw waste
was determined as 67% (see chapter 4). Therefore, any predicted values beyond that
point are necessarily deemed to be overestimation and need to be replaced with the
maximum mass loss. The time to reach the BVS may be variable depending on the
landfill conditions such as percentage of moisture, temperature and waste compositions.
In addition to the natural conditions, the time to reach BVS can be advanced by moisture
addition and air injection.
5.4 Discussion
5.4.1 Compression Index
Table 5-3 represents the modified compression indices of other studies in
comparison with results of this lysimeter study. In is noted that (Ca)'min values of
different case studies are similar while (Ca)'max values appear to be quite different. In
many cases compression index value of settlement occurred by mechanical interaction
would be less than 0.1. Sowers (1973) reported that the compression indices might
increase as organic content increased. Sowers also suggested that the compression index
could be a function of void ratio. In order words, compression indices may increase as
void ratio increases. Since the void volume is defined as the ratio of the volume of voids
to the volume of solids (Das, 2002), various modem landfill technologies such as


104
4.3.3 Biodegradability of Excavated Wastes
The biodegradability of each component was determined using the methane yield
data from the BMP assays. Figure 4-4 presents the changes in cumulative CH4 volume of
each component of waste layer 2-3 over time. The cumulative CH4 volume of all of the
components fell between cellulose (positive control) and the blank (sludge only; negative
control). Among the four biodegradable wastes, the greatest volume of methane was
produced from office paper fraction and the least methane volume was observed from the
wood fraction. Biodegradability of the lignocellulosic waste can be arranged as follows
(from largest to lowest): office paper > cardboard > newspaper > wood. Figure 4-5 shows
the differences of the methane yields of biodegradable wastes excavated from lysimeters
2 and 4 relative to those of the raw waste. Overall the methane yields of the solid wastes
excavated from the anaerobic lysimeter were statistically greater in comparison with
those of the aerobic lysimeter (p < 0.05). The higher the methane yields the waste
samples showed, the less decomposed they were. It is noted that higher methane yields of
newspaper were observed from all paper fractions of anaerobic lysimeter (Figures C-18
through C-20).
Table 4-4 presents the overall methane yields of each waste layer. These values
were calculated based on the methane yield of each component and the waste fraction
obtained from the garbage separation process. The greatest degree of waste
decomposition was observed from the middle layers (layer 2-2 and 2-3), and a relatively
low decomposition was observed with the top and bottom layers. For lysimeter 4,
however, though the overall methane yield was higher than that of the lysimeter 2, similar
methane yields were observed from most layers except for the layer 4-4. Among all waste
layers excavated from lysimeter 2 and 4, layer 4-4 showed the least waste decomposition.


CHAPTER 6
SUMMARY AND CONCLUSIONS
6.1 Summary
In this research, the overall performance of aerobic and anaerobic landfills was
compared. Four stainless-steel lysimeters were used as simulated landfills. Based on
waste stream data published by the U.S. EPA and FDEP, fabricated garbage was made
and loaded in each lysimeter. Constant pressure (2040 psf) was applied to the fabricated
waste to simulate MSW placed at a depth of 40 feet in the landfill. Leachate produced
was collected from each lysimeter on a weekly basis. Leachate collected was injected
back to the lysimeters, with deionized water used as makeup water, after analyzing for
metals and conventional water quality parameters. In addition, gas quality and settlement
were monitored during the entire operation. The period of time spent researching the
aerobic and anaerobic lysimeters differed (365 and 719 days, respectively). After the
lysimeter studies were completed, one each of the aerobic and anaerobic lysimeters was
dissembled and the decomposed wastes were excavated. Lignocellulosic materials
contained in the wastes were separated and analyzed for metal content and the degree of
biodegradation.
More than 90% of the leachate COD, BOD and TOC was reduced within 100 days
in the aerobic lysimeters. The concentrations of ammonia remained the same during the
acidic phase in the anaerobic condition; however, ammonia concentrations increased to
an amount four times higher in the methane phase than the initial concentration. No such
dramatic change in ammonia was observed from the aerobic lysimeter. It was noted that
141


BOD (mg/L) BOD (mg/L)
34
Days
Figure 2-5. Changes in BOD of aerobic and anaerobic lysimeters versus time


COD (mg/L) COD (mg/L)
33
Days
Figure 2-4. Changes in COD of aerobic and anaerobic lysimeters versus time


133
Table 5-3. Comparison of compress indices between current study and other studies
References
mjn
V'-'ctj max
24 landfill case (Bjargard and Edgers, 1990)
0.019
0.125
Laboratory large scale (Gandolla et al., 1992)
0.063
0.34
Laboratory large scale (Lee et al., 1995)
0.063
0.149
Lysimeter 1 (this study)
0.063
0.175
Lysimeter 2 (this study)
0.072
0.227
Lysimeter 3 (this study)
0.050
0.140
Lysimeter 4 (this study)
0.032
0.536
15-year-old landfills, Boston, MA
0.24
Old landfill, WV
0.30
10-year-old landfills, Elizabeth, NJ
0.02


30
Back Front


63
these metals might be influenced by the different environment of the aerobic lysimeters
such as an alkaline pH and oxidizing conditions.
3.4.2 Comparison to Other Studies
Generally, heavy metal concentrations found in anaerobic landfills are reported low
(Kjeldson et al., 2002). The presence of sulfide and the low solubility of metals at neutral
pH may reduce metal concentrations in leachate. Metal concentrations of the aerobic and
anaerobic lysimeters along with MSW leachate summarized from the literature are
presented in Table 3-6. Metal concentrations of the aerobic lysimeters listed in Table 3-6
are the average value of the lysimeters 1 and 2. For the anaerobic lysimeters, since
lysimeter 3 remained in acidic condition for most of a test period, metal results of
lysimeter 3 during the methanogenic phase were not included in Table 3-6. For the
anaerobic lysimeters, As and Zn concentrations were substantially higher than those of
MSW leachate during acidic phase. They were reduced then during the methanogenic
phase and similar to those of MSW leachate despite the presence of CCA-treated wood.
Cu concentrations of the anaerobic lysimeters were extremely low, and they were lower
than even drinking water standards. Although most metal concentrations of the anaerobic
lysimeters were greater than drinking water standards, they were similar or lower than
those of general MSW leachate during the methanogenic phase.
For the aerobic lysimeters, As, Cr, Fe, Mn and Zn concentrations were lower than
those of MSW leachate and the anaerobic lysimeters during the first acidic phase.
Aluminum and copper concentrations were greater than those of the anaerobic lysimeters.
Only Pb concentration was greater than that of MSW landfill leachate and the anaerobic
lysimeters. However, different aspects of metal leaching were observed from the aerobic


131
5.5 Conclusions
The landfill settlement behavior occurring in aerobic and anaerobic simulated
landfills was mathematically analyzed. The logarithm of mass loss can be linearly
correlated with volume loss as follows:
[Settlement, %] = (AH, %) = 16.90 x log [mass loss, %] 6.24
Assuming the same cross-sectional areas, volume loss and the percentage of settlement
can be identical. With these relationships, the secondary settlement of aerobic and
anaerobic simulated landfills could be mathematically modeled. The first-order
exponential functions could be used to describe waste decomposition.
Settlement data obtained from this research could provide a useful tool in determining the
decay value, the most difficult derivative in the bioconsolidation model. Assuming that
waste decomposition follows a first-order kinetic model, the decay coefficients
determined based on the settlement data were 0.379 and 0.377 yr'1 for the aerobic
lysimeters. Decay coefficients for the anaerobic lysimeters changed by waste
decomposition phase. Decay coefficients calculated were 0.015 and 0.022 yr'1 and 0.194
and 0.246 yr"1 for the acid and methane phases of the lysimeters 3 and 4, respectively.


4
lysimeters to compensate for the amount of leachate lost by leachate collection. The gas
volume and composition were monitored using a gas totalizer and gas chromatography.
To explore the fate of heavy metals leached out of the fabricated wastes under
aerobic and anaerobic conditions, heavy metal-containing wastes (e.g., CCA-treated
wood, cathode-ray tube (CRT) glass and pieces of sheet metal) were mixed with the other
fabricated wastes before loading into the lysimeters. After loading and compacting the
fabricated waste, leachate generated by the lysimeter was collected and analyzed for
copper, chromium, arsenic, lead, aluminum, zinc, manganese and iron. The change of
heavy metal concentrations in the leachate over time was monitored. After the lysimeter
work was completed, the wastes excavated from two columns were analyzed for heavy
metals in order to compare heavy metal concentrations absorbed on solid waste to those
released from the lysimeters through the leachate.
To explore the decomposition of lignocellulosic wastes in aerobic and anaerobic
landfill environments, lignocellulosic wastes including paper and wood blocks were
prepared. They were included in the fabricated waste and loaded in the lysimeters. After
the lysimeter study was completed, lignocellulosic wastes were excavated and separated.
Biochemical methane potential (BMP) assays were used to evaluate the degree of
biodegradation of each lignocellulosic waste. In order to evaluate the impact of air
addition on wood waste decomposition, cellulose, lignin and BMP of raw and excavated
wood blocks were compared with respect to cellulose and lignin concentrations and BMP
values.
To simulate landfill settlement in aerobic and anaerobic conditions as a function of
waste mass loss, overburden pressure was applied to the stainless steel lysimeters using a


162
Figure B-4. Schematic of gas volume measuring tool; before gas measurement, fill tap-
water up to the top scale (VI close and V2 and V3 open) and close V2.
After connecting an air-sampling bag to the top, open V1 and a valve on the
air-sampling bag to let water drain out. As water drains out by gravity force,
gas in air-sampling bag transfers into the pipe and replaces with water. When
no more gas is left in air-sampling bag, water draining stops by itself. Close
V1 and measure the water volume replaced with gas.


166


167
Figure B-10. Not well degraded office paper (aerobic lysimeter)
Figure B-l 1 Wood blocks excavated from aerobic lysimeter


201
Table C-7. NH3+-N concentrations of aerobic and anaerobic lysimeters
Date
Lys 1
Lys 2
Date
Lys 3
Lys 4
1
44.5737
26.0448
13
123.1292
101.564
13
52.794
19
129.5536
129.0838
16
43.9773
51.5659
29
129.0838
144.4716
21
98.4427
41.2616
40
158.783
174.5122
28
125.3132
43.1491
47
176.4246
179.6586
38
153.8117
83.6837
55
198.1744
232.5246
45
125.3132
62.3102
69
216.2289
250.9585
50
184.8154
112.8914
76
199.6195
262.141
58
166.1884
70.1399
82
196.0262
278.8412
70
146.9202
95.2447
89
201.0752
282.9228
74
126.157
144.5331
97
184.286
283.9525
91
114.8975
80.0651
104
173.8794
363.5242
107
144.0877
106.5606
118
191.2202
383.5107
115
187.1865
206.4054
125
201.2484
310.1469
120
189.5881
142.6142
132
151.6302
342.1821
138
125.5446
141.4073
139
190.4698
335.5207
159
179.0542
148.5193
145
198.8885
332.8926
160
338.2198
196.4142
155
196.5562
326.412
163
268.6915
395.4994
162
215.9736
221.551
178
334.7243
210.4829
225
178.0782
178.0782
187
102.4254
183.2859
363
156.0991
196.3619
193
166.1915
100.2522
375
171.3963
207.5146
204
256.9294
144.1649
378
166.373
207.5146
212
255.886
217.4478
383
204.8859
228
448.6748
230.1956
390
202.743
244.4261
236
152.6165
162.2223
400
354.8838
347.6131
242
168.9598
179.5944
407
228.8206
255.2256
250
194.031
181.062
412
451.3427
313.1798
257
303.5795
214.8094
420
328.1665
339.5144
264
348.6262
219.2248
432
364.9485
325.3893
271
289.1112
220.1187
436
211.072
212.8735
277
402.0369
266.8583
453
299.062
329.7673
286
404.2729
270.5288
469
543.3762
507.6597
293
342.8068
290.9748
477
236.5871
267.615
300
215.5594
259.967
482
605.1256
600.0046
308
356.8128
231.5037
500
664.4257
770.9723
314
350.5038
198.0456
521
209.6439
240.1807
322
319.1664
242.0624
522
608.4553
426.5548
336
301.1875
193.6781
525
923.2783
505.5622
343
506.625
324.996
540
919.5288
1068.944
349
496.8555
295.9966
549
912.0755
1348.001
357
502.6943
269.5848
555
868.6071
365
512.5787
265.4179
566
915.7946
1252.792
379
510.5864
253.2997
574
861.5665
1252.792
590
997.4957
1567.038
598
854.583
1432.845
604
1017.999
1612.317


27
Table 2-3. Comparison of initial and final characteristics of the aerobic lysimeters
Initial
Final (1 year)
Lys 1
Lys 2
Lys 1
Lys 2
Water quantity (mL)
19,000
19,000
15,137*
15,582(15,056*)
Dry waste quantity (g)
12,784
12,784
8,389*
8,740 (8,715*)
pH
5.7
5.7
8.5
8.5
COD (mg/L)
20,000
28,000
3,400
4,700
BOD
13,000
16,000
200
30
TOC
6,000
7,000
2,600
2,200
Ammonia
70
40
500
250
Fluoride
80
30
0
0
Chloride
200
130
1,200
1,700
Sodium
80
140
800
900
* predicted


Gas concentrations (%) Gas concentrations (%)
178
Figure C-10. The change in biogas produced from the anaerobic lysimeters


25
Table 2-1. MSW components.
Waste components
Sources
Processing for size reduction
Office paper
Mixed scrap paper purchased at
office supply store
Grind with a paper shredder
Cardboard
Mixed corrugated boxes
Scissors and razor blade
Newspaper
Local newspaper
Grind with a paper shredder
Plastics
PET bottles collected from a
plastics recycler
Scissors
Food waste
Commercial dog food
Grind with a coffee grinder (less
than 1/32)
Southern yellow pine (SYP)
Home improvement store
Cut with band-saw (2 x 2)
CCA-treated wood
Home improvement store
Gather saw dust after drilling
Galvanized steel
Home improvement store
Cut with metal cutter (1/2 x
1/2)
Aluminum
Home improvement store
Cut with metal cutter (1/2 x
1/2)
Cathode-Ray Tube(CRT) glass
CRT monitors
Crush with a hammer (1/4-1/8)
Mixed cullet
Mixed container glass
Crush with a hammer (1/4 1/8)


Metal concentrations (mg/L)
181
pH
Figure C-13. Cu concentration versus pH in leachate from the lysimeters


5
hydraulic cylinder and hand pump. To correlate mass loss and volume loss, a lab-scale
experiment was designed where waste was decomposed in simulated landfills in the
laboratory with both mass loss and volume loss being measured. A difficulty with using
lab experiments to simulate landfill settlement is that it is hard to simulate true landfill
conditions, especially, the large overburden pressure. In this research, the experiments
included the application of overburden pressure to make the laboratory condition closer
to the field conditions.
1.4 Outline of Dissertation
The dissertation is presented in six chapters. The current chapter presents the
problem statement, objectives and research approach. Chapter 2 presents the comparison
of gas and leachate qualities between aerobic and aerobic simulated landfills. Chapter 3
presents the fate of heavy metals in aerobic and anaerobic simulated landfills. The
evaluation of biodegradation of lignocellulosic materials is presented in chapter 4.
Settlement behavior with waste decomposition is presented in chapter 5. Chapter 6
presents a summary, conclusions and recommendations for future work. Background and
other analytical procedures used for this research are presented in appendix A.
Supplemental s are presented in appendix B. All other tables and s pertaining to leachate
data and BMP are presented in appendix C.


66
concentration of Fe (II) may increase the Cr solubility, but the change is very small. Cu
concentrations in anaerobic landfills are typically extremely low. However, high
concentrations of Cu may be found in both acid and alkaline phase of aerobic landfills.
Since the microorganisms which contribute Cu leaching may corrode all Cu-made
equipment connected to the landfill, it is important to avoid using any Cu-containing
equipment for gas and leachate collection systems. Pb has high solubility under oxidizing
conditions at pH < 6. Thus it would be recommendable to monitor leachate quality for the
first acid phase of aerobic landfills. However, air addition facilities are generally installed
after landfill closure, Pb concentrations may not be high over operation period. Pb
concentration in anaerobic landfill conditions is generally low.
As previously discussed, air addition into a current anaerobic bioreactor landfill
may enhance waste decomposition substantially. However, unlike anaerobic landfills,
concerns about high A1 and Cu concentration and a risk of Cr (VI) may arise. In order to
avoid these potential risks, it would be recommendable to inject air into the shallow well
rather than the deep well. Since Al, Cu and Cr are redox-sensitive, great amount of these
metals could be reduced after passing through the anaerobic zone. However, since air
injection into the shallow wells may reduce the air diffusion efficiency into a landfill,
further economical and efficiency of air distribution analysis are needed.
3.4.4 The Impact of Air on Metal Mobility
Although total amounts of metals leached were not considerably high, it is
necessary to pay great attention to certain metals due to changes of their toxicity by
different pH and redox conditions. For example, among Cr species dissolved in leachate,
Cr (III) can be dominant in current anaerobic sanitary landfills, however,
thermodynamically Cr (VI) becomes a major Cr component in the environment formed


146
Air addition into the lysimeters substantially enhanced waste decomposition.
Concentrations of oxygen demanding substrates and VFAs were reduced. More
than 90% of the dissolved organic carbon was decomposed within 100 days in the
aerobic lysimeters.
Air addition could be used to recover acid-stuck landfills.
Ammonia concentrations in anaerobic lysimeters during the methanogenic phase
increased by an amount four times greater than those in the acidic phase.
High VFA concentrations did not affect mass loss of the aerobic lysimeters.
Mass loss could be estimated by gas and leachate quality. Estimated mass loss
corresponded with actual mass loss; mass losses predicted for lysimeters 2 and 4
were 4,069 g and 3,787 g, while actual mass losses were 4,044 g and 3,525 g,
respectively.
Among 8 metals (Al, As, Cu, Cr, Fe, Pb, Mn and Zn), average concentrations of
As, Fe, Mn, and Zn in the anaerobic lysimeters were significantly greater than
those of the aerobic lysimeters. In contrast, greater concentrations of Al, Cu, Cr
and Pb were found in leachate from the aerobic lysimeters.
Toxicity of some metals potentially changed with their oxidation states at given
conditions such as redox potential and pH. Cr changed its oxidation states to Cr
(VI) at alkaline pH under oxidizing conditions.
In the presence of CCA-treated wood, As concentrations proved substantially
higher than Cr and Cu concentrations in anaerobic condition; Cr and Cu
concentrations were relatively very low in the anaerobic lysimeters.


31
CCA treated wood'
1%
Alumium
4%
Galvanized steel
4%
Plastic
15%
SYP
5%
CRT glass Mixed cullet
1% 6%
Paper office paper
27%
aper newsprint
6%
Food waste
15%
Paper cardboard
16%
Figure 2-2. The composition of fabricated municipal solid waste for this research.


152
A.2 Estimation of Biodegradable Volatile Solids (BVS)
Biodegradable volatile solids (BVS) can be estimated using methane potential as
follows:
BVS(%) =
a mass of waste (g) converted into CH 4
Initial dry mass (g)
x 100
(9)
In order to calculate a mass converted into CH4, total CH4 volume generated was
corrected to CH4 volume in standard temperature and pressure (STP, 0C and 1 atm)
using the ideal gas law:
P V P V
rl V1 r2 v2
(10)
Arranging equation (10) for V2, CH4 volume at STP, and proper values for V, P
and T were substituted into equation (2), as shown in the following expression:
[CH4 generated at STP (L)] = [CH. generated(L)]x ^ x (76 ~ *2)l"mHg
(273 + 35)K 760mmHg
(11)
where, 35 K is a set temperature of BMP; 760 mm Hg is equivalent to 1 atm; and
42 mm Hg is water vapor pressure at 35C. Since gas in the serum bottle was saturated by
water, pressure was corrected by subtracting the partial pressure of water vapor.
Corrected CH4 volume was converted into moles by dividing it by 22.4 L, which is
equivalent to 1 mole of gas at STP. According to the chemical reactions presented in
equation (2), 1 mole of glucose may convert into 3 moles of CH4. However, in reality, a
part of the 1 mole of glucose is used to generate energy, and the rest of it converts into
CH4. Rittman and McCarty (2001) reported that actual CH4 mole converted from 1 mole


64
lysimeters during the alkaline phase; the concentrations of most metals except for Fe
were greater than those of the anaerobic lysimeters.
The leaching behavior of CCA-treated wood of the aerobic and anaerobic
lysimeters was compared to a similar study (Jambeck, 2004). Jambeck (2004) researched
the leaching behavior of CCA-treated wood mixed with MSW through the 6.7-m high
PVC column tests. A total 2% of CCA-treated wood was included in the column and rain
water was used for Jambecks study while 1% of CCA-treated wood and DI water were
used for the present study (Table 3-7). In comparison metal leaching results from
Jambecks study proved similar to the anaerobic lysimeter here (Figure 3-18). Extremely
low Cu concentrations were observed in both studies. The range of Cr concentrations of
Jambecks study was higher than that of the anaerobic lysimeters during acid phase, but
the median of Cr concentrations of Jambecks study was bottom of the range. In contrast
to anaerobic condition, significantly different leaching trends of As and Cu were
exhibited from the aerobic lysimeters; overall As concentrations of the aerobic lysimeter
were lower than those of the anaerobic column studies during the acid phase. The 95th
percentile of As concentrations of the aerobic lysimeter was in the range of the anaerobic
system, but they were detected at the very beginning of the aerobic lysimeter operation.
After pH stabilized, the median As concentrations of both the aerobic and anaerobic
systems became identical. However, Cu concentrations were two orders of magnitude
higher than those of anaerobic lysimeters for both the acid and methane (alkaline) phase.
In comparison with other column study (Jambeck, 2004), it can be concluded that
leached As, Cu and Cr concentrations might not always be followed by the initial mass of
CCA-treated wood and metal concentrations contained. The differences between As and


APPENDIX C
LYSIMETER EXPERIMENT RAW DATA AND GRAPHS
C.l Graphs
Figure C-l. The change in COD of the lysimeters over the percentage of mass loss.
169


13
through C-l 1 in appendix C. These include the raw data and graphs of the leachate
parameters.
2.3.1 pH
Figure 2-3 depicts the change in pH over the course of the experiment. Both the
aerobic and anaerobic lysimeters remained in acidic condition during the beginning of the
experiment. The period of time required to stabilize the pH for the aerobic and anaerobic
lysimeters was 200 and 600 days, respectively. Average pH measurements of
approximately 8.9 (aerobic) and 7.1 (anaerobic) were observed at the end of the
experiment.
Two phases (acidic and alkaline or methane phase) of the pH of the aerobic and
anaerobic lysimeters were observed during a test period. The low pH occurring during the
initial phase of the research was attributed to a build up of organic acid concentrations
and the related microbial activities. Once the organic waste decomposition process began,
the biodegradable fraction of waste was converted into organic acids by various
biological reactions, and the accumulation of the organic acids lowered the pH. For the
aerobic lysimeters, air was injected through four front ports of the lysimeter using
manifolds. The pH was low (< 6) for the first 150 days, and high VFA and alkalinity
concentrations indicated that anaerobic conditions were predominant, suggesting that air
was not evenly distributed through the manifolds. An increase in the pH of the aerobic
lysimeters was observed after air was injected into the only bottom port. Typically, the
pH of the system increases to neutral conditions as the organic acids are consumed by
methanogenic bacteria. A large amount of CO2 production in an unbalanced ecosystem
may also contribute to lowering the pH as well. High concentrations of VFA and
alkalinity were measured in the anaerobic lysimeter leachate during the initial acid phase.


28
Table 2-4. Comparison of initial and final characteristics of the anaerobic lysimeters
Initial
Fina
(2 years)
Lys 3
Lys 4
Lys 3
Lys 4
Water quantity (mL)
19,000
19,000
18,844*
18,833 (18,704*)
Dry waste quantity (g)
12,784
12,784
11,290
9,258 (8,997*)
PH
4.5
4.9
6.5
7.4
COD (mg/L)
65,000
67,000
42,000
24,000
BOD
48,000
62,000
14,000
6,500
TOC
26,000
27,000
12,000
5,600
Ammonia
120
100
1,000
800
Fluoride
1,500
1,400
460
200
Chloride
1,450
1,400
670
500
Sodium
2,000
2,000
4,800
3,800
* predicted


124
5.2.5 Volume Loss versus Mass Loss
In this research, it was hypothesized that the mass loss resulting from waste
decomposition is the primary cause of volume loss. Microscopically, wastes consist of
small particles, and these particles may support each other to maintain their structure
against outer forces. A reduction in the number of particles disrupts this balance and
ultimately leads to loss of volume under the application of pressure. If the cross sectional
area is constant, such as the inside of lysimeters, the volume loss can be directly related
with a decrease of waste depth. Park and Lee (1997) explained the relationship between
landfill settlement and waste decomposition using an artificial model consisting of ice
and inert particles. As the volume of ice particles is reduced by melting, the overall
volume is reduced, a concept that can be applied to landfill settlement.
In order to evaluate the relationship between settlement and mass loss, the
percentage of settlement and overall mass loss obtained from each lysimeter at the end of
a test period were plotted. In order to find a better relation coefficient, linear and semi
logarithm correlations between the two parameters were conducted. This relationship can
be written as:
[settlement, %] = (AH, %) = a x/[mass loss, %] + P (3)
where, a and P were parameters obtained from the relationship. When deriving the
equation (3), several data points which deviated from this relation were excluded. These
data points were obtained during the first 30 days of the lysimeter studies. Thus, these
data points may have been influenced by primary mechanical settlement rather than
biological decomposition.


56
Another possibility of copper leaching from the aerobic lysimeters is the binding of
Cu with ammonium (NHj+). Since both Cu and ammonium are cations, their
complexations are present as an ionic form and can be dissolved in aquatic systems.
Arzutug et al (2004) reported that Cu leaching increased with ammonia concentrations.
However, complexation of Cu and ammonia may occur in a relatively narrow range of
ORP and pH (Hoar and Rothwell, 1970). When plotting Cu and ammonia data obtained
from the aerobic lysimeters, no clear evidence to prove the relationship between Cu
concentrations and ammonia was found (r2 = 0.021). Furthermore, since Cu complexes
with sulfide rather than ammonia in the presence of sulfide (Alymore, 2001), it may be
difficult to leach high concentrations of Cu under landfill conditions.
3.3.1.5 Lead
Figure 3-5 depicts the changes in lead concentrations over time. For aerobic
lysimeters, Pb concentrations dramatically increased to 1.7-2 mg/L within 30 days and
then gradually decreased. After the pH of the aerobic lysimeters stabilized, Pb
concentrations decreased to levels similar to the anaerobic lysimeters. In contrast, little
change in Pb concentrations was observed from the anaerobic lysimeters and low
concentrations of Pb were maintained over the test period. Generally, Pb concentrations
in landfill leachate have been reported to be very low (Charlatchka and Cambier, 2000;
Jang and Townsend, 2003). This is because lead may precipitate with various ligands
such as carbonate ions (CO3 ), sulfide, and volatile fatty acids (VFA).
Lead solubility is generally controlled by carbonate, or other Pb hydroxides and
phosphate in noncalcareous soils (Bradle, 2005). Charlatchka and Cambier (2000)
concluded that Pb solubility increased under oxidizing conditions at a pH of 6.2
However, Pb may precipitate as a form of PbS under reducing conditions in the presence


229
United States Environmental Protection Agency (USEPA), 2004, Monitoring approaches
for landfill bioreactors, EPA-600-R-04-31, Cincinnati, OH, USA.
United States Environmental Protection Agency (USEPA), 2005, Landfill gas emission
model ,LandGEM, version 3.02 users guide, EPA-600/R-05/047, Washington D.C.,
USA.
United States Environmental Protection Agency (USEPA), 2005, Municipal solid waste
in the United States: 2003 facts and s, EPA530-F-05-003, Office of solid waste and
emergency response, Wahington D. C., USA.
Valorga, 1985, Waste recovery as a source of methane and fertilizer The Valorga
process, 2nd Annual Internal Symposium on Industrial Resource Management,
Philadelphia, USA.
Vikman, M., Karjomaa, S., Kapanen, A., Wallenius, K. and Itavaara, M., 2002, The
influence of lignin content and temperature on the biodegradation of lignocellulose
in composting conditions, Applied Microbiology and Biotechnology, 59(4-5), 591-
598.
Vikman, M., Karjomaa, S., Kapanen, A., Wallenius, K. and Itavaara, M., 2002, The
influence of lignin content and temperature on the biodegradation of lignocellulose
in composting conditions, Applied Microbiology and Biotechnology, 59(4-5), 591 -
598.
Wall, D. K. and Zeiss, C., 1995, Municipal Landfill Biodegradation and Settlement,
Journal of Environmental Engineering-ASCE, 121(3), 214-224.
Wang, Q. H., Kuninobu, M., Ogawa, H. L, and Kato, Y., 1999, Degradation of volatile
fatty acids in highly efficient anaerobic digestion, Biomass & Bioenergy, 16(6),
407-416.
Warith M. A. and Takata, G. J., 2004, Effect of aeration on fresh and aged municipal
solid waste in a simulated landfill bioreactor, Water Quality Research Journal of
Canada, 39(3), 223-229.
Watsoncraik, I. A., James, A. G. and Senior, E., 1994, Use of multistage continuous-
culture systems to investigate the effects of temperature on the methanogenic
fermentation of cellulose-degradation intermediates, Water Science and
Technology, 30(12), 153-159.
White, R. H., 1987, Effect of lignin content and extractives on the higher heating value of
wood. Wood and Fiber Science, 19(4), 446-452.
Yazdani, R., Kieffer, J., Akau, H., and Augenstein, D., 2003, Monitoring the performance
of anaerobic landfill cells with fluids recirculation final report, Yolo County,
California, USA.


154
lysimeters due to iron oxidation found in the aerobic lysimeters. Paper samples, which
looked undegraded, were found in the center areas of the aerobic lysimeters; solid wastes
excavated from the anaerobic lysimeters looked uniformly degraded. Wood blocks were
separated immediately the excavation of the solid waste. It proved harder to separate
wood blocks from the solid waste taken from the aerobic lysimeters because the degraded
paper remained attached to the wood (Figure B-9). Almost all of the wood samples were
too dark to recognize identification numbers written with oil-based pen. Since all paper
and wood samples remained damp and attached, parts of the samples were dried at 103C
for two days in order to effectively separate each type.


170
Figure C-2. The change in BOD5 of the aerobic and anaerobic lysimeters over the
percentage of mass loss


213
Table C-15. Metal concentrations of lysimeter 2 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
7/28/2004
1.32
0.08
0.02
0.70
111.65
5.37
0.02
56.49
8/9/2004
4.54
0.03
0.03
0.36
88.62
2.67
0.12
93.43
8/29/2004
10.66
0.02
0.12
7.46
98.22
1.43
1.68
109.50
9/3/2004
10.35
0.01
0.08
2.52
32.56
0.76
0.25
35.76
9/17/2004
0.46
0.12
0.08
0.44
1.50
1.31
0.04
39.90
10/7/2004
8.69
0.01
0.06
3.43
23.45
0.56
0.31
31.35
10/14/2004
10.86
0.01
0.08
1.78
35.85
0.69
0.16
35.11
10/19/2004
8.97
0.01
0.06
2.22
20.94
0.49
0.25
37.84
11/6/2004
2.63
0.02
0.02
1.45
22.36
0.45
0.26
15.66
11/11/2004
6.22
0.04
0.09
5.07
40.55
1.32
0.63
17.47
12/8/2004
6.33
0.09
0.13
4.58
63.06
3.56
0.59
33.93
12/19/2004
2.24
0.03
0.04
1.26
51.02
0.95
0.22
12.90
1/6/2005
4.79
0.12
0.15
2.25
65.76
3.18
0.23
35.85
1/13/2005
1.44
0.11
0.13
1.76
37.96
2.36
0.14
22.78
2/16/2005
7.80
0.41
0.19
1.42
2.39
0.03
0.01
3.02
2/24/2005
11.96
0.54
0.31
1.66
6.63
0.07
0.04
6.82
3/1/2005
12.42
0.80
0.32
4.07
3.93
0.05
0.06
7.45
3/12/2005
9.08
0.51
0.21
2.45
2.45
0.03
0.02
4.71
3/20/2005
6.59
0.50
0.20
1.09
2.49
0.03
0.05
4.98
3/26/2005
9.28
0.70
0.23
1.52
2.70
0.05
0.03
7.36
4/3/2005
11.21
0.84
0.25
1.99
4.15
0.07
0.03
7.40
4/5/2005
10.34
0.13
0.11
0.80
3.78
0.03
0.10
3.60
4/17/2005
5.99
0.51
0.19
1.06
2.09
0.05
0.01
5.19
4/24/2005
16.58
0.82
0.24
0.79
3.38
0.09
0.03
7.50
4/30/2005
8.28
0.60
0.25
1.03
2.70
0.06
0.02
5.93
5/7/2005
6.37
0.59
0.25
1.42
2.70
0.06
0.01
6.25
5/16/2005
7.62
0.71
0.29
1.40
3.26
0.08
0.03
7.55
5/31/2005
9.30
0.79
0.33
1.06
3.66
0.10
0.01
8.50
6/6/2005
6.44
0.66
0.27
1.63
2.87
0.07
0.01
8.56
6/14/2005
6.65
0.66
0.28
1.27
2.99
0.07
0.01
8.64
6/28/2005
9.30
0.63
0.24
0.76
2.94
0.09
0.01
8.21
7/5/2005
4.96
0.59
0.24
1.09
2.93
0.07
0.03
8.80
7/11/2005
7.38
0.69
0.28
1.44
3.31
0.10
0.05
10.72
7/27/2005
5.56
0.61
0.24
0.84
2.72
0.07
0.01
8.46
8/10/2005
7.13
0.85
0.35
1.16
5.20
0.10
0.02
12.40


142
the pH of the aerobic lysimeters increased to a level more alkaline than the pH of the
anaerobic lysimeters. This resulted from the change in the carbonic system caused by
CO2 stripping by air injection. The highest recorded pH of the aerobic lysimeters was
9.17. Air injection was used in an attempt to recover one of the acid-stuck anaerobic
lysimeters. The initial pH was 6.2 and the methane content was 38%. After air injection
started, the pH increased to 7.3 within 8 days, and the methane content increased to 55%
within 11 days after air injection stopped.
Redox and pH changes in the aerobic and anaerobic lysimeters resulted in changes
in metal concentrations dissolved in leachate. Among the 8 metals under consideration,
the average concentrations of As, Fe, Mn, and Zn in the leachate of the anaerobic
lysimeters were significantly greater than those of the aerobic lysimeters. Most of these
metals were leached while the anaerobic lysimeters remained in an acidic condition. In
the presence of sulfide, all of these metals were potentially precipitated with sulfide at a
neutral pH. In contrast, Al, Cu, Cr, and Pb dissolved in leachate of the aerobic lysimeters
and exhibited significantly higher concentrations than observed in the anaerobic
lysimeters. Various redox, pH and biological reactions dictated the leaching patterns of
these metals.
The waste mass losses of aerobic and anaerobic lysimeters were estimated using
leachate and gas quality data. Mass removed from the wastes was primarily converted
into gas. After waste was removed from the lysimeters, the mass of waste excavated was
compared to the estimated mass loss. A good correlation (a 1% difference between the
two values) was found from the excavated aerobic lysimeter. For the anaerobic lysimeter,
the estimated mass loss was greater than the measured mass loss excavated by 7%. These


160
Top fringe
Hydraulic pressurizing unit
Figure B-2. The carriage system


194
Table C-3. Alkalinity
of aerobic and anaerobic lysimeters (unit: mg/L as CaCCb)
date
lys 1
lys 2
date
lys 3
lys 4
7/28/2004
5000
3500
8/26/2003
12,600.00
8/4/2004
4500
2600
9/9/2003
15,000.00
8/12/2004
4250
3000
9/16/2003
10,500.00
8/18/2004
6000
2200
10/8/2003
18,333.33
12,857.14
8/24/2004
9500
80
10/21/2003
12,142.86
12,857.14
9/10/2004
10750
700
10/28/2003
10,000.00
14,285.71
10/8/2004
16050
750
11/5/2003
12,000.00
14,000.00
10/19/2004
16000
11/12/2003
12,000.00
14,000.00
10/26/2004
11250
500
11/19/2003
10,500.00
12,000.00
11/6/2004
14000
1000
11/26/2003
8,500.00
10,500.00
11/13/2004
12300
2000
12/3/2003
11200
15000
11/24/2004
10500
2200
12/10/2003
10200
11000
12/8/2004
7050
5750
12/17/2003
12500
13000
12/20/2004
2969
3750
12/24/2003
10000
10800
1/5/2005
6000
8000
1/2/2004
11000
12000
1/25/2005
3690
2340
1/7/2004
12000
14000
2/5/2005
570
456
1/14/2004
7000
12500
2/16/2005
720
900
1/22/2004
11500
13000
2/24/2005
880
2000
1/29/2004
6000
13500
3/1/2005
700
1750
2/25/2004
12000
3/12/2005
700
1250
3/12/2004
6000
3/20/2005
1000
1400
3/16/2004
8000
11500
3/26/2005
1000
1300
7/28/2004
11000
11500
4/3/2005
1000
1100
8/4/2004
10000
11000
4/10/2005
1000
1100
8/12/2004
10000
11500
4/17/2005
1880
1000
8/18/2004
10000
11000
4/24/2005
1640
1000
8/24/2004
7500
8000
4/30/2005
2120
1000
9/10/2004
8000
9500
5/7/2005
2200
1000
10/8/2004
14550
12300
5/16/2005
2200
1000
10/19/2004
13500
12000
5/23/2005
2200
1000
10/26/2004
12150
11700
5/31/2005
2280
960
11/6/2004
12000
10500
6/6/2005
2280
1000
11/13/2004
12000
12075
6/14/2005
2240
1040
11/24/2004
12750
11250
6/28/2005
2400
1000
12/8/2004
13500
10950
7/12/2005
2200
1000
12/20/2004
13500
12750
7/27/2005
2200
1000
1/5/2005
14000
10710
8/10/2005
2040
1000
1/25/2005
13500
10200
2/5/2005
12880
12880
2/16/2005
14620
14790
2/24/2005
15000
15000
3/1/2005
14700
13650
3/12/2005
14900
15750
3/20/2005
15150
15000
3/26/2005
16300
16200


70
Table 3-4. Leachability of As, Cr, and Cu
Initial cone, (mg/lys)
mg released
% released
aerobic
anaerobic
aerobic
anaerobic
As
1279.1
mg
Lys 1
Lys 3
1.53
8.21
0.12%
0.64%
Lys 2
Lys 4
1.17
9.30
0.09%
0.73%
Cr
1573.1
mg
Lys 1
Lys 3
0.72
0.50
0.05%
0.03%
Lys 2
Lys 4
0.56
1.24
0.04%
0.08%
Cu
723.9
mg
Lys 1
Lys 3
9.82
0.10
1.36%
0.01%
Lys 2
Lys 4
6.37
0.34
0.88%
0.05%
Table 3-5. Comparison of cumulative mass of metal dissolved in leachate and adsorbed
on lignocellulosic materials (unit: mg)
lys 1
lys 2
lys 3
lys 4
Aerobic
(lys 2)
Anaerobic
(lys 4)
LC/AD
of lys 2
LC/AD
of lys 4
A1
27.3
25.9
7.8
8.7
41900
15700
0.06%
0.06%
As
1.5
1.2
8.2
9.3
110
67.6
1.06%
13.76%
Cr
0.7
0.6
0.5
1.2
222.2
243
0.25%
0.51%
Cu
9.8
6.4
0.1
0.3
238.1
151.2
2.68%
0.22%
Fe
161.3
64.2
1200
506.4
31100
19200
0.21%
2.63%
Mn
10.5
2
31
29.3
494.2
177.3
0.40%
16.51%
Pb
0.6
0.5
0.2
0.1
47.8
156.7
1.13%
0.08%
Zn
292.9
64.2
1,100
1,100
9,400
12,500
0.69%
8.60%


208
Table C-ll
continuec
)
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
6/14/2005
14119.6
2921.4
2853.6
5442.2
6/24/2005
11563.2
2668.4
2571.0
4663.6
7/5/2005
7374.9
2194.9
1865.6
3339.8
7/11/2005
8675.3
2771.5
2149.7
3920.2
7/19/2005
7789.0
3316.3
2432.0
3744.9


99
newspaper, cardboard, SYP and dog food. For decomposed waste, however, the BMP
assay of the decomposed dog food was excluded due to difficulties in separation. Instead,
the fine fraction that passed through the screen (0.475 cm) obtained from the garbage
separation process was used for BMP assay. The overall BMP was determined by the
method used for the evaluation of the overall BMP of waste excavated from a landfill
(Lee, 1996; Townsend et al., 1996). The overall BMP values of the waste samples were
determined as the sum of the multiplication of the VS fraction and BMP of each
biodegradable waste divided by the sum of the VS fraction. This calculation can be
summarized in the following expression:
([dry mass fraction,g]i x [VS%\i x [BMP(L / gVS)]i)
[Overall BMP (L/g VS was,e)] =
^ ([dry mass fraction, g]i x [VS%]i)
i=i
(1)
where n is the number of different kinds of biodegradable materials under consideration.
The overall BMP value obtained from equation (1) represents the average methane
potential per mass of dry volatile solids in the waste. In order to correct it to the methane
potential based on the total mass of waste, the average volatile solids percentage was
multiplied by the average total solids percentage and the total BMP value obtained from
equation (1):
[BMP (L/g waste)] = [overall BMP (L/g VS waste)] x [F5%]average x [7Y%]average
Where,
(2)
Average VS (%) = ^([dry mass fraction,%]/ x [VS%]i)
1=1
(3)


APPENDIX A
ADDITIONAL PROCEDURES AND CONCEPTS
A.l Prediction of Mass Loss by Gas and Leachate
Once solid wastes are subject to land disposal, decomposition processes start.
Cellulolytic bacteria convert biodegradable macromolecules into simpler substrates,
which are accessible to other bacteria that produce gases and/or other byproducts (Figure
A-l). In anaerobic systems, methane and carbon dioxide are the final products of
decomposition, while water and carbon dioxide are the major products in aerobic
systems. Various biochemical mechanisms are involved in these processes; they can be
expressed by simple chemical equations as follows:
C6H,o05 + 602 -Â¥ 6C02 + 5H20
(aerobic)
(1)
C6H,o05 + H20 -> 3CH4 + 3C02
(anaerobic)
(2)
According to the equations, (1) and (2), 6 moles of CO2 are produced from 1 mole
of glucose, a monomer of cellulose in aerobic decomposition. Three moles each of CH4
and CO2 are produced from 1 mole of glucose in anaerobic decomposition. At standard
temperature and pressure (STP, temperature = 273K and pressure = 1 atm), 1 mole of
gas occupies 22.4L. If gas expansion by temperature around a gas totalizer (20C for
average) is taken into consideration, the volume occupied by 1 mole of gas at 20C is
calculated using the Ideal Gas Law:
Vx
Tx
(3)
149


120
settlement as a function of site specific waste decomposition. However, no data are
available to characterize waste volume loss with respect to mass loss caused by
decomposition in landfills.
The objective of this research was to collect data on mass loss versus volume loss
for future use in settlement model development. In order to correlate mass loss and
volume loss, a lab-scale experiment was designed where waste was decomposed in
simulated landfills in the laboratory, with both mass loss and volume loss being measured.
A difficulty with using lab experiments to simulate landfill settlement is that it is hard to
simulate true landfill conditions, especially, the large overburden pressure. For this
reason, in this research, the experiments conducted in this research included the
application of overburden pressure to make the laboratory condition closer to the field.
5.2 Materials and Methods
5.2.1 Lysimeters
Four simulated landfill columns (lysimeters) were used in this research, They each
consisted of a primary stainless steel column and a carriage system component. Two
lysimeters were operated as aerobic lysimeters (lysimeter 1 and 2) and two were operated
as anaerobic lysimeters (lysimeter 3 and 4). Three parameters (temperature, air addition,
and overburden pressure) were controlled in an effort to simulate actual aerobic or
anaerobic bioreactor landfills. Details about the lysimeters and the carriage system are
described in chapter 2 and in Figures B-l through B-3.
Mixed fabricated waste samples were created and loaded into the columns as four
fractions to prevent waste component stratification in a particular place in the column
(composition of the fabricated waste fractions and their weight are summarized in
appendix C). Prior to loading, 6 inches (15.3 cm) of river rock was placed at the bottom


137
Figure 5-4. Relationship between settlement and overall mass loss of the aerobic and
anaerobic lysimeters


20
above 50% at the 9th day. The biogas generation rate of lysimeter 3, after air injection,
was 2.3 L/day on average for 10 days. During the rest of test period, the pH of the
lysimeter 3 went down to 6.5, but further decrease was not observed. In comparison with
the conditions of lysimeter 3 before air addition, the amount of biogas produced was
substantially increased and high percentage of methane (> 55%) was maintained (Figure
2-16).
2.4 Discussion
2.4.1 Differences between Aerobic and Anaerobic Lysimeters
The largest differences between the aerobic and anaerobic lysimeters can be found
from the enhancement of waste biodegradation. Based on the leachate quality results of
this study, a period required for the aerobic lysimeters to decompose 90% of BOD was
160 days in the aerobic lysimeters while more than 700 days were required for the
lysimeter 4.
Other differences between the two systems were the methane concentrations
contained in exit gas. Air addition to the aerobic lysimeters lowered CH4 concentrations
in the exit gas dramatically. Though a small amount of methane was found in the exit gas
of the aerobic lysimeter, it was less than 1 % of the CO2 gas generated.
It is noted that pH had a relatively low impact on waste decomposition in the
aerobic lysimeter. Though the pH of the aerobic lysimeters was acidic, settlement
consistently occurred (see Chapter 5). It was probable that the acidic condition was
localized only on the bottom part, where air was not supplied properly.
Tables 2-3 and 2-4 present the initial and final characteristics of the aerobic and
anaerobic lysimeters. The data presented in Table 2-4 for the anaerobic lysimeters
indicate that these systems were not stabilized yet. Water loss from the aerobic lysimeters


32
Figure 2-3. Comparison of pH between aerobic and anaerobic lysimeters versus time


165
Figure B-8. Solid samples excavated from one of the aerobic lysimeter


101
After drying at 75C for at least one hour, 300 mg of sample was taken from the washed
sample and 3mL of 72% H2SO4 was added. Samples with 72% H2SO4 were then placed
in a shaking incubator at 30C for one hour. After reaction with H2SO4, samples were
transferred to a 250 mL Erlenmeyer flask and diluted with 84 mL of deionized (DI)
water. The flasks were covered with aluminum foil, and autoclaved for an hour at 121C
and 15 psi. The residues were filtered through Gooch crucibles with glass fiber filters and
washed with 100 mL of hot DI water. Washed residues were dried at 105C, weighed and
ignited at 550C for an hour. The lignin content was determined by dividing the
difference in sample weights before and after ignition by the initial sample weight.
4.2.5 Data Analysis
In order to evaluate the impact of air addition on waste decomposition, all methane
yield results were statically compared through the ANOVA test at the 0.05 level of
significance. ANOVA tests were conducted to compare the differences of the methane
yields of the lignocellulosic materials, and the cellulose and lignin analysis results of the
ground SYP blocks.
4.3 Results
4.3.1 Methane Yield of Raw Waste
By combining methane yields of all biodegradable organic fractions, the overall
methane yield of the raw waste was estimated. The methane yield of each component is
presented in Table 4-1. The overall methane yield of the organic fraction of the raw waste
calculated was 0.337 L/g VS (0.191 L/g total waste, dry). In comparison with the
methane yields of the organic fraction of MSW reported, the methane yield of the
fabricated waste here is higher than mechanical-sorted MSW but close to other hand-
sorted or source-sorted MSW (Table 4-2). Mata-Alvarez et al. (1990) pointed out that the


226
Qian, X., Koemer, R. M. and Gray, D. H., 2002, Geotechnical aspects of landfill design
and construction, Upper Saddle River, N.J., Prentice Hall.
Rai, D., Sass, B. M. and Moore, D. A., 1987, Chromium(III) hydrolysis constants and
solubility of chromium(III) hydroxide, Inorganic Chemistry, 26(3), 345-349.
Rai, D., Eary, L. E., and Zachara, J. M., 1989, Environmental chemistry of chromium,
Science of the Total Environment, 86(1-2), 15-23.
Rajwanshi, P., Singh, V., Gupta, M. K. and Dass, S., 1997, Leaching of aluminium from
cookwares a review, Environmental Geochemistry and Health, 19(1), 1-18
Ravat, C., Monteil-Rivera, F. and Dumonceau, J., 2000, Metal ions binding to natural
organic matter extracted from wheat bran: application of the surface complexation
model, Journal of Colloid and Interface Science, 225(2), 329-339.
Read, A. D., Hudgins, M., Harper, S., Phillips, P. and Morris, J., 2001, The successful
demonstration of aerobic landfilling: The potential for a more sustainable solid
waste management approach?, Resource, Conservation and Recycling, 32, 115-
146.
Reid, I. D. and Seifert, K. A., 1982, Effect of an atmosphere of oxygen on growth,
respiration, and lignin degradation by white-rot fungi, Canadian Journal of Botany-
Revue Canadienne De Botanique, 60(3), 252-260.
Reinhart, D. R. and AlYousfi, A. B., 1996, The impact of leachate recirculation on
municipal solid waste landfill operating characteristics, Waste Management &
Research, 14(4), 337-346.
Reinhart, D. R. and Chopra, M. B., 2000, MSW landfill leachate collection systems for
the new millennium, report 00-13, Florida center for solid and hazardous waste
management, Gainesville, Florida, USA
Reinhart, D. R., McCreanor, P. T. and Townsend, T., 2002, The bioreactor landfill: its
status and future, Waste Management Research 20, 172-186.
Richard, F. C. and Bourg, A. C. M., 1991, Aqueous geochemistry of chromium a review,
Water Research, 25(7), 807-816.
Rittmann, B. E. and McCarty, P. L., 2001, Environmental biotechnology : principles and
applications, Boston, McGraw-Hill.
Rowell, R., 2005, Handbook of wood chemistry and wood composites, Boca Raton, FL
CRC press,
Sadiq, M., 1997, Arsenic chemistry in soils: an overview of thermodynamic predictions
and field observations, Water Air and Soil Pollution, 93(1-4), 117-136.


147
In the presence of CRT monitor glass, the highest Pb concentration was
observed from the aerobic lysimeters during the acidic condition.
When evaluating the total masses of metals adsorbed on lignocellulosic wastes
(office paper, newspaper, cardboard and wood blocks), greater masses of Fe, Mn,
As, A1 and Cu were found from the aerobic lysimeters.
Periods of time required for 20% settlement for the aerobic and anaerobic
lysimeters were 1 and 2 years, respectively.
The relationship between percentage of settlement and mass loss based on this
research can be written as: [settlement, %] = (AH, %) = 16.90 log [mass loss,
%] 6.24
The settlement and mass loss data obtained from this research could be used for
the development of the settlement models.
There was no significant impact of air addition on wood waste decomposition as
indicated by cellulose and lignin analysis, but methane yields of the SYP blocks
of the aerobic lysimeter were significantly lower than those of the anaerobic
lysimeters.
There was no evidence that the lignin component of the SYP blocks degraded in
either the aerobic or anaerobic lysimeters.
6.4 Future Work
The influences of air added to a landfill on gas and leachate quality, metal leaching
behavior, settlement and waste decompositions were explored. Air injection providing
various results distinguished as different from those of anaerobic landfills. Except for the
metal leaching results, much of the data obtained from this research gives validity to the


12
leachate was 6.11. Lab air was injected with 70 mL/min from the bottom of the lysimeter
for five days. The changes in pH and output gas qualities were monitored on a daily
basis. The impact of this addition is discussed in the results section of this chapter.
2.2.7 Prediction of Waste Mass Loss
As described in the following sections in this chapter, the aerobic lysimeters more
quickly stabilized the waste in comparison to the anaerobic lysimeters, and thus their
period of operation was shorter (379 days vs. 741 days). In an effort to normalize the
leachate measurements among the different columns, the biogas data, the leachate data
and the initial content of the waste was used to estimate the percentage of waste
decomposition for a column at any given time. The detailed procedure for this is
presented in appendix A, but, in short, the cumulative volume of biogas measured at any
given time (CH4 + CO2 for anaerobic columns and CO2 for aerobic columns) was used to
calculate the mass of initial waste degraded at that time. This was adjusted to account for
the mass of organic carbon solubilized in the leachate. The mass of waste estimated to be
degraded at a given time was divided by the estimated total potential mass loss in each
column (this total potential mass loss was estimated from measured methane yields of the
raw waste; see chapter 4 for details).
2.3 Results and Discussion
The data presented for the aerobic and anaerobic lysimeters in this dissertation
represent operation periods of 379 and 741 days, respectively. At the end of each
operation period, one each of the aerobic and anaerobic lysimeters was stopped and
emptied. The remaining lysimeters were left operational (data are not reported here).
Values of all leachate parameters analyzed for this research are presented in Table C-l


150
22AL V2
273K ~ (273 + 20)K
(4)
Such that,
V2 = 24.0 L (5)
Based on information obtained thus far, the mass loss of waste by gas generation
can be calculated as follows:
Aerobic condition:
1 L C02 generated
lmole C02
lmole C6H10O5
162 g
24.0 L C02
6 mole C02
1 mole C6H10O5
= 1.125 g mass loss / L C02 generated (6)
Anaerobic condition:
1 L biogas (CFI4
and C02) generated
lmole
biogas
lmole C6H10O5
162 g
24.0 L
biogas
3 mole biogas
1 mole C6H10O5
= 2.25 g mass loss / L biogas generated (7)
In order to predict mass loss more precisely at given conditions, organic carbons
remaining in leachate are considered as well. Since many different types of
microorganisms involved in biodegradation are ecologically linked, the biogas produced
by final gas-producers may not always reflect mass loss caused by the first attackers.
For example, wastes consumed by acid-forming bacteria may not convert into methane
gas immediately during the acid-forming stage of anaerobic decomposition due to the low
activity of methane-generators. Most microorganisms may not consume large wastes at
first. Once long and branched structures of cellulose are hydrolyzed by cellulolytic
bacteria, a first attacker, other bacteria may follow and consume the fragments of
cellulose. For this reason, mass loss may take place as cellulolytic bacteria hydrolyze
wastes, increasing dissolved organic carbons in leachate. Gas-generators consume part of


3-4. Changes of Cu concentrations over time 75
3-5. Changes of Pb concentrations over time 76
3-6. Changes of Fe concentrations over time 77
3-7. Changes of Mn concentrations over time 78
3-8. Changes of Zn concentrations over time 79
3-9. Distribution of As over a C-pH diagram 80
3-10. Potential- pH diagram of Cr 81
3-11. Distribution of Cu over a C-pH diagram 82
3-12. Adsorption of metal on solid wastes 83
3-13. The comparison of aerobic and anaerobic lysimeters in respect of total mass of
metals adsorbed on lignocellulosic materials 85
3-14. The comparison of metal concentrations adsorbed on organic (newspaper and
cardboard) and plastic waste 87
3-15. Fate of heavy metals thermodynamically occurred in aerobic (oxidizing) and
anaerobic (reducing) conditions 89
3-16. Comparison of concentrations of metal leached between aerobic and anaerobic
lysimeters 90
3-17. Changes in cumulative mass of meta released over a mass loss, % 92
3-18. Comparison of As, Cu and Cr leaching trend of the lysimeters to other study 94
4-1. The dry weight differences between predicted and measured remaining mass 113
4-2. Comparison of dry weights between raw and decomposed lignocellulosic wastes .114
4-3. The changes in the percentage of waste components after decomposition; (A) raw
waste components and (B) decomposed waste (aerobic) 115
4-4. Changes in cumulative methane volume of lignocellulosic materials over time 116
4-5. Methane yields and weight differences of lignocellulosic materials among raw
and two lysimeters (A) all lignocellulosic materials; (B) wood only 117
4-6. The comparison of dry masses measured and predicted by gas generated and
BMP assay 118
xi


Chloride (mg/L) Chloride (mg/L)
173
Figure C-5. The change in chloride (Cl) of the aerobic and anaerobic lysimeters over
time.


Cr, mg/L Cr, mg/L
74
0.5
0.4 -
0.3
0.2 -
0.1 -
Lys 1
O Lys 2
AEROBIC
cP
0.0
i
o


e
8 % o
o
o
o *
cO
o o
n r
I I
0 50 100 150 200 250 300 350
Figure 3-3 Changes of Cr concentrations over time


Cu, mg/L
75
100
~Sb
B
10 -
kO
o
o
AEROBIC
Lys 1
O Lys 2

$
8
o o
o O o o Q
O O
0.1
%
0.1 -
0.01
0.001
50
aaa^
A AA
100
S/***t A
V a.
4fc A
Below Detection Limit
150
200
250
I
300
350
ANAEROBIC
&A
Aaa A
A
A A
A A
A
A A
A AMA A
A A
AA
A Lys 3
A Lys 4
200
Days
400
600
Figure 3-4. Changes of Cu concentrations over time


203
Table C-8. Sulfide concentrations of aerobic and anaerobic lysimeters (uniti^g/L)
date
lys 1
lys 2
date
lys 3
lys 4
8/4/2004
335
65
8/13/2003
305
310
8/12/2004
165
25
8/15/2003
485
8/17/2004
230
40
8/26/2003
29
8/24/2004
110
0
8/29/2003
36
8/31/2004
175
5
9/9/2003
190
290
9/9/2004
145
45
9/16/2003
150
180
9/14/2004
110
0
9/24/2003
160
180
9/22/2004
40
0
9/29/2003
160
180
10/1/2004
60
0
10/8/2003
100
200
10/5/2004
65
20
10/15/2003
95
155
10/15/2004
120
30
10/21/2003
120
150
10/21/2004
230
10
10/28/2003
80
155
10/31/2004
125
45
11/5/2003
85
130
11/12/2004
95
15
11/12/2003
80
145
11/19/2004
65
5
11/19/2003
85
120
12/2/2004
45
55
11/26/2003
90
155
1/13/2005
345
125
12/3/2003
75
110
1/22/2005
820
645
12/10/2003
85
145
1/30/2005
1375
500
12/17/2003
65
135
2/8/2005
1063
563
12/19/2003
50
100
2/16/2005
425
575
1/2/2004
55
130
2/24/2005
125
1250
1/7/2004
65
125
3/1/2005
225
905
1/22/2004
80
110
3/12/2005
275
540
1/29/2004
70
85
3/20/2005
530
725
2/10/2004
88
136
3/25/2005
500
775
2/25/2004
80
120
4/3/2005
705
350
3/12/2004
120
108
4/13/2005
625
3/14/2004
46
117
4/25/2005
1450
900
6/1/2004
28
73
5/3/2005
2375
1225
7/28/2004
75
90
5/12/2005
2575
1025
8/4/2004
65
105
6/1/2005
2650
1075
8/12/2004
10
75
6/8/2005
2375
1075
8/17/2004
50
95
6/16/2005
2500
1150
8/24/2004
40
35
6/29/2005
2450
875
8/31/2004
45
75
7/11/2005
2275
875
9/9/2004
80
40
7/19/2005
2200
950
9/14/2004
35
75
7/27/2005
1950
900
9/22/2004
0
50
10/1/2004
15
60
10/5/2004
15
90
10/15/2004
40
90
10/21/2004
30
55
10/31/2004
115
140
11/12/2004
15
11/19/2004
35
105
12/2/2004
50
50


17
day 180, but they were still approximately 4 times lower than values of anaerobic
lysimeters. The trend of increases in ammonia concentrations also can be found in
operating bioreactor landfills (Reinhart and AlYousfi, 1996).
Since ammonia is generally produced from the deamination process of amino
acids (a monomer of proteins), elevated ammonia concentrations may be associated with
protein decomposition. Cali et al. (2005) reported that an active methanogenic bacteria
community increased the ammonia concentration. Several researchers have proposed that
the enhancement of waste decomposition and leachate recirculation in anaerobic
bioreactor landfills results in increased ammonia concentration (Reinhart and Al-Yousfi,
1996; Berge et al., 2005).
2.3.4 Dissolved Solids Content
Figure 2-9 shows the change in total dissolved solids (TDS) in the aerobic and
anaerobic lysimeters through the course of the experiment. For both aerobic and
anaerobic lysimeters, like the change in other organic matter, TDS concentrations were
lower as the pH was stabilized. TDS of the aerobic lysimeters were rapidly stabilized
approximately 8 to 10 g/L after day 200. TDS of the anaerobic lysimeters were still
greater than that of the aerobic lysimeters but TDS of lysimeter 4 fell below 20 g/L on
day 700. Typical TDS concentration in landfills is within the range of 2 to 60 g/L
(Kjeldsen, 2002).
Figure 2-10 depicts the change in alkalinity in the aerobic and anaerobic
lysimeters through the course of the experiment. For the aerobic lysimeters, the alkalinity
increased to 16,000 mg/L as CaCC>3 in the lysimeter 1, but another lysimeter showed low
alkalinity which was below 2,000 mg/L as CaCC>3 but it increased 8,000 mg/L as CaCC>3
again. The alkalinity was lowered below 2000 mg/L as CaCC>3 for both aerobic


204
Table C-8 (continued)
date
lys 1
lys 2
date
lys 3
lys 4
12/20/2004
0
65
1/13/2005
15
100
1/22/2005
0
355
1/30/2005
45
340
2/8/2005
85
400
2/16/2005
35
445
2/24/2005
55
485
3/1/2005
60
545
3/12/2005
55
700
3/20/2005
95
740
3/25/2005
75
1205
4/3/2005
75
975
4/13/2005
0
1014
4/25/2005
80
1165
5/3/2005
60
1420
5/12/2005
100
1350
6/1/2005
115
2400
6/8/2005
95
2525
6/16/2005
135
2400
6/29/2005
155
2450
7/11/2005
180
2150
7/19/2005
250
2350
7/27/2005
325
2750


ACKNOWLEDGMENTS
I would like to thank my advisor, Dr. Timothy G. Townsend, for showing such
great patience as a mentor. He gave me this great opportunity to study on the field of
solid waste. He also showed me the way of living as an engineer, professor, and a family
man. I cannot forget his tears when Mr. Townsend passed away. I would also like to
thank my committee members, Dr. Angela Lindner, Dr. Frank Townsend and Dr. Roger
Nordstedt, and my other spectacular faculty members, Dr. David Chynoweth, Dr. Gabriel
Bitton and Dr. Matthew Booth, who gave me great help.
I wish to thank my colleagues in the Solid and Hazardous Waste Research group, in
particular, Brajesh Dubey, Qiyong Xu, Kim Cochran, Steve Musson, Aaron Jordan,
Pradeep Jain, Jaehak Ko, Murat Semiz, Judd Larson, and Yong-Chul Jang, a faculty
memeber of Chung-Nam University in South Korea. I also thanks goes to my first mentor
and graduate advisor, professor, Byung-Ki Hur, a faculty of Inha University in South
Korea.
A special thanks goes to my mother as well as my father, who is fighting against
disease. Finally, greatest thanks go to my wife, Eunkyoung Choi, for her patience,
encouragement, and love.
IV


215
Table C-16 (continued)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
10/26/2004
0.48
1.11
0.03
0.02
539.75
6.61
0.09
227.72
11/6/2004
0.12
0.48
0.01
0.00
287.44
3.41
0.04
160.68
11/11/2004
0.45
1.13
0.03
0.02
597.42
6.50
0.10
220.40
11/24/2004
0.10
0.47
0.01
0.00
312.82
3.25
0.04
145.12
12/8/2004
0.39
1.02
0.03
0.02
585.70
5.77
0.09
193.25
1/6/2005
0.07
0.34
0.02
0.00
292.58
2.21
0.03
84.95
1/13/2005
0.30
0.80
0.05
0.02
591.98
4.27
0.08
135.46
1/21/2005
0.06
0.35
0.02
0.00
300.24
2.13
0.03
79.30
2/5/2005
0.16
0.72
0.04
0.01
539.54
3.69
0.06
121.26
2/16/2005
0.19
0.83
0.04
0.02
517.02
3.41
0.06
100.37
2/24/2005
0.26
0.83
0.04
0.02
486.62
3.17
0.06
94.80
3/1/2005
0.34
0.94
0.04
0.02
519.69
3.29
0.06
94.12
3/12/2005
0.15
0.89
0.04
0.02
468.42
2.94
0.05
85.92
3/20/2005
0.23
0.97
0.04
0.02
465.67
2.84
0.05
79.82
3/26/2005
0.14
0.90
0.04
0.02
441.47
2.69
0.04
80.03
4/3/2005
0.15
0.88
0.04
0.02
407.69
2.44
0.04
69.79
4/17/2005
0.12
0.84
0.04
0.02
356.34
2.11
0.04
60.49
4/24/2005
0.17
0.33
0.02
0.01
58.06
3.29
0.03
57.00
4/30/2005
0.18
0.29
0.02
0.02
89.49
3.22
0.03
67.41
5/7/2005
0.07
0.26
0.02
0.02
179.21
2.66
0.02
63.49
5/31/2005
0.24
0.49
0.04
0.02
335.52
2.47
0.04
61.55
6/6/2005
0.03
0.46
0.04
0.01
253.10
1.78
0.02
54.65
6/14/2005
0.00
0.41
0.03
0.00
211.36
1.37
0.02
45.09
6/28/2005
0.00
0.51
0.03
0.00
208.26
1.24
0.01
39.54
7/5/2005
0.04
0.67
0.03
0.00
220.61
1.29
0.03
37.57
7/11/2005
0.05
0.70
0.04
0.00
208.28
1.21
0.02
33.97
7/27/2005
0.06
0.72
0.04
0.00
165.58
0.88
0.01
23.89


96
samples. Generally, wood wastes are classified as recalcitrant materials in anaerobic
landfill conditions because of the high content of lignin (Chandler et al., 1980; Holt and
Jones, 1983; Stinson and Ham, 1995). The lignin content in wood differs according to
species, but typically, 27-33% of softwood species such as Douglas fir and Southern Pine
is composed of lignin, and 18-20% of hardwood species is composed of lignin (White,
1987).
Aerobic bioreactor landfills, which introduce air into a landfill to more rapidly
stabilize lignocellulosic wastes, are reported to have same advantages over conventional
anaerobic landfill (Reinhart et al, 2002). It is reported that lignin can be degraded through
the pretreatment or composting process of lignocellulosic materials (Fox and Noike,
2004; Vikman et al., 2002). Stinson and Ham (1995) reported that the degradation of total
cellulose contained in high lignin-containing materials such as newspaper dramatically
increased as lignin content decreased. If it is possible to decompose and/or depolymerize
lignin in aerobic landfills, greater decomposition of lignocellulosic materials can be
expected, which will lead to a reduction in the ultimate disposal capacity needed for the
landfill.
In this research, the biodegradation of lignocellulosic waste under aerobic and
anaerobic landfill conditions was evaluated. As described in chapter 2, four stainless steel
lysimeters were operated, two each aerobic and anaerobic bioreactors. Two of these (one
aerobic and one anaerobic) were excavated at the end of the experiment. The methane
yields of various lignocellulosic wastes contained in the lysimeters were measured. In
addition to the methane yields, the cellulose and lignin content of the wood waste


BIOGRAPHICAL SKETCH
Hwidong Kim was bom on Dec 17, 1970 in Wonju, South Korea. He graduated
with a Bachelor of Engineering degree from the Department of Biochemical Engineering
in 1992 from Inha University, Inchon, Korea. He enrolled the graduate school of Inha
University in the spring of 1993. After 1 academic year of graduate study, he enlisted in
military service for two years from 1994 through 1996. He received Masters degree in
Biochemical Engineering from Inha University in 1997.
He enrolled in the graduate program in the Department of Agricultural and
Biological Engineering at the University of Florida in August, 1999 and transferred to the
Department of Environmental Engineering Sciences to study solid and hazardous waste.
He earned a graduate research assistantship and teaching assistantship to complete his
study.
231


Metal concentrations (mg/kg)
87
Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic) (organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic) (organic) (organic) (platic) (plastic)
Figure 3-14. The comparison of metal concentrations adsorbed on organic (newspaper
and cardboard) and plastic waste


LIST OF TABLES
Table page
2-1. MSW components 25
2-2. Parameters and methods for analysis 26
2-3. Comparison of initial and final characteristics of the aerobic lysimeters 27
2-4. Comparison of initial and final characteristics of the anaerobic lysimeters 28
2-5. Comparison of leachate parameters with other aerobic landfill studies 29
2-6. Comparison of leachate parameters with other anaerobic landfill studies 29
3-1. Heavy metal sources in fabricated waste stream 69
3-2. Results of statistical analysis of metal leached between aerobic and anaerobic 69
3-3. The amount of leachate produced and used for analysis 69
3-4. Leachability of As, Cr, and Cu 70
3-5. Comparison of cumulative mass of metal dissolved in leachate and adsorbed on
lignocellulosic materials 70
3-6. Comparison of average metal concentrations of the aerobic and anaerobic
lysimeters with MSW leachate and regulatory levels 71
3-7. Comparison of characteristics of CCA-treated wood used for Jambeck (2004) and
this study 71
4-1. Methane yields, VS and mass fraction of the lignocellulosic materials in raw
waste 109
4-2. Comparison of methane yields of MSW with other studies 109
4-3. Biodegradable volatile solid (BVS) of organic fraction of the raw waste 109
4-4. The physical characteristics of excavated waste 110
4-5. Overall methane yields of waste layers of the lysimeters 2 and 4 Ill
viii


192
Table C-2. Conductivity of the aerobic and anaerobic lysimeters (unit: fas)
date
lys 1
lys 2
date
lys 3
lys 4
10/8/2004
17800
1900
8/8/2003
12540
28000
10/26/2004
15200
1700
8/13/2003
17300
17900
11/6/2004
17500
2600
8/15/2003
16400
16500
11/13/2004
16100
5400
8/19/2003
14560
15380
11/24/2004
10500
7000
8/22/2003
13630
14890
12/8/2004
12000
2000
8/26/2003
16100
16200
12/20/2004
2500
4700
9/9/2003
17300
17600
1/5/2005
12200
15300
9/12/2003
17200
17600
1/17/2005
8600
12100
9/16/2003
15440
15300
1/25/2005
11200
7600
9/19/2003
19680
1302
1/30/2005
7600
3300
9/24/2003
18200
18000
2/5/2005
7400
2200
9/26/2003
21900
20200
2/8/2005
6800
5800
9/29/2003
20100
19600
2/16/2005
8100
6600
10/3/2003
22400
> 20000
2/24/2005
7600
9800
10/8/2003
16900
19300
3/1/2005
5300
9200
10/10/2003
19000
> 20000
3/12/2005
7100
8300
10/15/2003
19700
22800
3/20/2005
7500
7800
10/17/2003
18600
19400
3/26/2005
6500
7600
10/21/2003
19000
22300
4/3/2005
7500
7600
10/26/2003
14800
16600
4/10/2005
7600
7400
10/31/2003
17100
22600
4/17/2005
8500
6800
11/5/2003
16800
18600
4/18/2005
8300
6800
11/7/2003
17200
19000
4/30/2005
10000
7000
11/12/2003
19400
19000
5/7/2005
9400
7200
11/15/2003
17400
22400
5/16/2005
7800
6600
11/19/2003
17000
19500
5/23/2005
8000
6300
11/26/2003
18000
5/31/2005
8400
3900
11/28/2003
18000
22100
6/6/2005
8200
6800
12/3/2003
17600
18100
6/14/2005
8700
6500
12/10/2003
16000
18500
6/28/2005
8400
7600
12/17/2003
16000
17800
7/5/2005
8000
6100
12/23/2003
14230
15940
7/12/2005
7700
7200
1/2/2004
15400
16000
7/19/2005
6700
6200
1/7/2004
19200
1/14/2004
18600
17520
2/12/2004
12400
12400
2/25/2004
10910
11040
3/16/2004
11700
11400
3/19/2004
12200
14500
6/1/2004
13500
11100
10/8/2004
16300
13600
10/19/2004
16500
15100
10/26/2004
15900
15300
11/6/2004
17200
15800
11/13/2004
16500
17200
11/24/2004
19200
14000


Settlement, % Settlement, %
138
(a)
Figure 5-5. Relationship between percentage of settlement and mass loss


Copyright 2005
by
Hwidong Kim


190
Table C-l (continued'
date
lys 1
lys 2
date
lys 3
lys 4
4/30/2005
9.01
8.814
1/22/2004
5.222
5.298
5/7/2005
9.161
8.752
1/29/2004
5.189
5.295
5/16/2005
8.855
8.563
2/12/2004
5.157
5.182
5/23/2005
8.839
8.665
2/25/2004
5.138
5.206
5/31/2005
8.497
8.573
3/16/2004
5.14
5.168
6/6/2005
8.564
8.549
3/19/2004
5.093
5.134
6/14/2005
8.59
8.55
5/9/2004
5.05
5.094
6/28/2005
8.4
8.555
6/1/2004
5.214
5.246
7/5/2005
8.586
8.518
7/28/2004
5.069
5.198
7/12/2005
8.587
8.552
8/4/2004
5.129
5.046
7/19/2005
8.77
8.78
8/12/2004
5.167
5.203
7/27/2005
8.545
8.621
8/18/2004
5.235
5.233
8/10/2005
8.518
8.533
8/24/2004
5.203
5.161
8/31/2004
5.239
5.182
9/10/2004
5.264
5.178
9/15/2004
5.15
5.078
9/23/2004
5.273
5.148
10/8/2004
5.582
5.43
10/19/2004
5.61
5.432
10/26/2004
5.343
5.206
11/6/2004
5.448
5.252
11/13/2004
5.442
5.276
11/24/2004
5.521
5.308
12/8/2004
5.527
5.352
12/13/2004
6.2
5.8
12/16/2004
6.2
5.7
12/20/2004
6.153
6.264
1/5/2005
5.8
5.8
1/13/2005
5.875
5.97
1/17/2005
5.532
5.54
1/21/2005
5.8
5.95
1/25/2005
5.892
5.982
1/30/2005
5.855
5.993
2/5/2005
5.93
6.24
2/8/2005
5.92
6.31
2/16/2005
5.89
6.471
2/24/2005
5.881
6.7
3/1/2005
5.897
6.761
3/12/2005
5.95
6.9
3/20/2005
6.04
7.18
3/26/2005
6.102
7.292
4/3/2005
6.071
7.25
4/10/2005
6.105
7.485
4/17/2005
6.116
7.338
4/24/2005
7.182
7.377
4/30/2005
6.873
7.392


210
Table C-12. (continued)
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
6/14/2005
4462.3
4249.4
630.0
247.1
6/24/2005
4883.7
6142.8
782.9
87.4
7/5/2005
2960.0
5356.2
623.3
51.5
7/11/2005
1712.2
3791.7
423.2
32.9
7/19/2005
1729.3
4328.7
333.5
79.8


211
Table C-13. Fabricated waste in lysimeters (unit: g)
CB
NP
cullets
CRT
A1
Plastic
lys 1 fraction 1
544.4
205.5
170.1
34
136.4
510.6
lys 1- fraction 2
544.3
204.1
170.1
34
136.1
510.3
lys 1- fraction 3
544.3
204.1
170.1
34
136.2
510.4
lys 1- fraction 4
544.4
204.3
170.1
34
136.1
510.3
lys 2- fraction 1
544.3
204.1
170.1
34
136.2
510.3
lys 2- fraction 2
544.3
204.2
170.1
34
136.2
510.3
lys 2- fraction 3
544.3
204.1
170.1
34
136.1
510.3
lys 2- fraction 4
544.3
204.3
170.2
34
136.2
510.3
Lys 3 Fraction 1
544.5
204.1
188.2
34
136.7
510.4
Lys 3 Fraction 2
544.1
204.5
187.8
34
136.6
510.7
Lys 3 Fraction 3
544.5
204.4
186.6
34
136.3
510.6
Lys 3 Fraction 4
544.2
204.1
187.4
34
136.2
510.6
Lys 4 Fraction 1
544.7
204.8
187.7
34
136.5
510
Lys 4 Fraction 2
544.2
204.3
187.1
34
136.2
510
Lys 4 Fraction 3
544.4
204.6
187
34
136
510.7
Lys 4 Fraction 4
544.2
204
187.1
34
136
510
steel
OP
dogfood
SYP
CCA
Total
lys 1 fraction 1
136.1
952.5
510.8
171
34.1
3405.0
lys 1 fraction 2
136.2
952.5
510.4
170
34
3402.2
lys 1- fraction 3
136.2
952.5
510.4
170
34
3402.5
lys 1- fraction 4
136.2
952.5
510.4
171
34.1
3402.9
lys 2- fraction 1
136.1
952.5
510.3
170
34.1
3402.2
lys 2- fraction 2
136.1
952.5
510.3
170
34
3402.1
lys 2- fraction 3
136.2
952.5
510.3
170
34
3402.0
lys 2- fraction 4
136.2
952.5
510.3
171
34
3402.8
Lys 3 Fraction 1
136.2
953
510
170
34
3420.9
Lys 3 Fraction 2
136.5
953
510
170
34
3421.0
Lys 3 Fraction 3
136.6
953
510
170
34
3419.8
Lys 3 Fraction 4
136
952.6
510
170
34
3419.0
Lys 4 Fraction 1
136.4
953
510
170
34
3421.0
Lys 4 Fraction 2
136.4
953
510
170
34
3419.4
Lys 4 Fraction 3
136.3
953.5
510
170
34
3420.9
Lys 4 Fraction 4
136
953
510
170
34
3417.8


57
of sulfur. Lead is generally present in an ionic form at a pH < 6 under oxidizing
conditions (Drever, 1988).
As shown in Figure 3-5, Pb leached from the aerobic lysimeters significantly
greater than from the anaerobic lysimeters. Most samples with high concentrations were
distributed in the acidic phases. This leaching pattern corresponds to the characteristics of
Pb previously discussed. For both aerobic and anaerobic lysimeters, Pb concentrations
decreased with an increase in pH. Most Pb in alkali conditions may be precipitated as
forms of PbCC>3 or PbS depending upon redox potentials.
3.3.1.6 Iron
As shown on Figure 3-6, initial Fe concentrations of the aerobic lysimeters
(110mg/L for both aerobic lysimeters) were higher than those of the anaerobic lysimeters
(20-22mg/L). Iron concentrations of the aerobic lysimeters increased to 250mg/L on the
30th day and then gradually decreased. During changes in pH of the aerobic lysimeters to
alkali conditions, Fe concentrations lowered substantially to below 10mg/L. In contrast to
the aerobic lysimeters, Fe concentrations of the anaerobic lysimeters increased from 20 to
600mg/L for the first 450 days and then decreased with increasing pH. Although
lysimeters 3 and 4 are both anaerobic lysimeters, the final Fe concentrations were
substantially different (165mg/L and 5.6 mg/L for lysimeters 3 and 4, respectively).
Generally, free Fe concentration is strongly associated with the redox condition of
the system. Iron is present in aquatic systems in two oxidation states; Fe (III) and Fe (II).
Ferric (Fe3+) and ferrous (Fe2+) irons can be transformed to each other depending upon
the redox conditions. Iron (III) is precipitated as a mineral deposit such as Fe23 or
Fe(OH)3 at a pH > 5. Iron (III) is also involved in complexation with metals. Under
moderately oxidizing and reducing condition, Fe (II) ions are dominant in the pH range


216
Table C-17. Metal concentrations of lysimeter 4 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
8/3/2003
9.36
2.47
0.64
0.23
20.55
6.75
0.02
78.37
8/13/2003
10.82
2.77
0.72
0.26
23.00
7.61
0.01
86.55
8/15/2003
12.97
2.47
0.64
0.22
20.80
7.25
0.01
81.52
8/19/2003
18.42
2.53
0.65
0.19
29.50
7.18
0.04
83.08
8/22/2003
10.57
2.61
0.58
0.16
30.88
7.26
0.02
79.50
8/26/2003
5.79
2.83
0.52
0.14
34.59
7.69
0.02
84.37
9/9/2003
2.70
2.17
0.37
0.08
32.60
6.66
0.03
85.14
9/17/2003
3.15
2.67
0.42
0.18
38.39
7.96
0.04
103.91
9/24/2003
3.70
2.99
0.41
0.08
32.52
8.33
0.02
138.19
10/9/2003
3.11
3.01
0.43
0.09
34.50
8.84
0.04
152.48
10/15/2003
2.57
2.79
0.39
0.06
31.39
8.43
0.06
156.50
10/21/2003
2.80
3.24
0.43
0.09
36.51
9.69
0.09
166.70
10/28/2003
1.11
3.23
0.45
0.08
35.60
9.47
0.03
332.44
11/5/2003
0.99
3.19
0.42
0.06
34.97
9.74
0.03
365.16
11/12/2003
0.88
3.12
0.40
0.07
35.35
9.80
0.02
376.90
11/19/2003
0.71
2.89
0.34
0.06
33.88
9.40
0.02
389.22
11/26/2003
0.71
3.04
0.34
0.05
36.16
9.97
0.03
402.26
12/3/2003
0.65
3.18
0.34
0.05
38.87
10.60
0.02
417.96
12/10/2003
0.40
2.18
0.29
0.04
29.52
9.35
0.02
392.13
12/17/2003
0.46
2.89
0.29
0.03
39.36
9.85
0.03
406.29
12/23/2003
0.54
2.97
0.27
0.04
44.25
10.11
0.02
418.35
1/2/2004
0.49
2.77
0.26
0.03
47.92
9.61
0.02
407.58
1/9/2004
0.48
2.80
0.24
0.03
52.53
9.85
0.02
415.38
1/14/2004
0.45
2.68
0.22
0.03
56.31
9.38
0.02
406.26
1/22/2004
0.42
2.60
0.21
0.03
63.99
9.20
0.02
407.18
2/12/2004
0.43
2.50
0.20
0.03
85.93
9.59
0.01
455.51
3/16/2004
0.36
2.20
0.17
0.02
77.91
8.14
0.02
380.16
3/26/2004
0.44
2.66
0.21
0.03
89.91
9.87
0.02
422.14
6/1/2004
0.42
1.45
0.10
0.01
89.47
6.59
0.02
353.83
7/28/2004
0.38
1.27
0.08
0.01
111.89
5.99
0.02
328.98
8/9/2004
0.00
0.28
0.02
0.00
244.59
1.70
0.03
60.70
8/12/2004
0.23
1.12
0.10
0.02
129.38
5.66
0.12
290.65
8/17/2004
0.29
0.88
0.06
0.02
116.11
4.58
0.01
263.51
8/24/2004
0.22
1.11
0.10
0.02
158.65
5.84
0.02
288.44
9/3/2004
0.23
0.73
0.05
0.00
117.19
3.99
0.01
234.78
9/10/2004
0.21
1.09
0.08
0.01
194.40
6.17
0.02
300.74
9/15/2004
0.21
0.68
0.05
0.00
127.40
3.91
0.01
228.76
9/23/2004
0.16
0.98
0.07
0.01
197.70
5.77
0.02
286.15
10/7/2004
0.17
0.60
0.04
0.00
150.94
3.75
0.01
212.72
10/14/2004
0.23
1.06
0.08
0.01
288.81
6.66
0.03
309.69
10/19/2004
0.17
0.58
0.04
0.00
165.30
3.71
0.01
211.07
10/26/2004
0.04
0.65
0.05
0.00
200.54
4.26
0.02
227.80
10/30/2004
0.00
0.70
0.02
0.00
82.41
0.71
0.01
31.18
11/6/2004
0.22
0.60
0.04
0.00
203.20
3.99
0.02
220.41
11/11/2004
0.08
0.66
0.04
0.01
281.63
4.81
0.03
239.09
11/24/2004
0.20
0.63
0.04
0.00
363.98
4.46
0.03
210.94


Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy
COMPARATIVE STUDIES OF AEROBIC AND ANAEROBIC LANDFILLS
USING SIMULATED LANDFILL LYSIMETERS
By
Hwidong Kim
December 2005
Chair: Timothy G. Townsend
Major Department: Department of Environmental Engineering Sciences
Many proposals suggest that air injection into bioreactor landfills enhance waste
composition; several potential benefits of air addition have been hypothesized, yet little
has been proven about the overall performance of aerobic landfills compared with current
anaerobic landfills. Utilizing research conducted with six-foot tall stainless steel
simulated landfill lysimeters, complete with fabricated wastes, this Ph.D. dissertation
compares aerobic and anaerobic landfills with respect to gas and leachate quality, fate of
metals, settlement behavior and biodegradation of lignocellulosic materials.
Through air injection, a large enhancement of waste decomposition was observed.
More than 90% of the maximum chemical oxygen demand (COD), biochemical oxygen
demand (BOD) and total organic carbon (TOC) concentrations decreased within 100 days.
During the methanogenic phase in the anaerobic condition, concentrations of ammonia
increased by an amount four times greater than the initial concentrations. A large change
of ammonia was not observed from the aerobic lysimeters.
xiv


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23
Without high alkalinity, the pH of the landfill may decrease again when air addition is
stopped. This could happen when methanogenic bacterial population was not enough to
adapt to the new condition.
As Reinhart et al. (2002) pointed out, the reduction of leachate volume due to air
stripping could be one of the advantages of the aerobic landfills. In this research, a total
of 31m3 of air was added during a test period (1 year). Comparing the volume of water
initially added with final leachate volume, approximately 21 % of leachate volume was
reduced. Reduced volume of leachate implies that the operation of aerobic landfills can
be economical in terms of saving the cost for the leachate treatment.
2.4.4. Limitations
Since CH4 gas is one of the gases causing global warming, CH4 reduction can be
one of the advantages of the aerobic lysimeters. However, landfill gas released without a
flare system could be adverse to the environment. Berge et al. (2005) and Reinhart et al.
(2002) pointed out that various kinds of volatile organic compounds (VOC) and nitrous
oxide, a more potent greenhouse gas than methane, can be emitted without the flare
system. Future research is required to identify the gas constituents and develop the
filtering system as an alternative.
As previously mentioned, an extra monitoring job may be required to check
moisture content and gas contents around the gas injection well. Certain ratios of methane
and oxygen can be flammable according to Coward and Jones (1952) and Liao et al.
(2005). While air was added, CH4 concentrations were low, but the unpredictable
changes in O2 and CH4 concentrations were observed from the aerobic landfill (Read et al,
2001) and high concentrations of CH4 and 02 could coexist when air addition starts (Lee
et al, 2002).


1600
1400
1200
1000
800
600
400
200
0
14000
12000
10000
8000
6000
4000
2000
0
:e C-8
176
Lys 1
O Lys 2
*
O

o
o


ko
o*
o
o o
o
o , Qz>P-
50
100
150
200
i
250
i
300
350
Lys 3
O Lys 4
Day
800
change in sodium (Na) of the aerobic and anaerobic lysimeters over time


218
Table C-18. ANOVA results of metals and organic absorbence
A1 (mg/L)
As
Cr
Cu
CB
aerobic
Avg
108.627
0.266
0.576
0.603
anaerobic
Avg
18.536
0.092
0.351
0.216
P
2.6E-4
8.22E-08
4.7E-4
2.72E-09
NP
aerobic
Avg
91.152
0.223
0.412
0.486
anaerobic
Avg
33.307
0.112
0.427
0.317
P
0.012
0.001
0.744
0.023
OP
aerobic
Avg
52.442
0.128
0.254
0.222
anaerobic
Avg
22.368
0.086
0.252
0.174
P
0.025
0.007
0.925
0.040
PL
aerobic
Avg
7.584
0.019
0.051
0.026
anaerobic
Avg
2.010
0.023
0.071
0.053
P
0.003
0.443
0.076
0.015
WD
aerobic
Avg
3.326
0.047
0.068
0.097
anaerobic
Avg
1.555
0.026
0.159
0.031
P
0.150
0.048
0.001
0.004
Fe
Mn
Pb
Zn
CB
aerobic
Avg
66.453
1.175
0.125
21.936
anaerobic
Avg
21.938
0.206
0.117
15.392
P
7.52E-05
1.13E-08
0.701
0.031
NP
aerobic
Avg
42.885
0.801
0.070
14.747
anaerobic
Avg
22.349
0.227
0.180
20.363
P
0.044
0.002
0.060
0.068
OP
aerobic
Avg
41.314
0.648
0.061
13.558
anaerobic
Avg
33.843
0.301
0.348
17.405
P
0.242
0.001
0.008
0.064
PL
aerobic
Avg
2.196
0.071
0.010
1.173
anaerobic
Avg
8.273
0.073
0.035
5.074
P
0.008
0.942
0.001
0.001
WD
aerobic
Avg
26.832
0.206
0.017
4.261
anaerobic
Avg
10.274
0.050
0.022
5.027
P
0.131
0.031
0.487
0.556


189
C.2 Raw Data
Table C-l.
dH of the aerobic and anaerobic
lysimeters
date
lys 1
lys 2
date
lys 3
lys 4
7/28/2004
5.669
5.695
8/8/2003
4.546
4.868
8/4/2004
5.29
5.373
8/13/2003
5.342
5.457
8/12/2004
5.14
5.07
8/15/2003
5.237
5.321
8/18/2004
5.264
4.862
8/19/2003
4.971
5.308
8/24/2004
5.45
4.562
8/22/2003
5.157
6.063
8/31/2004
5.448
4.747
8/26/2003
5.54
5.926
9/10/2004
5.389
4.747
9/9/2003
5.31
5.49
9/15/2004
5.25
4.569
9/12/2003
5.239
5.471
9/23/2004
5.316
4.395
9/16/2003
5.318
5.447
10/8/2004
5.66
4.871
9/19/2003
5.233
5.303
10/19/2004
5.663
4.9
9/24/2003
5.186
5.262
10/26/2004
5.471
4.66
9/26/2003
5.159
5.205
11/6/2004
5.525
4.995
9/29/2003
5.262
5.257
11/13/2004
5.442
5.05
10/3/2003
5.196
5.26
11/24/2004
5.396
5.329
10/8/2003
5.206
5.242
12/8/2004
5.156
5.509
10/10/2003
5.146
5.193
12/20/2004
5.006
5.42
10/15/2003
5.195
5.243
1/5/2005
5.373
5.855
10/17/2003
5.212
5.195
1/7/2005
5.4
5.94
10/21/2003
5.206
5.213
1/9/2005
5.296
5.932
10/26/2003
5.246
5.23
1/10/2005
5.494
5.946
10/28/2003
5.143
5.171
1/13/2005
5.758
6.149
10/31/2003
5.158
5.158
1/17/2005
7.474
6.16
11/5/2003
5.142
5.165
1/21/2005
7.14
6.45
11/7/2003
5.151
5.177
1/25/2005
7.404
6.962
11/12/2003
5.177
5.221
1/30/2005
8.279
8.646
11/15/2003
5.154
5.172
2/5/2005
8.71
8.82
11/19/2003
5.11
5.16
2/8/2005
8.82
8.87
11/26/2003
5.094
5.1
2/16/2005
8.88
8.86
11/28/2003
5.038
5.051
2/24/2005
9.165
9.024
12/3/2003
5.288
5.357
3/1/2005
8.973
9.045
12/10/2003
5.088
5.151
3/12/2005
8.974
9.015
12/17/2003
5.161
5.19
3/20/2005
8.85
8.89
12/19/2003
5.11
5.204
3/26/2005
8.908
8.926
12/23/2003
5.143
5.179
4/3/2005
8.972
8.883
1/2/2004
5.214
5.255
4/10/2005
8.998
8.87
1/7/2004
5.08
5.274
4/17/2005
8.982
8.851
1/14/2004
5.251
5.271
4/18/2005
9.035
8.774
1/16/2004
5.126
5.193


191
Table C-l (continued
date
lys 1
lys 2
date
lys 3
lys 4
5/7/2005
6.822
7.436
5/16/2005
6.724
7.305
5/23/2005
6.658
7.363
5/31/2005
6.622
7.597
6/6/2005
6.544
7.357
6/14/2005
6.493
7.368
6/28/2005
6.501
7.431
7/12/2005
6.459
7.471
7/19/2005
6.57
7.46
7/27/2005
6.468
7.362
8/10/2005
6.459
7.372


164
Figure B-7. (A) Blue water phenomenon observed from gas collection system of
aerobic lysimeters; (B) a hole on copper tube caused by corrosion of Cu


2!
ero'
c
-!
OI
U>
K)
rT
O
3
5'
c
n>
D-
Adsorption capacity (mg/g)
o
o
p
La O La
to to u> p
O La O La
£
N
3
i.
r
¡III
^ o z o
ti ti m
800
Adsorption capacity (mg/g)
ooooooppp
ooooo *r*r
ON)4^0s00OK)^0v
00
-P^


118
80
-a
-a
ed
Ui
60
u
-o
C/
>
m
60
40 -
20 -
Based on mass
Y////A Based on gas produced
Based on methane yield
Anaerobic
t = 2 years
Figure 4-6. The comparison of dry masses measured and predicted by gas generated and
BMP assay


8
for the application of an external load to the fabricated waste. The carriage system
consisted of a hydraulic cylinder, carriage, steel shaft, and steel plate. A small port
located on the top flange was used as a pathway for liquid addition. Perforations in the
steel plate allowed added liquid to percolate into the waste (see Figure B-2 for a detail of
the carriage system).
2.2.2 Temperature Control
The temperature at the center of a full-scale landfill usually remains constant
because the garbage and cover soil serve to insulate the system (McBean et al., 1995). In
a laboratory environment, however, the heat produced by biologically degrading waste is
not sufficient to maintain a temperature close to those normally encountered in a landfill.
Thus, a temperature control system was designed and constructed (Figure B-3).
The temperature of each lysimeter was measured using a type T thermocouple wire
(SRT201-160, Omega) fixed on the outside of each lysimeter. Two temperature
controllers (MC240, Electrothermal) were utilized in series to maintain desired
temperatures without extreme fluctuations. The lysimeters were insulated with 5-cm-
thick fiberglass and bubble insulation to minimize heat loss. Prior to operation, the
lysimeters were filled with tap water and the temperature controllers were tested by
measuring the temperature of the water.
The temperature of the aerobic lysimeters was maintained at a constant 55C for
the entire operating period. The anaerobic lysimeters were started at 35C and at day 400,
the temperature was increased from 35C to 55C at a rate of 2C per day. Although 55C
is in the optimum range for thermophilic anaerobic waste decomposition (Rittmann and
McCarty, 2001) and is often encountered in landfills (Watsoncraik et al., 1994;


29
Table 2-5. Comparison of leachate parameters with other aerobic landfill studies
Parameters
(mg/L
except for
pH)
Compost
study3
Lysimeter
study lb
Lysimeter
study 2C
Lysimeter
study 3d
Lysimeter
study 4e
This study
Air flow
rate
20L/min
(once per a
week)
38L/min for
30min. at
12-h
intervals
8.4-
1300L/min
20mL/min
70-
120mL/min
COD
2434-31812
861-22026
130-23000
500-5000
2-1000
3400-47000
BODs
8-11571
100-10000
10-2000
30-45000
Ammonia
98-558
260-630
7-400
2-100
40-700
TDS
3300-11400
700-7700
EH
7.1-8.2
5.17-7.98
5.24-7.5
7-9
7.8
4.5-9.1
Krogmann and Woyczechowski, 2000; Agada and Sponza, 2004; cWarith and Takata, 2004; Stessel and
Murphy, 1992; eBorglin et al., 2004
Table 2-6. Comparison of leachate paramel
ers with other anaerobic landfi
1 studies
Parameters
(mg/L
except for
pH)
Conventional
landfill Ia
Conventional
landfill 2b
Bioreactor
landfills'
Bioreactor
landfill lab
scaled
This study
COD
140-152000
1000-40000
20-17000
100-88000
2000-80000
BODj
20-57000
50-25000
0-10000
6600-60000
TOC
30-29000
7000-19000
Ammonia
50-2200
50-1500
76-1850
100-1600
TDS
2000-60000
2000-25000
18000-50000
PH
4.5-9
3-7.5
5.4 -8.6
4-7.5
4.5-7.5
Ca
10-7200
300-4000
20-4000
aKjeldsen et a
., 2002; bPokhrel, 2004; cReinhart and Al-Yousfi, 1996;
dPohland and Kim, 1999


222
El-Fadel, M, Findikakis, A. N. and Leckie, J. O., 1989, A numerical-model for methane
production in managed sanitary landfills, Waste Management & Research, 7(1), 31-
42.
Fang, M., J. Wong, W. C, Ma, K. K. and Wong, M. H., 1999, Co-composting of sewage
sludge and coal fly ash: nutrient transformations, Bioresource Technology, 67(1),
19-24.
Fannin, K., 1987, Anaerobic digestion of biomass, edited by Chynoweth, P. D., and
Roinsaacson, Elsevier applied science, New York, USA
Faure, G., 1991, Principles and applications of inorganic geochemistry : a comprehensive
textbook for geology students, McMillan Pub. Co., New York
Fleming, I. R., Rowe, R. K. and Cullimore, D. R., 1999, Field observations of clogging in
a landfill leachate collection system, Canadian Geotechnical Journal, 36(4), 685-
707.
Florida Department of Environmental Protection (FDEP), 2002, Solid Waste Mangement
Annual Report, Tallahassee, FL, USA.
Fox, M. and Noike, T., 2004, Wet oxidation pretreatment for the increase in anaerobic
biodegradability of newspaper waste, Bioresource Technology, 91(3): 273-281.
Gerritse, R., Vriesema, G., R., Dalenberg, J. W. and Deroos, FI. P., 1982, Effect of
sewage-sludge on trace-element mobility in soils, Journal of Environmental Quality,
11(3), 359-363.
Gibson, R. E. and Lo, K. Y., 1961, A theory of consolidation for soils exhibiting
secondary compression, ACTA Polytechnic Scandianavica, 10, 296
Grima, S., Bellon-Maurel, V., Feuilloley, P. and Silvestre, F., 2000, Aerobic
biodegradation of polymers in solid-state conditions: A review of environmental
and physicochemical parameter settings in laboratory simulations, Journal of
Polymers and the Environment, 8(4), 183-195.
Gujer, W. and Jenkins, D., 1975, Nitrification Model for Contact Stabilization Activated-
Sludge Process, Water Research, 9(5-6), 561-566.
Gunaseelan, V. N., 1997, Anaerobic digestion of biomass for methane production: a
review, Biomass & Bioenergy, 13(1-2), 83-114.
Hoar, T. P. and Rothwell, G. P., 1970, Potential/Ph Diagram for a Copper-Water-
Ammonia System Its Significance in Stress-Corrosion Cracking of Brass in
Ammoniacal Solutions, Electrochimica Acta 15(6), 1037.


198
Table C-5
continued)
date
lys 1
lys 2
date
lys 3
lys 4
3/12/2005
14576
10648
3/20/2005
15017
9954
3/24/2005
9636
3/26/2005
15017
9203
4/3/2005
13539
4/17/2005
13484
9526
4/24/2005
12010
9368
4/30/2005
9303
4/30/2005
13470
5/7/2005
13302
8205
5/16/2005
12293
8637
5/23/2005
12580
8490
6/6/2005
12990
6/14/2005
12505
6/28/2005
12930
3231
7/5/2005
13160
6493
7/11/2005
11760
6775
7/19/2005
7990
7/27/2005
12230
5625
8/10/2005
7685


69
Table 3-1.Heavy metal sources in fabricated waste Stream-
Waste components
Contained heavy metals
% of component in fabricated
waste
CCA treated wood
Copper, Chromium and
Arsenic
1%
Cathode-ray Tube (CRT) glass
Lead
1%
Aluminum sheet
Aluminum
4%
Galvanized steel sheet
Zinc, Manganese and Iron
4%
Table 3-2. Results of statistical analysis of metal leached between aerobic and anaerobic
Average concentrations (mg/L)
F
P-value
F-crit
Aerobic
Anaerobic
A1
7.89
1.28
206.89
8.3E-33
3.89
As
0.40
1.28
66.15
4. IE-14
3.89
Cr
0.19
0.10
40.67
1.19E-09
3.89
Cu
2.87
0.02
81.10
1.58E-16
3.89
Fe
35.06
167.68
53.09
6.92E-12
3.89
Mn
1.91
4.57
35.80
9.75E-09
3.89
Pb
0.22
0.03
31.32
7.07E-08
3.89
Zn
54.36
201.12
66.30
3.87E-14
3.89
Table 3-3. The amount of leachate produced and used for analysis
lys 1
lys 2
lys 3
lys 4
leachate produced (mL)
8,717
9,747
18,571
15,979
leachate released (mL)
(used for analysis)
4,024
4,081
6,135
6,111


50
hydrogen peroxide and then analyzed for heavy metals using ICP-AES following US
EPA, SW-846 Method 601 OB (USEPA, 2003). Digested samples were filtered using ash
free cellulose filters and analyzed for heavy metals and cations using Inductively Coupled
Plasma-Atomic Emission Spectrometry (ICP-AES) (Thermo Electronics, USA). Leachate
samples preserved with concentrated nitric acid were analyzed for a total of 8 metals (As,
Cu, Cr, Mn, Zn, Pb, Fe and Al).
3.3 Results and Discussions
3.3.1 Changes in Metal Concentrations versus Time and the Percentage of Mass
Loss
The following section presents the results (for each metal) of the aerobic and
anaerobic lysimeters as separated plots. The experimental time period for the aerobic and
anaerobic lysimeters differed. Because of their time scale difference, the cumulatitive
mass of metal leaching was plotted for all lysimeters as a function of waste mass loss.
The estimation of mass loss is described in appendix A. The total amounts of leachate
produced and used for the analysis are summarized in Table 3-3.
3.3.1.1 Aluminum
The changes in Al concentration in aerobic and anaerobic conditions over a period
of time are depicted in Figure 3-1. High Al concentrations were observed from both
aerobic and anaerobic lysimeters (18 and 20mg/L) for the first 10 to 20 days. Whereas the
Al concentrations of anaerobic lysimeters dramatically decreased to below 0.5mg/L
within 100 days, great changes in Al concentrations were not observed from the aerobic
lysimeters. The changes in Al concentrations in aerobic lysimeters were mainly
controlled by pH; high concentrations of Al were observed from both lysimeters 1 and 2
at pH < 6 and pH > 8, and lowest Al concentrations (< 1 mg/L) were observed at 6 < pH


3
o To compare leachate and gas quality between aerobic and anaerobic bioreactor
landfills,
o To explore the fate of heavy metals leached from the fabricated wastes in
aerobic and anaerobic bioreactor landfills,
o To explore the decomposition of lignocellulosic wastes in anaerobic and
aerobic bioreactor landfills, and
o To evaluate the loss of mass versus the loss of volume in aerobic and anaerobic
bioreactors for use in future settlement model development.
1.3 Research Approach
Four stainless steel lysimeters were constructed: two were operated aerobically and
two were operated anaerobically. These lysimeters were designed and constructed as part
of a previous research experiment (Sheridan, 2003). After operating the aerobic and
anaerobic lysimeters for 1 and 2 years, respectively, one aerobic and one anaerobic
lysimeter were dismantled. Waste samples were collected and characterized. The
remaining aerobic and anaerobic lysimeters were kept in operation so that waste
stabilization could be completely researched; the results of this extended operation will
be presented elsewhere.
To compare leachate and gas quality between the aerobic and anaerobic bioreactors,
two pairs of simulated landfill lysimeters containing fabricated wastes were operated as
aerobic and anaerobic bioreactors. The fabricated wastes were loaded into the lysimeters,
compacted, and mixed with water and seed (either anaerobic sludge or aerobic compost).
Leachate generated by the lysimeters was collected and analyzed for leachate quality
parameters. A mixture of collected leachate and deionized water was added back to the


129
5.4.3 Application
The development of the landfill settlement model using the biological reactions
may require many case studies since biological waste decomposition is affected by many
parameters. Through the many research efforts, the first-order kinetics model was
reported as one of the most suitable to describe the biological decomposition (El-Fadel et
al., 1989). However, it may be difficult to express all landfill settlement using the aspect
of waste decomposition. Once waste is disposed in landfills, initial compression and
primary settlement may take place by its own elasticity and mechanical interactions
caused by overburden pressure. After a period of time, the waste component, which can
be readily decomposed, may be depleted, and mass loss of waste is mainly controlled by
hydrolysis of waste (Park and Lee, 1997). The settlement models using this hydrolysis of
waste are called a bioconsolidation model (El-Fadel and Khoury, 2000). Park and Lee
(1997) proposed the bioconsolidation model using the hydrolysis coefficient, k:
E (t) = E tot-dec (1 6 ) (5)
where, e tot-dec= total amount of compression due to the waste decomposition; k =
hydrolysis coefficient (or decay rate); and t = time (year)
Based upon the data obtained from this research, mass loss also can be expressed
by the equation (6):
Mass loss, % = ^ x 100 = 100(1 e~h) (6)
M o
If this equation (6) is substituted into equation (4), a mass loss-settlement
relationship, a similar equation can be derived to the bioconsolidation model.
= H, (0.7244 0.169 log(l e*))
(7)


Sulfate (mg/L) Sulfate (mg/L)
174
Figure C-6. The change in sulfate of the aerobic and anaerobic lysimeters over time


Mn, mg/L Mn, mg/L
78
o.oi
50
100
150
200
250
300
350
Figure 3-7. Changes of Mn concentrations over time


202
Table C-7 (continuec
D
Date
Lys 1
Lys 2
Date
Lys 3
Lys 4
612
886.4613
1444.554
619
879.276
1310.144
626
993.4448
1468.26
633
872.1489
1409.711
639
837.3708
1257.9
648
669.5123
977.0133
655
702.8366
1046.08
662
771.0181
883.8794
670
654.4303
718.6858
676
799.8811
943.3915
684
716.8409
730.0174
698
699.709
886.2881
705
926.4855
880.748
711
970.8101
857.0626
719
1053.542
771.5203
727
1061.779
780.5869
741
1029.21
817.9315


122
daily difference of settlement was recorded by measuring the movement of the length of
the shaft visible above the flange after adjusting the pressure to 510 psi (3516 kPa).
5.2.3 Compression Index and Phase Separate Method
The measured settlement data were fit to several different relationships that had
been previously proposed to model landfill settlement in the secondary phase (without
consideration of condition with decomposition rate). These relationships included the
modified secondary index proposed by Sowers (1973). Bjangard and Edgers (1990)
developed the modified secondary index and explained landfill settlement mechanisms by
separating the landfill settlement curve by different phases (phase separate method).
Solving for parameter described on part of the relationship allowed comparison with
other studies.
A modified secondary index (Ca) was used to describe settlement behavior of the
aerobic and anaerobic lysimeters. Originally, the secondary compression index (Ca) was
used to describe the secondary settlement for soil tests, but Sowers (1973) first applied
this concept to landfill settlement. Since it is hard to estimate void volume in the field, the
secondary compression index was modified. The modified compression index (Ca) is
used to estimate the settlement that occurred after the first mechanical settlement. The Ca
and Ca can be expressed as follows:
C =
log(t2/t,)
(1)
C. =
AH
G
H-\og(t2/ti) 1 + e0
(2)


200
Table C-6
continued)
1/21/2005
52350
37850
2/5/2005
48550
38650
2/8/2005
49600
36900
2/24/2005
51200
40150
3/12/2005
51500
36500
3/20/2005
52050
37000
3/26/2005
64850
33100
4/3/2005
55650
32550
4/17/2005
48650
32150
4/26/2005
40850
28650
4/30/2005
43000
29950
5/7/2005
45050
28650
5/16/2005
44250
27450
5/23/2005
45000
26850
5/31/2005
39250
22650
6/6/2005
41750
24400
6/14/2005
43800
23400
6/28/2005
42050
21400


7
leachate and landfill gas that result from aerobic and anaerobic operation of identical
MSW streams. The experiments conducted involved a technique long employed in the
study of landfills: waste-filled columns constructed and operated to simulate landfill
processes, referred to here as lysimeters (Pohland, 1980). The columns were designed
and operated to control several parameters not traditionally simulated in such
experiments, such as temperature and overburden pressure. The objective was to compare
leachate and landfill gas quality between each type of system so that similarities and
differences can be better understood and to assist in future decision-making, design and
operation efforts. Several complementary objectives were evaluated as part of this
experiment and they are described in greater detail in Chapters 3 (fate of metals), 4
(comparison of decomposition) and 5 (comparison of settlement).
2.2 Material and Methods
Four lysimeters were used in this research, and each consisted of a stainless steel
column and a carriage system component. The original design and construction of the
lysimeters used for this research were described previously by Sheridan (2003). Two
were operated aerobically (lysimeter 1 and 2) and two were operated anaerobically
(lysimeter 3 and 4). Three parameters, temperature, air addition, and overburden pressure,
were controlled in an effort to simulate actual aerobic or anaerobic bioreactor landfills.
2.2.1 General Description of the Lysimeter
A schematic of each lysimeter type is presented in Figure 2-1 (see Figure B-l for
additional detail). The 6-ft stainless steel main body contained 5 front ports, 2 back ports
and 1 valve at the bottom for leachate collection. The front ports were used for air
addition (in the case of the aerobic lysimeters). The carriage system component was
designed to support a hydraulic pressurizing unit installed at the top of each lysimeters


Metal concentrations (mg/kg)
88
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Figure 3-14. (continued)


100
Average TS (%) = ^ ([dry mass fraction,%]/ x [TS%]i) (4)
1=1
Based on the methane yield of the new and decomposed waste, the biodegradable
fraction contained in volatile solids was determined as:
, ,,, . .... a mass of waste (g) converted into biogas ...
Biodegradable fraction (%) = ><100
Initial dry mass (g)
Biodegradable volatile solids (BVS, %) = (VS %) x (Biodegradable fraction %)
Details about all calculations are described in appendix A.
4.2.4 Cellulose and Lignin Determination
In order to determine the cellulose content of selected solid waste samples (SYP),
lg of dried sample was placed in a 250 mL Erlenmeyer flask containing 6 mL of
deionized water, 24 mL of glacial acetic acid and 2mL of concentrated nitric acid. After
boiling for 20 minutes, 50 mL of toluene was added and swirled for 2 minutes. Waiting
until solids were settled, toluene was decanted through the Gooch filter apparatus. This
same procedure was repeated using 50 mL of ethyl ether. All solid materials were then
transferred from the flask to a Gooch crucible with glass-fiber filters using acetone.
Solids contained in the crucibles were washed by pouring 75 mL of hot toluene, hot
methanol and ether through the crucibles successively. Washed solid samples were dried
at 103C for 1 day and their dry weights were recorded. After removal of the organic
portion at 550C for 1 hour, weights of the residual materials were recorded again.
Cellulose content was then determined by multiplying 100 by the ratio of the organic
portion ignited to the initial weight of the sample.
To determine the lignin content, 1 g of sample was placed in a Gooch crucible with
a glass-fiber filter and washed with 150 mL of toluene and ethanol mixed in the 2:1 ratio.


80
60
40
20
0
100
80
60
40
20
0
eC
187
Days
. Change in methane yields of the waste layer 4-1 and 4-2


221
Charlatchka, R. and Cambier, P., 2000, Influence of reducing conditions on solubility of
trace metals in contaminated soils. Water Air and Soil Pollution 118(1-2), 143-167.
Chynoweth, D. P. and Isaacson, R., 1987, Anaerobic digestion of biomass. London ; New
York, Elsevier Applied Science.
Chynoweth, D. P., Jerger, D. E. and Srivastava, V. J., 1985, Biological gasification of
woody biomass, Proceedings of the 20th intersociety Energy Conversion
Engineering Conference, Warrendale, PA, Society of Automotive Engineers, Inc. 1,
573-579.
Coward, H. F. and Jones, G. W., 1952, Limits of flammability of gases and vapors.
Bulletin 503, U.S. Bureau of Mines.
Critchley, M. M., Pasetto, R. and O'Halloran, R. J., 2004, Microbiological influences in
'blue water' copper corrosion, Journal of Applied Microbiology 97(3), 590-597.
Cummings, S. P. and Stewart, C. S., 1994, Newspaper as a Substrate for Cellulolytic
Landfill Bacteria, Journal of Applied Bacteriology, 76(2), 196-202.
Das, B. M., 2002, Principles of geotechnical engineering, Pacific Grove, CA, Brooks
Cole/Thompson Learning.
De Baere, L. and Verstrate, W., 1984, High rate anaerobic composting with biogas
recovery, Biocycle, 25, 30-31.
De Boer, I. J. M., 2003, Environmental impact assessment of conventional and organic
milk production. Livestock Production Science 80(1-2), 69-77.
De. Baere, L., 1984, High rate dry anaerobic composting process for the organic fraction
of solid waste, 7th Symposium on Biotechnology for Fuel and Chemicals,
Gatlinburg, Tennessee, 1984.
Drever, J. I., 1988, The geochemistry of natural waters. Englewood Cliffs, N.J., Prentice
Hall.
Dubey, B. and Townsend, T., 2004, Arsenic and lead leaching from the waste derived
fertilizer ironite. Environmental Science & Technology 38(20), 5400-5404.
Eary, L. E., 1999, Geochemical and equilibrium trends in mine pit lakes, Applied
Geochemistry, 14(8), 963-987.
Edwards, M., Jacobs, S. and Taylor, R. J., 2000, The blue water phenomenon, Journal
American Water Works Association, 92(7), 72-82.
El-Fadel, M. and Khoury, R., 2000, Modeling settlement in MSW landfills: a critical
review, Critical Reviews in Environmental Science and Technology, 30(3), 327-
361.


132
Table 5-1. (Ca)min and (Ca)max values of lys 1 through 4
(cy
min
(C
1 max
Todays
t2
AH, %
(Ca) mjn
t.
t2
AH, %
(Ca) max
lys 1
23
70
3.05
0.063
70
316
11.45
0.175
lys 2
33
60
1.88
0.072
60
360
17.64
0.227
lys 3
46
260
3.76
0.050
260
761
6.52
0.140
lys 4
29
440
3.75
0.032
440
718
11.41
0.536
Table 5-2. k values of aerobic and anaerobic lysimeters
Lys 1
Lys 2
Lys 3
Lys 4
Overall (yr1)
0.379
0.377
0.0202
0.1048
delayed phase
(lys 3: 0 ~ 650 days; lys 4: 0~ 400 days)


0.0145
0.0219
Loe phase
0.379
0.377
0.1944
0.2458


49
local hardware store and cut into 1.5 cm 1.5 cm square. Galvanized steel served as a
source of both Fe and Zn. CCA-treated wood was used as a source of Cr, Cu and As.
Crushed cathode ray tube (CRT) monitor glass was used as a Pb source. Total Cu, Cr,
and As concentrations were 2350 50, 2890 56, and 1330 10 mg/kg, respectively.
Crushed CRT monitor glass used in this research was a mixture of the funnel sections of
30 CRT color monitors. Jang and Townsend (2003) reported that 413 mg/L of Pb leached
from CRT funnel glass using the toxicity characteristics leaching procedure (TCLP).
3.2.2 Sampling Methods
Leachate samples were collected weekly via a sampling port located at the bottom
of each lysimeter. A portion of the leachate collected was used for analysis of general
water quality parameters and the remainder was injected back into the lysimeters. A 50
mL aliquot was preserved with concentrated nitric acid and used for heavy metal
analysis.
Lysimeter studies were conducted for 379 and 741 days for aerobic and anaerobic
lysimeters, respectively. After the lysimeter studies were completed, solid wastes were
removed from single aerobic and anaerobic lysimeter and analyzed for heavy metals. The
samples were divided by depth into 4 fractions. Details about these fractions are
summarized in Table A-2 in appendix A. Each fraction was then separated into 5
categories which include office paper, cardboard, newspaper, wood blocks and plastics.
The separated samples were ground using an Urschell mill (Fritsch, German).
3.2.3 Analytical Methods
Leachate samples were digested with nitric and hydrochloric acids following EPA
method 3050B and 3010A for solid and liquid digestion, respectively (USEPA, 2003).
Approximately 2 g of the ground samples were digested using nitric acid and 30%


40
30
20
10
0
40
30
20
10
0
177
140
Lys 1
l
I:
I
1
I ;
| r
ll
ll
11
ll
is
II
II
I I
! 1 1
I I
n
I I
l I
I I
I
l
L-
i
i
i i
'V
11
I
/
-O-J
I /
1/
- 120
- 100
- 80
60
- 40
20
100
200
300
400
140
120
- 100
- 80
- 60
40
- 20
Lys 2
"I r-
ll
H
H
|!
ll
n
I I
I I
I I
I
I
i r.
trj
u
a Lrv~V""/
-^Ij'l
l|
I j l /
U i\! V
i
100
i
300
200
Days
400
?. The change in biogas produced from the aerobic lysimeters
Air injection rate (mL/min) Air injection rate (mL/min)


Ammonia (mg/L) Ammonia (mg/L)
38
Figure 2-8. Changes in ammonia concentrations versus time


The fate of metals leached from the various metal sources including cathode ray
tube (CRT) monitor glass and ground CCA-treated wood were explored. Metal leaching
trends observed varied from anaerobic to aerobic lysimeters; the average concentrations
of As, Fe, Mn, and Zn in the anaerobic lysimeters proved significantly greater in
concentration than observed in the anaerobic lysimeters. Furthermore, significantly
greater concentrations of Al, Cu, Cr, and Pb were detected in the aerobic lysimeters as
compared to the anaerobic lysimeters.
Using leachate and gas measurements, mass losses from the aerobic and anaerobic
lysimeters were estimated. Mass removed from the wastes was primarily converted into
gas; after the water was removed from the lysimeters, the mass of waste excavated from
each lysimeter was compared with the estimated loss mass. For wood waste, no great
influence on air addition was observed through cellulose/lignin analysis. Methane
potential of lignoceliulosic materials other than wood waste resulted in great differences
of biodegradation between aerobic and anaerobic lysimeters.
The landfill settlement behavior occurring in aerobic and anaerobic simulated
landfills was mathematically analyzed. The logarithm of mass loss was linearly correlated
with the percentage of settlement. With this relationship, the secondary settlement of
bioreactor landfills could be mathematically modeled using the first-order exponential
function.
xv


Metal concentrations (mg/L)
183
Figure C-15. Pb concentration versus pH in leachate from the lysimeters


60
50
40
30
20
10
0
60
50 -
40
30
20
10
0
re 2
>>
39
AEROBIC
OQ
100
200
300
400
i
100
I
200
i
600
300
400
Days
500
700
800
h Changes in TDS of the aerobic and anaerobic lysimeters versus time


112
Table 4-6. Summary of cellulose and lignin content of the wood samples
Sample ID
Cellulose
Lignin
Measurement
Average
Measurement
Average
raw
53.2%
50.47%
29.2%
28.80%
47.7%
28.4%
2-1
44.4%
42.34%
31.3%
30.46%
40.3%
29.6%
2-2
46.3%
46.09%
30.9%
30.88%
45.9%
30.8%
2-3
49.0%
47.26%
32.1%
31.80%
45.5%
31.5%
2-4
48.7%
48.94%
29.3%
30.40%
49.1%
31.5%
4-1
47.4%
46.75%
27.6%
28.62%
46.2%
29.7%
4-2
45.4%
45.38%
28.5%
27.96%
45.4%
27.4%
4-3
50.8%
50.22%
27.4%
28.27%
49.6%
29.2%
4-4
49.1%
49.87%
27.9%
28.26%
50.7%
28.6%


134
Figure 5-1. The changes in settlement, cumulative gas (CO2) and pH over time


89
OXIDIZED REDUCED
Methylation
Q : precipitated
o : precipitated in the presence of sulfur
Figure 3-15. Fate of heavy metals thermodynamically occurred in aerobic (oxidizing) and
anaerobic (reducing) conditions (Bridle, 2004; Drever, 1988; Sadiq, 1997;
Masscheleyn et al, 1991; Richard and Bourg, 1991; Benjamin, 2000; McBride
and Blasiak, 1979)


106
observed in comparison to the raw (50.5 %) and the anaerobic lysimeter blocks. A
comparison of SYP block biodegradability by methane yields was determined more
valuable.
In contrast to the cellulose and lignin results of the wood samples, the methane
yields of SYP excavated from the aerobic lysimeters were statistically lower than those of
raw and the anaerobic lysimeters (p < 0.05) (Figure 4-5). Average BMP assay results of
wood from the anaerobic lysimeters were statistically the same as the raw SYP.
There was no evidence to find if the lignin component was degraded in both the
aerobic and anaerobic lysimeters. Overall, the lignin concentrations of raw SYP blocks
were not significantly different from those of the both lysimeters.
4.4 Discussion
Through this research, waste decomposition was evaluated by conducting the BMP
assay on solid waste samples. The BMP assay results indicated that the methane yield of
the aerobic lysimeter was lower than that of the anaerobic lysimeter despite different test
periods (1 year and 2 years for the aerobic and anaerobic lysimeters, respectively). The
largest difference of methane yield between the raw and decomposed waste was observed
with office paper. These differences decreased following the order: office paper >
cardboard > newspaper > wood. This result indicated that waste may be decomposed in a
landfill in the same order. It was confirmed by the changes in the percentage of organic
fraction as waste decomposed (Figure 4-3).
The methane yield results of the newspaper and SYP blocks indicated that
lignocellulosic materials with high lignin contents can be more decomposable in aerobic
condition relative to anaerobic condition. It is reported that the volume reduction of
general lignocellulosic materials was not much different between aerobic and anaerobic


60
50
40
30
20
10
0
7.0
6.5
6.0
46
Days
-16. Changes in gas concentrations, pH and gas generation rate after air injection
into lysimeter 3


N-NH3+(mg/L)
171
Figure C-3. The change in ammonia of the aerobic and anaerobic lysimeters over time


43
Days
Figure 2-13. The changes in air injection rate and gas concentrations of aerobic lysimeter
Air injection rate (mL/min)


As, mg/L As, mg/L
73
1.0
0.8 -
0.6 -
0.4 -
0.2 -
AEROBIC
Lys 1
O Lys 2
M
O O o
Oo
tPo
o o



o
O


0.0
p
- Qd O CTO f -
Qo
100
200
300
400
ANAEROBIC
A Lys 3
A Lys 4
A

A
A
A M^*A
A A A A
ViA
AA a4£^-2AAa
i 1
400 600 800
Days
Figure 3-2. Changes of As concentrations over time


Cumulative Biogas (CH4 and C02), L
45
Figure 2-15. Cumulative biogas vs. days in aerobic and anaerobic lysimeters
800


53
effect of pH on the total oxyaionic arsenate concentrations by pH. This changes in
solubility results in the relatively higher concentrations of As at alkaline conditions
observed in the aerobic lysimeters (Figure 3-2).
3.3.1.3 Chromium
The initial Cr concentrations of the anaerobic lysimeters were higher than those of
the aerobic lysimeters (Figure 3-3). However, Cr concentrations in the anaerobic
lysimeters gradually decreased to below 0.05mg/L by the day 453. As the pH of lysimeter
4 changed to moderately alkali (pH > 7.4) after day 464, minor increases in Cr
concentrations were observed. In contrast, clear Cr leaching trends were not exhibited by
the aerobic lysimeters before day 100, but an increase in Cr concentration did occur
following day 150. This increase in Cr concentration corresponds to an increase in pH.
Overall, the average Cr concentrations of the aerobic lysimeters were significantly greater
than those of the anaerobic lysimeters. This may be because thermodynamically Cr can
be present as an ionic form at alkaline pH under oxidized condition.
The toxicity of Cr is determined by its oxidation state. Among the various Cr
oxidation states, only trivalent and hexavalent forms are taken into consideration in
natural aquatic systems. Hexavalent Cr is considered more toxic than trivalent Cr due to
its high mobility and solubility. Cr (VI) may be reduced to Cr (III) at low ORP potential.
Cr (VI) becomes unstable and is reduced to Cr (III) at low pH. In order to maintain the
oxidation state of Cr as Cr (VI) at a low pH, it is necessary to keep highly oxidizing
conditions (Richard and Bourg, 1991). In contrast to other metals such as As and Cu, Cr
(III) is not likely to precipitate with sulfide. Chromium solubility is mainly controlled by
Cr(OH)3(s). Generally, Cr(OH)3(s) is formed in a pH range of 6.5 to 7 under moderately
oxidizing or reducing conditions.


CHAPTER 1
INTRODUCTION
1.1 Problem Statement
Landfills remain the predominant method for managing municipal solid waste
(MSW) in the U.S. Although modem engineered landfills protect the environment from
groundwater contamination and in some cases gas emissions, they are most often
operated in a fashion where only a small amount of the disposed waste is permitted to
biodegrade to a more stabilized state. This results in large amounts of undegraded waste
being stored for many years in the future; their management will continue to demand
resources and may pose a long-term environmental risk.
Alternatively, many innovative and more environmental-friendly strategies for
operation of MSW landfills have been proposed (Stegmann, 1983; Barlaz et al., 1992;
Komilis et al., 1999). Among these techniques, leachate recirculation has been found to
be the most practical approach for enhancing waste decomposition and stabilization in
landfills (Reinhart et al., 2002). This process stabilizes landfilled waste more rapidly
because of the increased moisture content and the more effective distribution of nutrients
and microorganisms in the landfill. This result creates a very favorable environment for
the existing anaerobic organisms responsible for waste degradation. If controlled,
methane produced can be utilized as a resource. This technique has changed the concept
of a landfill from a historical garbage dump to a bioreactor, where various biochemical
reactions are managed in a controlled fashion.
1


Fluoride (mg/L) Fluoride (mg/L)
172
Figure C-4. The change in fluoride of the aerobic and anaerobic lysimeters over time.


COMPARATIVE STUDIES OF AEROBIC AND ANAEROBIC LANDFILLS
USING SIMULATED LANDFILL LYSIMETERS
By
HWIDONG KIM
A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY
UNIVERSITY OF FLORIDA
2005


APPENDIX B
SUPPLEMENTAL S
1. Peristaltic pump
2. air purging
3. leachate collection port
4. Wet-tip gas totalizer
5. Temperature controller
6. Gas-sampling bag
7. Hydraulic cylinder
8. Heating tape
9. Hydraulic jack
10. Plunger
11. Fabricated wastes
AEROBIC ANAEROBIC
LYSIMETER LYSIMETER
Figure B-l. Schematics of aerobic and anaerobic lysimeters used for this research
159


52
concentrations in the aerobic lysimeters decreased initially and increased with increasing
pH. The As leaching pattern of the aerobic lysimeters appears similar to A1 leachate
trends. The lowest As concentration of lysimeter 1 was 0.12 mg/L on the day 163. For
lysimeter 2, extremely low As concentrations were observed, with several samples below
the detection limit (0.011 mg/L) despite a pH < 6. Overall, As dissolved in the leachate of
the aerobic lysimeters was significantly lower than that of the anaerobic lysimeters (p <
0.05).
Figure 3-9 depicts the distribution of As concentrations observed from the aerobic
and anaerobic lysimeters at various pH conditions. It has been reported that As solubility
changes with pH and is characterized by a U-shaped curve in oxidizing conditions
(Drever, 1988). However, Carbonell-Barrachina et al (1999) reported that in the presence
of sulfide, Fe, and Mn, the solubility of As was dramatically lower and did not follow a
U-shaped solubility curve. However, As concentrations were not impacted by these
constituents in the anaerobic lysimeters. The most likely explantion is that the anaerobic
lysimeters had poor-anoxic conditions during the first phase. The low sulfide
concentrations at a pH < 6 confirm that the redox potential was not low enough for
sulfide to become involved in As precipitation. For the aerobic lysimeters, As
concentrations were low under acidic conditions and increased up to 1 mg/L at a pH of 9.
Masscheleyn et al (1991) found that As solubility decreased substantially as the
redox potential increased. The changes in As solubility are also associated with the
oxidation state of iron; Fe (III) has a strong affinity for arsenate. Therefore, low arsenic
concentrations are likely dictated by the low solubility of arsenate. Under oxidizing
conditions, As solubility may increase or decrease by pH changes. This is because of the


Metal concentration (mg/L)
94
3
1
0.01 -
0.001 -
0.0001
10
0.01 -
0.001 n 1 1 1 1 1 1
Jambeck Jambeck Aerobic Aerobic Anaerobic Anaerobic
(acid) (methane) (acid) (alkali) (acid) (methane)
Figure 3-18. Comparison of As, Cu and Cr leaching trend of the lysimeters to other study
(Jambeck, 2004)


54
According to the potential-pH diagram of Cr (Figure 3-10), total Cr obtained from
both aerobic and anaerobic lysimeters in an acidic environments is likely to be Cr (III) as
Cr(OH)2+. In contrast, dissolved Cr from the aerobic lysimeter at a pH 9 could be Cr (VI)
as CrC>42\ Since all Cr species presented on the potential-pH diagram are based upon
assuming thermodynamic equilibrium, all Cr obtained from the aerobic lysimeters at high
pH may not be Cr (VI).
It is noted that an increase in Cr was observed from lysimeter 4 around a neutral pH
(A in Figure 3-3). The most likely explanation for this is the lower Fe concentrations of
lysimeter 4 than of those of lysimeter 3 (Figure 3-6). In the presence of Fe, Cr may be
precipitated with Fe rather than OH' due to rapid kinetics. The complexation of Fe and Cr
decreases Cr solubility lower than the complexation of OH and Cr (Eary and Rai, 1987).
Therefore, an increase of Cr concentration at the end of lysimeter 4 would be the result of
a decrease of Fe concentrations.
3.3.1.4 Copper
Overall copper concentrations of the aerobic lysimeters were one to three orders of
magnitude higher than those of the anaerobic lysimeters (Figure 3-4). For the aerobic
lysimeters, clear Cu leaching patterns over time were not observed, but relatively large
changes in Cu concentrations were observed at lysimeter 1 from the day 140 to 190. This
period of time corresponded to a pH change from 5.5 to 9. For anaerobic lysimeters, Cu
concentrations gradually decreased for the first 450 days. The initial Cu concentrations of
lysimeter 3 and 4 were 0.082 and 0.234 mg/L, respectively. Although the concentrations
slightly increased after day 450, final concentrations of Cu remained lower than the initial
values.


Pb, mg/L Pb> m§/L
76
10
1 -
0.1 -
6
0.01 -
AEROBIC
Lys 1
O Lys 2
\

O
0
o*o
O
o o
o
0*0 ,
o
0 o
0.001
50
100
150
200
250
i
300
350
A Lys 3
A Lys 4
ANAEROBIC
0.1 -
ASAVa
a^aaa^
a Aa A
* aA
AaAAA aa
A .A
aa a
0.01 -
"VA
AA -A ^
A \
Aa a a
A
Below detection limit of ICP for Pb
0.001
~r
100
200
300
400
Days
500
T
600
700
Figure 3-5. Changes of Pb concentrations over time


220
Basso, M. C., Cerrella, E. G. and Cukierman, A. L., 2002, Lignocellulosic materials as
potential biosorbents of trace toxic metals from wastewater. Industrial &
Engineering Chemistry Research, 41(15), 3580-3585.
Benjamin, M. M., 2002, Water chemistry. Boston, McGraw-Hill.
Berge, N. D., Reinhart, D. R. and Townsend, T. G., 2005, The fate of nitrogen in
bioreactor landfills. Critical Reviews in Environmental Science and Technology
35(4), 365-399.
Bissen, M. and Frimmel, F. H., 2003, Arsenic a review. part 1: occurrence, toxicity,
speciation, mobility. Acta Hydrochimica Et Hydrobiologica 31(1), 9-18.
Bjamgard, A. B. and Edgers, L., 1990, Settlement of municipal solid waste landfills,
Proceedings of the 13th Annual Madison Waste Conference, University of
Wisconsin, Madison, WI, 192-205.
Bleiker, D. E., Farquhar, G. and Mcbean, E., 1995, Landfill Settlement and the Impact on
Site Capacity and Refuse Hydraulic Conductivity, Waste Management & Research,
13(6), 533-554.
Borglin, S. E., Hazen, T. C., Oldenburg, C. M. and Zawislanski, P. T., 2004, Comparison
of aerobic and anaerobic biotreatment of municipal solid waste, Journal of the Air
& Waste Management Association, 54(7), 815-822.
Bozkurt, S., Moreno, L. and Neretnieks, I., 1999, Long-term fate of organics in waste
deposits and its effect on metal release, The Science of the Total Environment, 228,
135-152.
Bradl, H. B., 2005, Heavy metals in the environment: [origin, interaction and
remediation]. Amsterdam ; Boston, Elsevier Academic Press.
Brown, A., 1985, Review of lignin in biomass. Journal of Applied biochemistry 7, 371-
387.
Calli, B., Mertoglu, B., Inane, B. and Yenigun, O., 2005, Community changes during
start-up in methanogenic bioreactors exposed to increasing levels of ammonia,
Environmental Technology, 26(1), 85-91.
Carbonell-Barrachina, J., DeLaune, A., Patrick, R. D., Burlo, W. H., Sirisukhodom, F.,
and Anurakpongsatom, P., 1999, The influence of redox chemistry and pH on
chemically active forms of arsenic in sewage sludge-amended soil. Environment
International 25(5), 613-618.
Chandler, J. A., Jewell, W. J., Gossett, J. M., Vansoest, P. J. and Robertson, J. B., 1980,
Predicting Methane Fermentation Biodegradability, Biotechnology and
Bioengineering 22, 93-107.


Ill
Table 4-5. Overall met
iane yields of waste la
yers of the lysimeters 2 and 4.
Waste layer
L/g-VS
vs
L/g
2-1
0.148
83.4%
0.1236
2-2
0.092
83.3%
0.0763
2-3
0.093
81.4%
0.0760
2-4
0.133
85.7%
0.1136
4-1
0.167
84.0%
0.1399
4-2
0.156
87.9%
0.1373
4-3
0.163
86.5%
0.1414
4-4
0.225
86.5%
0.1943
Raw waste
0.337
88.4%
0.2974


Al concentrations, mg/L
179
4 5 6 7 8 9 10
pH
Figure C-l 1. Al concentration versus pH in leachate from the lysimeters.


COMPARATIVE STUDIES OF AEROBIC AND ANAEROBIC LANDFILLS
USING SIMULATED LANDFILL LYSIMETERS
By
HWIDONG KIM
A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY
UNIVERSITY OF FLORIDA
2005

Copyright 2005
by
Hwidong Kim

This document is dedicated to my parents and loving wife

ACKNOWLEDGMENTS
I would like to thank my advisor, Dr. Timothy G. Townsend, for showing such
great patience as a mentor. He gave me this great opportunity to study on the field of
solid waste. He also showed me the way of living as an engineer, professor, and a family
man. I cannot forget his tears when Mr. Townsend passed away. I would also like to
thank my committee members, Dr. Angela Lindner, Dr. Frank Townsend and Dr. Roger
Nordstedt, and my other spectacular faculty members, Dr. David Chynoweth, Dr. Gabriel
Bitton and Dr. Matthew Booth, who gave me great help.
I wish to thank my colleagues in the Solid and Hazardous Waste Research group, in
particular, Brajesh Dubey, Qiyong Xu, Kim Cochran, Steve Musson, Aaron Jordan,
Pradeep Jain, Jaehak Ko, Murat Semiz, Judd Larson, and Yong-Chul Jang, a faculty
memeber of Chung-Nam University in South Korea. I also thanks goes to my first mentor
and graduate advisor, professor, Byung-Ki Hur, a faculty of Inha University in South
Korea.
A special thanks goes to my mother as well as my father, who is fighting against
disease. Finally, greatest thanks go to my wife, Eunkyoung Choi, for her patience,
encouragement, and love.
IV

TABLE OF CONTENTS
page
ACKNOWLEDGMENTS iv
LIST OF TABLES viii
LIST OF FIGURES x
CHAPTERS
1. INTRODUCTION 1
1.1 Problem Statement 1
1.2 Objectives 2
1.3 Research Approach 3
1.4 Outline of Dissertation 5
2. COMPARATIVE STUDIES OF LEACHATE AND GAS QUALITY OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILL BIOREACTORS 6
2.1 Introduction 6
2.2 Material and Methods 7
2.2.1 General Description of the Lysimeter 7
2.2.2 Temperature Control 8
2.2.3 Fabricated Waste Stream 9
2.2.4 Air Injection 10
2.2.5 Leachate and Gas Analysis 10
2.2.6 Recovery of the Anaerobic Lysimeters 11
2.2.7 Prediction of Waste Mass Loss 12
2.3 Results and Discussion 12
2.3.1 pH 13
2.3.2 Organic Carbon Concentration 14
2.3.3 Nitrogen 16
2.3.4 Dissolved Solids Content 17
2.3.5 Oxidation Reduction Conditions 18
2.3.7 Gas Quality 19
2.4 Discussion 20
2.4.1 Differences between Aerobic and Anaerobic Lysimeters 20
2.4.2 The Comparison of Leachate Parameters with Other Studies 21
2.4.3 Implications for Full-scale Application 22
v

2.4.4Limitations 23
2.5 Conclusions 24
3. THE FATE OF HEAVY METALS IN SIMULATED LANDFILL
BIOREACTORS UNDER AEROBIC AND ANAEROBIC CONDITIONS 47
3.1 Introduction 47
3.2 Materials and Methods 48
3.2.1 Heavy Metal Sources in Synthetic Waste 48
3.2.2 Sampling Methods 49
3.2.3 Analytical Methods 49
3.3 Results and Discussions 50
3.3.1 Changes in Metal Concentrations versus Time and the Percentage of
Mass Loss 50
3.3.1.1 Aluminum 50
3.3.1.2 Arsenic 51
3.3.1.3 Chromium 53
3.3.1.4 Copper 54
3.3.1.5 Lead 56
3.3.1.6 Iron 57
3.3.1.7 Manganese and Zinc 58
3.3.2 Organic Wastes as Absorbents of Heavy Metals 59
3.4 Discussion 61
3.4.1 Overall Comparison of Metal Behavior 61
3.4.2 Comparison to Other Studies 63
3.4.3 Implication for Disposal of Heavy Metals 65
3.4.4 The Impact of Air on Metal Mobility 66
3.5 Conclusions 67
4. THE EVALUATION OF LIGNOCELLULOSIC WASTE DECOMPOSITION OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILLS 95
4.1 Introduction 95
4.2 Materials and Methods 97
4.2.1 Composition of Fabricated Waste 97
4.2.2 Excavation and Processing of Decomposed Solid Waste 97
4.2.3 Methane Yield Determination 98
4.2.4 Cellulose and Lignin Determination 100
4.2.5 Data Analysis 101
4.3 Results 101
4.3.1 Methane Yield of Raw Waste 101
4.3.2 Solid Waste Excavation 102
4.3.2 Mass Loss for Individual Components 103
4.3.3 Biodegradability of Excavated Wastes 104
4.3.4 Biodegradability of Wood Waste 105
4.4 Discussion 106
4.5 Conclusions 108
vi

5. LANDFILL SETTLEMENT BEHAVIOR WITH WASTE DECOMPOSITION 119
5.1 Introduction 119
5.2 Materials and Methods 120
5.2.1 Lysimeters 120
5.2.2 Application of Overburden Pressure 121
5.2.3 Compression Index and Phase Separate Method 122
5.2.4 Estimation of Mass Loss 123
5.2.5 Volume Loss versus Mass Loss 124
5.3 Results 125
5.3.1 Settlement Behavior over Time 125
5.3.2 The Relationship between The Settlement and Mass Loss 126
5.3.3 Ultimate Settlement 127
5.4 Discussion 127
5.4.1 Compression Index 127
5.4.2 Correlation of Mass Loss and Volume Loss 128
5.4.3 Application 129
5.5 Conclusions 131
6. SUMMARY AND CONCLUSIONS 141
6.1 Summary 141
6.2 The Implication of This Research 143
6.3 Conclusions 145
6.4 Future Work 147
APPENDIX
A. ADDITIONAL PROCEDURES AND CONCEPTS 149
A. 1 Prediction of Mass Loss by Gas and Leachate 149
A.2 Estimation of Biodegradable Volatile Solids (BVS) 152
A.3 Lysimeter Dismantlement 153
B. SUPPLEMENTAL FIGURES 159
C. LYSIMETER EXPERIMENT RAW DATA AND GRAPHS 169
C.l Graphs 169
C.2 Raw Data 189
LIST OF REFERENCES 219
BIOGRAPHICAL SKETCH 231
vii

LIST OF TABLES
Table page
2-1. MSW components 25
2-2. Parameters and methods for analysis 26
2-3. Comparison of initial and final characteristics of the aerobic lysimeters 27
2-4. Comparison of initial and final characteristics of the anaerobic lysimeters 28
2-5. Comparison of leachate parameters with other aerobic landfill studies 29
2-6. Comparison of leachate parameters with other anaerobic landfill studies 29
3-1. Heavy metal sources in fabricated waste stream 69
3-2. Results of statistical analysis of metal leached between aerobic and anaerobic 69
3-3. The amount of leachate produced and used for analysis 69
3-4. Leachability of As, Cr, and Cu 70
3-5. Comparison of cumulative mass of metal dissolved in leachate and adsorbed on
lignocellulosic materials 70
3-6. Comparison of average metal concentrations of the aerobic and anaerobic
lysimeters with MSW leachate and regulatory levels 71
3-7. Comparison of characteristics of CCA-treated wood used for Jambeck (2004) and
this study 71
4-1. Methane yields, VS and mass fraction of the lignocellulosic materials in raw
waste 109
4-2. Comparison of methane yields of MSW with other studies 109
4-3. Biodegradable volatile solid (BVS) of organic fraction of the raw waste 109
4-4. The physical characteristics of excavated waste 110
4-5. Overall methane yields of waste layers of the lysimeters 2 and 4 Ill
viii

4-6. Summary of cellulose and lignin content of the wood samples 112
5-1. (Ca)min and (Ca)max values of lys 1 through 4 132
5-2. k values of aerobic and anaerobic lysimeters 132
5-3. Comparison of compress indices between current study and other studies 133
A-l. Actual mass loss and predicted values of the aerobic and anaerobic lysimeter 155
A-2. Mass and density of wastes excavated by depth 155
C-l. pH of the aerobic and anaerobic lysimeters 189
C-2. Conductivity of the aerobic and anaerobic lysimeters 192
C-3. Alkalinity of aerobic and anaerobic lysimeters 194
C-4. Total dissolved solids (TDS) of aerobic and anaerobic lysimeters 196
C-5. Total organic contents (TOC) of aerobic and anaerobic lysimeter 197
C-6. Chemical oxygen demand (COD) of aerobic and anaerobic lysimeter 199
C-l. NH3+-N concentrations of aerobic and anaerobic lysimeters 201
C-8. Sulfide concentrations of aerobic and anaerobic lysimeters 203
C-9. Volatile fatty acids (VFA) of lysimeter 1 205
C-10. Volatile fatty acids (VFA) of lysimeter 2 206
C-l 1. Volatile fatty acids (VFA) of lysimeter 3 207
C-l2. Volatile fatty acids (VFA) of lysimeter 4 209
C-l3. Fabricated waste in lysimeters 211
C-14. Metal concentrations of lysimeter 1 212
C-l5. Metal concentrations of lysimeter 2 213
C-l 6. Metal concentrations of lysimeter 3 214
C-17. Metal concentrations of lysimeter 4 216
C-l 8. ANOVA results of metals and organic absorbence 218
IX

LIST OF FIGURES
Figure Page
2-1. Schematic of the lysimeter 30
2-2. The composition of fabricated municipal solid waste for this research 31
2-3. Comparison of pH between aerobic and anaerobic lysimeters versus time 32
2-4. Changes in COD of aerobic and anaerobic lysimeters versus time 33
2-5. Changes in BOD of aerobic and anaerobic lysimeters versus time 34
2-6. Changes in VFAs of aerobic and anaerobic lysimeters versus time (A) acetic acid
only and (B) acetic acid, propionic acid and butyric acid 36
2-7 Changes in the ratio BOD/COD of the aerobic and anaerobic lysimeters over time ..37
2-8. Changes in ammonia concentrations versus time 38
2-9. Changes in TDS of the aerobic and anaerobic lysimeters versus time 39
2-11. Changes in sulfide and pH versus time 41
2-12. The changes in sulfate and sulfide versus time in the presence of dissolved
oxygen 42
2-13. The changes in air injection rate and gas concentrations of aerobic lysimeter 43
2-14. Changes in gas concentrations of anaerobic lysimeter 4 44
2-15. Cumulative biogas vs. days in aerobic and anaerobic lysimeters 45
2-16. Changes in gas concentrations, pH and gas generation rate after air injection into
lysimeter 3 46
3-1. Changes of A1 concentrations over time 72
3-2. Changes of As concentrations over time 73
3-3. Changes of Cr concentrations over time 74
x

3-4. Changes of Cu concentrations over time 75
3-5. Changes of Pb concentrations over time 76
3-6. Changes of Fe concentrations over time 77
3-7. Changes of Mn concentrations over time 78
3-8. Changes of Zn concentrations over time 79
3-9. Distribution of As over a C-pH diagram 80
3-10. Potential- pH diagram of Cr 81
3-11. Distribution of Cu over a C-pH diagram 82
3-12. Adsorption of metal on solid wastes 83
3-13. The comparison of aerobic and anaerobic lysimeters in respect of total mass of
metals adsorbed on lignocellulosic materials 85
3-14. The comparison of metal concentrations adsorbed on organic (newspaper and
cardboard) and plastic waste 87
3-15. Fate of heavy metals thermodynamically occurred in aerobic (oxidizing) and
anaerobic (reducing) conditions 89
3-16. Comparison of concentrations of metal leached between aerobic and anaerobic
lysimeters 90
3-17. Changes in cumulative mass of meta released over a mass loss, % 92
3-18. Comparison of As, Cu and Cr leaching trend of the lysimeters to other study 94
4-1. The dry weight differences between predicted and measured remaining mass 113
4-2. Comparison of dry weights between raw and decomposed lignocellulosic wastes .114
4-3. The changes in the percentage of waste components after decomposition; (A) raw
waste components and (B) decomposed waste (aerobic) 115
4-4. Changes in cumulative methane volume of lignocellulosic materials over time 116
4-5. Methane yields and weight differences of lignocellulosic materials among raw
and two lysimeters (A) all lignocellulosic materials; (B) wood only 117
4-6. The comparison of dry masses measured and predicted by gas generated and
BMP assay 118
xi

5-1. The changes in settlement, cumulative gas (CO2) and pH over time 134
5-2. The changes in settlement, cumulative gas (CO2 and CH4) and pH over time 135
5-3. Settlement behaviors and compression coefficients of aerobic and anaerobic
lysimeter over a period of time 136
5-4. Relationship between settlement and overall mass loss of the aerobic and
anaerobic lysimeters 137
5-5. Relationship between percentage of settlement and mass loss 138
5-6. Correlation of logarithm of mass loss of the aerobic lysimeters over time 139
5-7. Different k values of anaerobic lysimeters at lag and log phases 139
5-8. Settlement prediction of the aerobic lysimeters 140
A-l. Schematic of mass loss by waste decomposition 156
A-2. Waste mass loss by TOC and gas generation 158
B-l. Schematics of aerobic and anaerobic lysimeters used for this research 159
B-2. The carriage system 160
B-3. A schematic of the temperature control system 161
B-4. Schematic of gas volume measuring tool; before gas measurement, fill tap-water
up to the top scale 162
B-5. The nation-wide composition of discarded municipal solid waste in 2003 163
B-6. The composition of municipal solid waste in Florida in 2000 163
B-7. (A) Blue water phenomenon observed from gas collection system of aerobic
lysimeters; (B) a hole on copper tube caused by corrosion of Cu 164
B-8. Solid samples excavated from one of the aerobic lysimeter 165
B-9. Decomposed papers were commingled together (aerobic lysimeter) 166
B-10. Not well degraded office paper (aerobic lysimeter) 167
B-l 1 Wood blocks excavated from aerobic lysimeter 167
C-l. The change in COD of the lysimeters over the percentage of mass loss 169
xii

C-2. The change in BOD5 of the aerobic and anaerobic lysimeters over the percentage
of mass loss 170
C-3. The change in ammonia of the aerobic and anaerobic lysimeters over time 171
C-4. The change in fluoride of the aerobic and anaerobic lysimeters over time 172
C-5. The change in chloride (Cl) of the aerobic and anaerobic lysimeters over time 173
C-6. The change in sulfate of the aerobic and anaerobic lysimeters over time 174
C-7. The change in calcium (Ca) of the aerobic and anaerobic lysimeters over time 175
C-8. The change in sodium (Na) of the aerobic and anaerobic lysimeters over time 176
C-9. The change in biogas produced from the aerobic lysimeters 177
C-10. The change in biogas produced from the anaerobic lysimeters 178
C-l 1. A1 concentration versus pH in leachate from the lysimeters 179
C-12. Cr concentration versus pH in leachate from the lysimeters 180
C-l3. Cu concentration versus pH in leachate from the lysimeters 181
C-14. Mn concentration versus pH in leachate from the lysimeters 182
C-l5. Pb concentration versus pH in leachate from the lysimeters 183
C-l6. Zn concentration versus pH in leachate from the lysimeters 184
C-l 7. Change in methane yields of the waste layer 2-1 and 2-2 185
C-l 8. Change in methane yields of the waste layer 2-3 and 2-4 186
C-l 9. Change in methane yields of the waste layer 4-1 and 4-2 187
C-20. Change in methane yields of the waste layer 4-3 and 4-4 188
xiii

Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy
COMPARATIVE STUDIES OF AEROBIC AND ANAEROBIC LANDFILLS
USING SIMULATED LANDFILL LYSIMETERS
By
Hwidong Kim
December 2005
Chair: Timothy G. Townsend
Major Department: Department of Environmental Engineering Sciences
Many proposals suggest that air injection into bioreactor landfills enhance waste
composition; several potential benefits of air addition have been hypothesized, yet little
has been proven about the overall performance of aerobic landfills compared with current
anaerobic landfills. Utilizing research conducted with six-foot tall stainless steel
simulated landfill lysimeters, complete with fabricated wastes, this Ph.D. dissertation
compares aerobic and anaerobic landfills with respect to gas and leachate quality, fate of
metals, settlement behavior and biodegradation of lignocellulosic materials.
Through air injection, a large enhancement of waste decomposition was observed.
More than 90% of the maximum chemical oxygen demand (COD), biochemical oxygen
demand (BOD) and total organic carbon (TOC) concentrations decreased within 100 days.
During the methanogenic phase in the anaerobic condition, concentrations of ammonia
increased by an amount four times greater than the initial concentrations. A large change
of ammonia was not observed from the aerobic lysimeters.
xiv

The fate of metals leached from the various metal sources including cathode ray
tube (CRT) monitor glass and ground CCA-treated wood were explored. Metal leaching
trends observed varied from anaerobic to aerobic lysimeters; the average concentrations
of As, Fe, Mn, and Zn in the anaerobic lysimeters proved significantly greater in
concentration than observed in the anaerobic lysimeters. Furthermore, significantly
greater concentrations of Al, Cu, Cr, and Pb were detected in the aerobic lysimeters as
compared to the anaerobic lysimeters.
Using leachate and gas measurements, mass losses from the aerobic and anaerobic
lysimeters were estimated. Mass removed from the wastes was primarily converted into
gas; after the water was removed from the lysimeters, the mass of waste excavated from
each lysimeter was compared with the estimated loss mass. For wood waste, no great
influence on air addition was observed through cellulose/lignin analysis. Methane
potential of lignoceliulosic materials other than wood waste resulted in great differences
of biodegradation between aerobic and anaerobic lysimeters.
The landfill settlement behavior occurring in aerobic and anaerobic simulated
landfills was mathematically analyzed. The logarithm of mass loss was linearly correlated
with the percentage of settlement. With this relationship, the secondary settlement of
bioreactor landfills could be mathematically modeled using the first-order exponential
function.
xv

CHAPTER 1
INTRODUCTION
1.1 Problem Statement
Landfills remain the predominant method for managing municipal solid waste
(MSW) in the U.S. Although modem engineered landfills protect the environment from
groundwater contamination and in some cases gas emissions, they are most often
operated in a fashion where only a small amount of the disposed waste is permitted to
biodegrade to a more stabilized state. This results in large amounts of undegraded waste
being stored for many years in the future; their management will continue to demand
resources and may pose a long-term environmental risk.
Alternatively, many innovative and more environmental-friendly strategies for
operation of MSW landfills have been proposed (Stegmann, 1983; Barlaz et al., 1992;
Komilis et al., 1999). Among these techniques, leachate recirculation has been found to
be the most practical approach for enhancing waste decomposition and stabilization in
landfills (Reinhart et al., 2002). This process stabilizes landfilled waste more rapidly
because of the increased moisture content and the more effective distribution of nutrients
and microorganisms in the landfill. This result creates a very favorable environment for
the existing anaerobic organisms responsible for waste degradation. If controlled,
methane produced can be utilized as a resource. This technique has changed the concept
of a landfill from a historical garbage dump to a bioreactor, where various biochemical
reactions are managed in a controlled fashion.
1

2
Air addition has been suggested as another means, in concert with leachate
recirculation, to achieve rapid landfill stabilization. It has been reported that waste
decomposes more rapidly in aerobic systems relative to anaerobic systems (Read et al.,
2001). Additional reports suggest that air injection may stop the production of methane
(one of the most serious greenhouse gases), change the leachate quality for the better,
reduce the amount of volatile organic compounds (VOCs), and improve the degradability
of anaerobically recalcitrant materials (Grima et al., 2000; Read et al., 2001; Lee et al.,
2002; Reinhart et al., 2002). Some of these potential benefits have been investigated at
the lab scale (Stessel and Murphy, 1992), and some positive outcomes have been reported
from field studies (Read et al., 2001; Lee et al, 2002). However, in order to apply this
new technique successfully to full-scale operating landfills, further investigation is
necessary. While anaerobic bioreactors have been heavily simulated in previous studies,
there are few cases involving the simulation of aerobic landfills. It is also rare to find
side-by-side simulations on the same waste stream under the same field conditions
comparing aerobic and anaerobic systems.
1.2 Objectives
The main objective of this research was to compare aerobic and anaerobic landfills
using simulated landfill lysimeters. In the early development of anaerobic bioreactors,
several fundamental simulated landfill experiments were performed that have provided
much of our understanding of such processes to date (Pohland, 1980). This research
presents the results of parallel aerobic and anaerobic simulated bioreactors. Several
different parameters of concern were investigated: leachate and gas quality, settlement,
heavy metal fate, and decomposition of lignocellulosic materials. The following were
specific objectives of this research:

3
o To compare leachate and gas quality between aerobic and anaerobic bioreactor
landfills,
o To explore the fate of heavy metals leached from the fabricated wastes in
aerobic and anaerobic bioreactor landfills,
o To explore the decomposition of lignocellulosic wastes in anaerobic and
aerobic bioreactor landfills, and
o To evaluate the loss of mass versus the loss of volume in aerobic and anaerobic
bioreactors for use in future settlement model development.
1.3 Research Approach
Four stainless steel lysimeters were constructed: two were operated aerobically and
two were operated anaerobically. These lysimeters were designed and constructed as part
of a previous research experiment (Sheridan, 2003). After operating the aerobic and
anaerobic lysimeters for 1 and 2 years, respectively, one aerobic and one anaerobic
lysimeter were dismantled. Waste samples were collected and characterized. The
remaining aerobic and anaerobic lysimeters were kept in operation so that waste
stabilization could be completely researched; the results of this extended operation will
be presented elsewhere.
To compare leachate and gas quality between the aerobic and anaerobic bioreactors,
two pairs of simulated landfill lysimeters containing fabricated wastes were operated as
aerobic and anaerobic bioreactors. The fabricated wastes were loaded into the lysimeters,
compacted, and mixed with water and seed (either anaerobic sludge or aerobic compost).
Leachate generated by the lysimeters was collected and analyzed for leachate quality
parameters. A mixture of collected leachate and deionized water was added back to the

4
lysimeters to compensate for the amount of leachate lost by leachate collection. The gas
volume and composition were monitored using a gas totalizer and gas chromatography.
To explore the fate of heavy metals leached out of the fabricated wastes under
aerobic and anaerobic conditions, heavy metal-containing wastes (e.g., CCA-treated
wood, cathode-ray tube (CRT) glass and pieces of sheet metal) were mixed with the other
fabricated wastes before loading into the lysimeters. After loading and compacting the
fabricated waste, leachate generated by the lysimeter was collected and analyzed for
copper, chromium, arsenic, lead, aluminum, zinc, manganese and iron. The change of
heavy metal concentrations in the leachate over time was monitored. After the lysimeter
work was completed, the wastes excavated from two columns were analyzed for heavy
metals in order to compare heavy metal concentrations absorbed on solid waste to those
released from the lysimeters through the leachate.
To explore the decomposition of lignocellulosic wastes in aerobic and anaerobic
landfill environments, lignocellulosic wastes including paper and wood blocks were
prepared. They were included in the fabricated waste and loaded in the lysimeters. After
the lysimeter study was completed, lignocellulosic wastes were excavated and separated.
Biochemical methane potential (BMP) assays were used to evaluate the degree of
biodegradation of each lignocellulosic waste. In order to evaluate the impact of air
addition on wood waste decomposition, cellulose, lignin and BMP of raw and excavated
wood blocks were compared with respect to cellulose and lignin concentrations and BMP
values.
To simulate landfill settlement in aerobic and anaerobic conditions as a function of
waste mass loss, overburden pressure was applied to the stainless steel lysimeters using a

5
hydraulic cylinder and hand pump. To correlate mass loss and volume loss, a lab-scale
experiment was designed where waste was decomposed in simulated landfills in the
laboratory with both mass loss and volume loss being measured. A difficulty with using
lab experiments to simulate landfill settlement is that it is hard to simulate true landfill
conditions, especially, the large overburden pressure. In this research, the experiments
included the application of overburden pressure to make the laboratory condition closer
to the field conditions.
1.4 Outline of Dissertation
The dissertation is presented in six chapters. The current chapter presents the
problem statement, objectives and research approach. Chapter 2 presents the comparison
of gas and leachate qualities between aerobic and aerobic simulated landfills. Chapter 3
presents the fate of heavy metals in aerobic and anaerobic simulated landfills. The
evaluation of biodegradation of lignocellulosic materials is presented in chapter 4.
Settlement behavior with waste decomposition is presented in chapter 5. Chapter 6
presents a summary, conclusions and recommendations for future work. Background and
other analytical procedures used for this research are presented in appendix A.
Supplemental s are presented in appendix B. All other tables and s pertaining to leachate
data and BMP are presented in appendix C.

CHAPTER 2
COMPARATIVE STUDIES OF LEACHATE AND GAS QUALITY OF AEROBIC
AND ANAEROBIC SIMULATED LANDFILL BIOREACTORS
2.1 Introduction
The operation of municipal solid waste (MSW) landfills as bioreactors for the
purpose of rapid landfill stabilization has historically been proposed as an anaerobic
process. Conditions within the landfill are controlled to accelerate the activity of the
anaerobic microorganisms responsible for waste decomposition. The addition of air has
also been proposed as a method to enhance landfill stabilization (Stessel and Murphy,
1992), and recently this technique has gained more attention (Read et al., 2001; Reinhart
et al., 2002). In addition to an enhancement of waste decomposition that is more rapid
than anaerobic operation, a major benefit often cited for air addition is the reduction in
methane emissions relative to anaerobic landfills (Borglin et al., 2004). These studies also
find that the overall strength of leachate (with respect to readily degradable carbon
compounds and oxygen demand) is lower in aerobic systems, offering a potential
advantage with respect to leachate treatment.
Research examining the relative differences in leachate quality between aerobic and
anaerobic systems is very limited. Though the performance of aerobic bioreactor landfills
has been simulated in several studies (Agdag and Sponza, 2004; Warith and Takata,
2004), these studies are often limited with respect to their ability to control several key
parameters, and their lack of a complementary anaerobic system for comparison purposes.
This chapter reports the results of research performed to examine the characteristics of
6

7
leachate and landfill gas that result from aerobic and anaerobic operation of identical
MSW streams. The experiments conducted involved a technique long employed in the
study of landfills: waste-filled columns constructed and operated to simulate landfill
processes, referred to here as lysimeters (Pohland, 1980). The columns were designed
and operated to control several parameters not traditionally simulated in such
experiments, such as temperature and overburden pressure. The objective was to compare
leachate and landfill gas quality between each type of system so that similarities and
differences can be better understood and to assist in future decision-making, design and
operation efforts. Several complementary objectives were evaluated as part of this
experiment and they are described in greater detail in Chapters 3 (fate of metals), 4
(comparison of decomposition) and 5 (comparison of settlement).
2.2 Material and Methods
Four lysimeters were used in this research, and each consisted of a stainless steel
column and a carriage system component. The original design and construction of the
lysimeters used for this research were described previously by Sheridan (2003). Two
were operated aerobically (lysimeter 1 and 2) and two were operated anaerobically
(lysimeter 3 and 4). Three parameters, temperature, air addition, and overburden pressure,
were controlled in an effort to simulate actual aerobic or anaerobic bioreactor landfills.
2.2.1 General Description of the Lysimeter
A schematic of each lysimeter type is presented in Figure 2-1 (see Figure B-l for
additional detail). The 6-ft stainless steel main body contained 5 front ports, 2 back ports
and 1 valve at the bottom for leachate collection. The front ports were used for air
addition (in the case of the aerobic lysimeters). The carriage system component was
designed to support a hydraulic pressurizing unit installed at the top of each lysimeters

8
for the application of an external load to the fabricated waste. The carriage system
consisted of a hydraulic cylinder, carriage, steel shaft, and steel plate. A small port
located on the top flange was used as a pathway for liquid addition. Perforations in the
steel plate allowed added liquid to percolate into the waste (see Figure B-2 for a detail of
the carriage system).
2.2.2 Temperature Control
The temperature at the center of a full-scale landfill usually remains constant
because the garbage and cover soil serve to insulate the system (McBean et al., 1995). In
a laboratory environment, however, the heat produced by biologically degrading waste is
not sufficient to maintain a temperature close to those normally encountered in a landfill.
Thus, a temperature control system was designed and constructed (Figure B-3).
The temperature of each lysimeter was measured using a type T thermocouple wire
(SRT201-160, Omega) fixed on the outside of each lysimeter. Two temperature
controllers (MC240, Electrothermal) were utilized in series to maintain desired
temperatures without extreme fluctuations. The lysimeters were insulated with 5-cm-
thick fiberglass and bubble insulation to minimize heat loss. Prior to operation, the
lysimeters were filled with tap water and the temperature controllers were tested by
measuring the temperature of the water.
The temperature of the aerobic lysimeters was maintained at a constant 55C for
the entire operating period. The anaerobic lysimeters were started at 35C and at day 400,
the temperature was increased from 35C to 55C at a rate of 2C per day. Although 55C
is in the optimum range for thermophilic anaerobic waste decomposition (Rittmann and
McCarty, 2001) and is often encountered in landfills (Watsoncraik et al., 1994;

9
Townsend et al., 1996), 35C was used as the starting point because the anaerobic seed
used was from a mesophilic digester.
2.2.3 Fabricated Waste Stream
The waste stream fabricated for this research was based on typical MSW
composition estimates previously reported for the U. S. and Florida (see Figure B-4 and
B-5). For simplification purposes, several minor components, such as textiles and tires,
were excluded from the fabricated waste stream. A greater portion of commingled paper
was allotted as a substitute for those excluded materials. The relative amount of office
paper, cardboard and newsprint in commingled paper (4.6 : 2.6 : 1) was again estimated
from previous published data (FDEP, 2003 and USEPA, 2005). Figure 2-2 presents the
fabricated waste stream composition used. Table 2-1 presents a description of each
component, the source, and the method of sample preparation. Commercial grade dog
food (Pedigree, USA) was used as the food waste portion of the fabricated waste stream.
To support complementary research on the fate of certain heavy metals in aerobic and
anaerobic landfill environment (chapter 3), a part of the wood waste fraction was
comprised of CCA and a part of the glass fraction was comprised of leaded cathode ray
tube (CRT) glass. Detailed waste components and their sources are presented in Table 2-
1 and Figure 2-2.
Mixed fabricated waste samples were created and loaded into the columns as four
distinct fractions to prevent waste component stratification in a particular place in the
column, (composition of the fabricated waste fractions and their weight are summarized
in appendix C). Prior to loading, 6 inches (15.3 cm) of river rock was placed at the
bottom of each lysimeter, and a geotextile was placed between the rock and waste. Each
waste fraction was then loaded and compacted until it occupied 25 % of the depth of the

10
lysimeter. Two liters of DI water were added along with the compaction of each waste
fraction. After loading, 11 L of additional water was added from the top of each lysimeter.
The goal of adding water was to bring the waste in each lysimeter, at the beginning of the
experiment, to field capacity. A capacity of 58% was targeted as this was the field
capacity measured for this waste under the initial compaction conditions of the lysimeter.
The waste was compacted to a density of 30 lb/ft dry (480.6 kg/m dry)-
2.2.4 Air Injection
Two computer-controlled pump drives (Model No. 7550-10, Cole-Parmer) were
used for air addition. Air was saturated and warmed prior to injection to keep moisture in
the waste from evaporating. Air was injected on the ports located at the side of the
aerobic lysimeters using a manifold from day 1 to day 164 and changed to the most
bottom port from day 164 to the end of a test period. A flow rate of 70 mL/min was found
to be suitable for control purposes and to maintain low exit gas oxygen concentrations.
The flow rate was adjusted several times during the experiment when oxygen
concentrations in the exit gas became less than 1% to maintain aerobic conditions.
2.2.5 Leachate and Gas Analysis
Leachate samples were collected on a weekly basis. Leachate was analyzed for
sulfide and dissolved oxygen immediately; analysis for pH, alkalinity, and conductivity
was carried out within one hour after collection. After this initial analysis, 15 mL of
leachate was preserved with sulfuric acid and placed in acid-rinsed high-density
polyethylene (HDPE) bottles for later analysis of chemical oxygen demand (COD), total
organic carbon (TOC), volatile fatty acids (VFA) and ammonia. For metal analysis, 50
mL of leachate was preserved with concentrated nitric acid and stored at 4C. The
remaining leachate was recirculated back to the top of the lysimeters. Deionized water

11
was added to make up for the amount of leachate used for analysis. Table 2-2 summarizes
the parameters and methods used for each analysis.
Biogas samples generated from both the aerobic and anaerobic lysimeters were
collected and analyzed for methane, carbon dioxide, and oxygen. For the aerobic
columns, the gas volume was measured using a wet-tip gas meter. For the anaerobic
lysimeters, the gas was gathered in 5-L and 10-L air-sampling bags, and the volume
contained in the bags was measured using the water-gas replacement method (see Figure
B-4). A LANTEC GEM 500 (SAIC, San Diego, CA) gas meter was used for gas analysis
for both the aerobic and anaerobic lysimeters. Additionally, gas samples collected from
the anaerobic lysimeters were analyzed for CH4 and CO2 using a gas chromatograph
equipped with a GS-Carbon plot column (Agilent Technology, Palo Alto, CA) to confirm
the measurements analyzed by LANTEC GEM500 gas meter.
2.2.6 Recovery of the Anaerobic Lysimeters
Since both of the anaerobic lysimeters (lys 3 and 4) remained in an acidic condition
(pH < 6) for 500 days, 100 g of sodium bicarbonate was added as a buffer to the top of
each lysimeter on day 300. The pH of the top part of the lysimeters changed to neutral,
but the pH of leachate collected from the bottom port remained low (5 to 5.5). The pH of
the leachate from lysimeter 4 increased to pH 7 from day 400. Since only minimum
changes in leachate pH of the lysimeter 3 were observed after buffer addition, additional
sodium bicarbonate was added to the bottom port rather than to the top of the lysimeters;
a total of lOOg of sodium bicarbonate was added (20 g each were added on days 420,
453, 469, 532, and 555 again). Only a temporary increase in pH was observed after this
addition. As a next step in increasing pH, lab air was injected into lysimeter 3 on day 627.
Before air injection, the methane concentration of the lysimeter 3 was 35%, and the pH of

12
leachate was 6.11. Lab air was injected with 70 mL/min from the bottom of the lysimeter
for five days. The changes in pH and output gas qualities were monitored on a daily
basis. The impact of this addition is discussed in the results section of this chapter.
2.2.7 Prediction of Waste Mass Loss
As described in the following sections in this chapter, the aerobic lysimeters more
quickly stabilized the waste in comparison to the anaerobic lysimeters, and thus their
period of operation was shorter (379 days vs. 741 days). In an effort to normalize the
leachate measurements among the different columns, the biogas data, the leachate data
and the initial content of the waste was used to estimate the percentage of waste
decomposition for a column at any given time. The detailed procedure for this is
presented in appendix A, but, in short, the cumulative volume of biogas measured at any
given time (CH4 + CO2 for anaerobic columns and CO2 for aerobic columns) was used to
calculate the mass of initial waste degraded at that time. This was adjusted to account for
the mass of organic carbon solubilized in the leachate. The mass of waste estimated to be
degraded at a given time was divided by the estimated total potential mass loss in each
column (this total potential mass loss was estimated from measured methane yields of the
raw waste; see chapter 4 for details).
2.3 Results and Discussion
The data presented for the aerobic and anaerobic lysimeters in this dissertation
represent operation periods of 379 and 741 days, respectively. At the end of each
operation period, one each of the aerobic and anaerobic lysimeters was stopped and
emptied. The remaining lysimeters were left operational (data are not reported here).
Values of all leachate parameters analyzed for this research are presented in Table C-l

13
through C-l 1 in appendix C. These include the raw data and graphs of the leachate
parameters.
2.3.1 pH
Figure 2-3 depicts the change in pH over the course of the experiment. Both the
aerobic and anaerobic lysimeters remained in acidic condition during the beginning of the
experiment. The period of time required to stabilize the pH for the aerobic and anaerobic
lysimeters was 200 and 600 days, respectively. Average pH measurements of
approximately 8.9 (aerobic) and 7.1 (anaerobic) were observed at the end of the
experiment.
Two phases (acidic and alkaline or methane phase) of the pH of the aerobic and
anaerobic lysimeters were observed during a test period. The low pH occurring during the
initial phase of the research was attributed to a build up of organic acid concentrations
and the related microbial activities. Once the organic waste decomposition process began,
the biodegradable fraction of waste was converted into organic acids by various
biological reactions, and the accumulation of the organic acids lowered the pH. For the
aerobic lysimeters, air was injected through four front ports of the lysimeter using
manifolds. The pH was low (< 6) for the first 150 days, and high VFA and alkalinity
concentrations indicated that anaerobic conditions were predominant, suggesting that air
was not evenly distributed through the manifolds. An increase in the pH of the aerobic
lysimeters was observed after air was injected into the only bottom port. Typically, the
pH of the system increases to neutral conditions as the organic acids are consumed by
methanogenic bacteria. A large amount of CO2 production in an unbalanced ecosystem
may also contribute to lowering the pH as well. High concentrations of VFA and
alkalinity were measured in the anaerobic lysimeter leachate during the initial acid phase.

14
The pH of the aerobic lysimeters measured in the latter half of the experiment (9.0)
was more alkaline than the pH measured from the anaerobic lysimeters (7.2). According
to other lysimeter studies, higher pH was observed from the aerobic lysimeters in
comparison with that of the anaerobic lysimeter. The range of pH of aerobic lysimeters
has been reported as 7 9 (Stessel and Murphy, 1992; OKeefe and Chynoweth, 2000;
Agdag and Sponza, 2004). Summerfelt et al. (2003) also observed an increase of pH
when air was injected into their aquaculture system. They reported that this increase was
because of CO2 stripping by air; a decrease in CO2 leads to a decrease of carbonic acid
(H2CO3) and bicarbonate concentrations (HCO3) consuming H+ ions. These relationships
can be described by carbonate systems as follows:
C02gas<-> H2CO3 (1)
H2CO3 HC03' + H+ (2)
HC03 C03 + H+ (3)
They additionally concluded that, because the dehydration of carbon acid is rate-limiting,
pH may not increase instantaneously.
2.3.2 Organic Carbon Concentration
Figure 2-4 depicts the change of COD concentrations for the lysimeters versus
time. The initial average COD concentrations in the leachate of the aerobic and anaerobic
lysimeters were 36,000 mg/L and 66,000 mg/L, respectively. The COD values for the
aerobic lysimeters increased up to greater than 84,000 mg/L and decreased rapidly after
pH was stabilized. Although one of the aerobic lysimeters showed high COD (70,000
mg/L) at day 50, the overall COD concentrations of the aerobic lysimeters were lower
than values in the anaerobic lysimeters. Similar trends occurred for COD as were
observed for BOD5 (Figure 2-5). The BOD values of the aerobic lysimeters decreased

15
rapidly down to below 100 mg/L from day 200 while BOD values of the anaerobic
lysimeters decreased relatively slowly.
The primary contributor to high COD or BOD concentrations in landfill leachate
is volatile fatty acids (McBean et al., 1995). Figure 2-6 (a) depicts the changes in acetic
acid, one of the major volatile fatty acids (VFA), as a function of mass loss. Acetic acid is
used as a substrate by methanogenic bacteria and contributes to the formation of an acidic
environment when they are unbalanced with the growth of methanogenic bacteria. For
these reasons, VFA concentrations are used as an indicator to assess landfill conditions
(USEPA, 2004). For example, the decrease in acetic acid concentration in lysimeter 3 and
4 corresponds to the point when the pH began to rise. In the aerobic lysimeters, high
concentrations of acetic acid were noted during the first phase of the experiment due to
improper air distribution as discussed earlier. Acetic acid in leachate from the aerobic
lysimeters was degraded to less 1 mg/L by day 200, which corresponds with the time
required to deplete COD and BOD.
Among different types of short carbon chain fatty acids, acetic, propionic and
butyric acids are known as major VFAs that are involved in biodegradation processes in
anaerobic conditions. Production and degradation of these major VFAs in selected
aerobic and anaerobic lysimeters are presented in Figure 2-6 (b). In both aerobic and
anaerobic lysimeters, the concentration of VFAs was mainly: acetic acids > butyric acids
> propionic acids. These results are similar to those found by Parawira et al (2004). They
also explained that high butyric acids were mainly attributed to high carbohydrates in
waste. Under the same condition, the degradation of VFAs in anaerobic condition was
found to be in the following order: butyric acids > acetic acids > propionic acids. Wang et

16
al. (1999) explained that various enzymatic reactions in microorganisms dictate a greater
decreasing rate of butyric acids than that of other VFAs. However, more biosynthetic
processes are involved in butyric acid production than acetic acid due to longer carbon
chains. In aerobic lysimeters, all three major VFAs were depleted together like other bulk
organic carbon.
The ratio of BOD5 to COD is often used to assess the biodegradability of the
organic matter in leachate, and thus to assess the degree of landfill stabilization. In old
stabilized landfills, the BOD5/COD ratio is below 0.10 (Kjeldsen et al, 2002). A low
BOD5/COD suggests that a leachate is low in biodegradable organic carbon and relatively
high in hard-to-biodegrade organic compounds such as humic compounds. In this
research, low BOD5/COD ratios were observed with the aerobic lysimeters after day 200
(0.04 on average) (Figure 2-7). Relatively high BOD5/COD ratios were exhibited from
the anaerobic lysimeters (0.36 on average). These values fall into the range of average
BOD5/COD ratios proposed by Kjeldsen et al. (2002) for the acid phase (0.58) and the
methanogenic phase (0.06).
2.3.3 Nitrogen
Figure 2-8 shows the changes in ammonia-nitrogen in the lysimeter during the
course of experiment. Ammonia concentrations from the aerobic lysimeters remained
relatively constant, showing a general increase during the course of the experiment. In a
different fashion, ammonia concentrations in the anaerobic leachate increased
dramatically at a point corresponding to an increase in system pH. Ammonia
concentrations in the anaerobic lysimeters increased to concentration in the range of
1000-1600 mg/L. These values then dropped to 800-1000 mg/L and stabilized. Small
increases of ammonia concentrations in leachate of aerobic column were observed after

17
day 180, but they were still approximately 4 times lower than values of anaerobic
lysimeters. The trend of increases in ammonia concentrations also can be found in
operating bioreactor landfills (Reinhart and AlYousfi, 1996).
Since ammonia is generally produced from the deamination process of amino
acids (a monomer of proteins), elevated ammonia concentrations may be associated with
protein decomposition. Cali et al. (2005) reported that an active methanogenic bacteria
community increased the ammonia concentration. Several researchers have proposed that
the enhancement of waste decomposition and leachate recirculation in anaerobic
bioreactor landfills results in increased ammonia concentration (Reinhart and Al-Yousfi,
1996; Berge et al., 2005).
2.3.4 Dissolved Solids Content
Figure 2-9 shows the change in total dissolved solids (TDS) in the aerobic and
anaerobic lysimeters through the course of the experiment. For both aerobic and
anaerobic lysimeters, like the change in other organic matter, TDS concentrations were
lower as the pH was stabilized. TDS of the aerobic lysimeters were rapidly stabilized
approximately 8 to 10 g/L after day 200. TDS of the anaerobic lysimeters were still
greater than that of the aerobic lysimeters but TDS of lysimeter 4 fell below 20 g/L on
day 700. Typical TDS concentration in landfills is within the range of 2 to 60 g/L
(Kjeldsen, 2002).
Figure 2-10 depicts the change in alkalinity in the aerobic and anaerobic
lysimeters through the course of the experiment. For the aerobic lysimeters, the alkalinity
increased to 16,000 mg/L as CaCC>3 in the lysimeter 1, but another lysimeter showed low
alkalinity which was below 2,000 mg/L as CaCC>3 but it increased 8,000 mg/L as CaCC>3
again. The alkalinity was lowered below 2000 mg/L as CaCC>3 for both aerobic

18
lysimeters after the pH was stabilized at alkaline condition. For the anaerobic lysimeters,
relatively high alkalinity was maintained over the entire test period. Generally, alkalinity
could be generated by CO2 accumulation and ammonification (Fannin, 1987). It is also
consumed by nitrification (Gujer and Jenkins, 1974).
2.3.5 Oxidation Reduction Conditions
Figure 2-11 depicts the change of sulfide concentrations in the aerobic and
anaerobic lysimeters over a period of time. Little changes of sulfide were observed during
the acid phase of both aerobic and anaerobic lysimeters. However, rapid increases in
sulfides along with an increase of pH were exhibited from the aerobic lysimeters and
lysimeter 4. The sulfide level of lysimeter 1 was lowered on day 210 again, but increased
up to 2,600 pg/L during the alkaline phase. The trends of change in sulfide concentration
of lysimeter 2 exhibited are similar to that of lysimeter 1. The highest sulfide level found
in lysimeter 2 was 1,200 pg/L. In lysimeter 4, sulfide concentrations increased as the pH
increased.
It is notable that great concentrations of sulfide were found in the system where air
injection had been taking place. It is hard to understand how sulfide could be presented in
an aerobic environment. In comparison with sulfate concentrations, sulfide was formed
by sulfate reduction (Figure 2-12). However, Figure 2-10 shows that high dissolved
oxygen concentration was also found in the same condition. Snoeyink and Jenkins (1980)
reported that sulfide could be detected under aerobic conditions. They explained that this
phenomenon was caused by a non-equilibrium situation for the reaction between oxygen
and sulfide. Therefore, it can be concluded that sulfide can be found before it oxidizes for
a second time by dissolved oxygen. This result indicates that anaerobic zones were

19
presented in the aerobic lysimeters producing sulfide. Relatively high ammonia
concentrations found from the aerobic lysimeters (Figure 2-9) also indicated the presence
of anaerobic zones in the aerobic lysimeters.
2.3.7 Gas Quality
Biogas emitted from the aerobic and anaerobic lysimeters was measured for O2,
CO2 and CH4. Figures 2-13 and 2-14 depict the changes in gas concentrations of aerobic
and anaerobic lysimeters over a period of time. The initial air injection rate of the aerobic
lysimeters was 70 mL/min. The air injection rate was regulated by changes of oxygen
levels within the range of 70 to 120 mL/min. High CO2 concentrations were observed
from aerobic lysimeters during the first 50 days, but decreased to lower than 20%. The
concentrations of CH4 and CO2 of the anaerobic lysimeters changed during the acidic
phase, but stabilized to approximately 60% CH4 and 40% CO2 during the methane phase.
Overall, a total of 40,100 liters (1,400 ft3) of gas was injected into each aerobic
lysimeters for a test period, and 45% and 43% of the oxygen included in the air added
was converted into CO2 in lysimeter 1 and 2, respectively. In the anaerobic lysimeters,
500 and 1,600 liters of biogas (CO2 and CH4) were produced from lysimeters 3 and 4.
Most of the gas generated was mainly concentrated on the methanogenic phase in
anaerobic lysimeters while a relatively steady gas generation was exhibited over time in
aerobic lysimeters as summarized in Figure 2-15.
Lab air was added to recover the lysimeter 3, which had remained in acidic
condition (pH <6) for 600 days. Figure 2-16 depicts the change in gas concentrations, gas
generation rate and pH during air injection. The pH was adjusted to 7.1 at day 4, and air
injection was stopped on day 5. After oxygen was depleted in the lysimeter, methane
concentrations substantially increased along with biogas generation rate and reached

20
above 50% at the 9th day. The biogas generation rate of lysimeter 3, after air injection,
was 2.3 L/day on average for 10 days. During the rest of test period, the pH of the
lysimeter 3 went down to 6.5, but further decrease was not observed. In comparison with
the conditions of lysimeter 3 before air addition, the amount of biogas produced was
substantially increased and high percentage of methane (> 55%) was maintained (Figure
2-16).
2.4 Discussion
2.4.1 Differences between Aerobic and Anaerobic Lysimeters
The largest differences between the aerobic and anaerobic lysimeters can be found
from the enhancement of waste biodegradation. Based on the leachate quality results of
this study, a period required for the aerobic lysimeters to decompose 90% of BOD was
160 days in the aerobic lysimeters while more than 700 days were required for the
lysimeter 4.
Other differences between the two systems were the methane concentrations
contained in exit gas. Air addition to the aerobic lysimeters lowered CH4 concentrations
in the exit gas dramatically. Though a small amount of methane was found in the exit gas
of the aerobic lysimeter, it was less than 1 % of the CO2 gas generated.
It is noted that pH had a relatively low impact on waste decomposition in the
aerobic lysimeter. Though the pH of the aerobic lysimeters was acidic, settlement
consistently occurred (see Chapter 5). It was probable that the acidic condition was
localized only on the bottom part, where air was not supplied properly.
Tables 2-3 and 2-4 present the initial and final characteristics of the aerobic and
anaerobic lysimeters. The data presented in Table 2-4 for the anaerobic lysimeters
indicate that these systems were not stabilized yet. Water loss from the aerobic lysimeters

21
was calculated using the water carrying capacity of the exit gas assuming that the gas was
100% saturated with water vapor. The overall performance of the aerobic lysimeters with
respect to waste decomposition and leachate quality was substantially greater than those
of the anaerobic lysimeters. However, the concentrations of sodium in the aerobic
lysimeters were still too high to meet drinking water standards. The final concentration of
ammonia in the aerobic lysimeters was also substantially higher than the criteria value of
ambient water (0.897 at 30C and pH 8.0) (USEPA, 1999). Although large quantities of
waste were decomposed, leachate of the anaerobic lysimeters still contained high
concentrations of organics, ammonia and anions (Table 2-4). Leachate generated would
be used for recirculation, but excessive volume of leachate must be treated at an on-site
or off-site wastewater treatment plant.
2.4.2 The Comparison of Leachate Parameters with Other Studies
Tables 2-5 and 2-6 summarize the comparison of leachate constituent
concentrations from this study with those from other studies. For aerobic landfill studies,
the maximum concentrations of COD, BOD and ammonia in this study appeared to be
greater than those of other studies, but they were in a similar range overall. The pH of the
aerobic lysimeters of this study was, however, greater than other studies. As previously
discussed, this would be because of the relocation of carbonic acids, bicarbonate, and
carbonate due to CO2 removal by air stripping (Summerfelt, 2003). The high pH of the
aerobic lysimeters implies that great concentrations of carbonate ions were dissolved due
to high partial pressure of CO2, and these carbonate ions might consume more H+ ions
when CO2 was removed. If alkalinity data of other studies were available, it would be

22
clear to describe the differences between this and other studies by comparing the
concentrations of carbonate ions.
2.4.3 Implications for Full-scale Application
Unlike the lab-scale simulated landfill, it is extremely difficult to aerate an entire
large-scale landfill. Highly compacted wastes make it difficult for an air stream to
penetrate into the recesses of a landfill. Moreover, leachate characteristics resulted from
air addition may be variable. As the analytical results have shown, leachate
characteristics of lysimeters 1 and 2 were different, despite starting with the same waste
stream and the same operational condition. The leachate characteristics of lysimeter 1
were similar to those of the anaerobic lysimeters during the first 180 days showing great
concentration of organic matter despite air addition. This was because of the large
anaerobic zones formed at the bottom of the lysimeter by improper air addition to the
bottom.
Aerobic zones can be formed around air injection wells but anaerobic zones may
still be present in the same landfill. However, coexistence of the aerobic and anaerobic
zones can be used for recovery of acid-stuck sour landfills. In this research, the air
addition was conducted under the hypothesis that environments formed by aeration for a
short period can be favorable to anaerobic microorganisms. A great amount of VFAs,
which caused acidic conditions, may be rapidly consumed by aerobes living in a
relatively wide pH range. Conversion of carbonic acid (HaCCV) to CO2 caused by air
stripping may increase the pH. With air addition with low flow rate, the anaerobic zones
may be protected from oxygen intrusion because oxygen may be depleted by the
respiration of aerobes. An additional technical strategy would be to add buffer such as
lime along with air addition. Buffer added may increase the alkalinity concentration.

23
Without high alkalinity, the pH of the landfill may decrease again when air addition is
stopped. This could happen when methanogenic bacterial population was not enough to
adapt to the new condition.
As Reinhart et al. (2002) pointed out, the reduction of leachate volume due to air
stripping could be one of the advantages of the aerobic landfills. In this research, a total
of 31m3 of air was added during a test period (1 year). Comparing the volume of water
initially added with final leachate volume, approximately 21 % of leachate volume was
reduced. Reduced volume of leachate implies that the operation of aerobic landfills can
be economical in terms of saving the cost for the leachate treatment.
2.4.4. Limitations
Since CH4 gas is one of the gases causing global warming, CH4 reduction can be
one of the advantages of the aerobic lysimeters. However, landfill gas released without a
flare system could be adverse to the environment. Berge et al. (2005) and Reinhart et al.
(2002) pointed out that various kinds of volatile organic compounds (VOC) and nitrous
oxide, a more potent greenhouse gas than methane, can be emitted without the flare
system. Future research is required to identify the gas constituents and develop the
filtering system as an alternative.
As previously mentioned, an extra monitoring job may be required to check
moisture content and gas contents around the gas injection well. Certain ratios of methane
and oxygen can be flammable according to Coward and Jones (1952) and Liao et al.
(2005). While air was added, CH4 concentrations were low, but the unpredictable
changes in O2 and CH4 concentrations were observed from the aerobic landfill (Read et al,
2001) and high concentrations of CH4 and 02 could coexist when air addition starts (Lee
et al, 2002).

24
2.5 Conclusions
In this research, the gas and leachate quality from aerobic and anaerobic
simulated bioreactor landfills were compared. Waste streams referenced from EPA and
FDEP were loaded into 4 stainless steel lysimeters with a density of 3500 kPa. All
lysimeters were prepared with the same conditions, and two of them were assigned for
aerobic and two for anaerobic bioreactor landfill simulations. Leachate and gas generated
from the lysimeters were analyzed for chosen parameters to make comparisons between
aerobic and anaerobic landfills.
Leachate analysis results indicated that organic compounds as measured by COD,
TOC, BOD and VFAs in the aerobic lysimeters were degraded more rapidly than those in
anaerobic lysimeters. Except for the acidic phase, the pH of the aerobic lysimeters rapidly
increased and was stabilized around pH 9.0 while anaerobic lysimeters had remained in
acidic phase for more than 400 days, and stabilized exhibiting pH 7.3. The concentrations
of ammonia in anaerobic lysimeters increased along with an increase of pH. Ammonia
concentrations in aerobic lysimeters varied little over time, but ammonia levels were
significantly lower than those of anaerobic lysimeters. Sulfide results imply that both
aerobic and anaerobic zones were coexisting in aerobic lysimeters. This may be caused
by the limit of oxygen distribution in the lysimeters because of high density and low
hydraulic conductivity of wastes under overburden pressure.

25
Table 2-1. MSW components.
Waste components
Sources
Processing for size reduction
Office paper
Mixed scrap paper purchased at
office supply store
Grind with a paper shredder
Cardboard
Mixed corrugated boxes
Scissors and razor blade
Newspaper
Local newspaper
Grind with a paper shredder
Plastics
PET bottles collected from a
plastics recycler
Scissors
Food waste
Commercial dog food
Grind with a coffee grinder (less
than 1/32)
Southern yellow pine (SYP)
Home improvement store
Cut with band-saw (2 x 2)
CCA-treated wood
Home improvement store
Gather saw dust after drilling
Galvanized steel
Home improvement store
Cut with metal cutter (1/2 x
1/2)
Aluminum
Home improvement store
Cut with metal cutter (1/2 x
1/2)
Cathode-Ray Tube(CRT) glass
CRT monitors
Crush with a hammer (1/4-1/8)
Mixed cullet
Mixed container glass
Crush with a hammer (1/4 1/8)

26
Table 2-2. Parameters and methods for analysis.
Parameters
Method
Alkalinity
Standard Method 2320B
Ammonia
Standard Method 4500-D
BOD
Standard Method 521 OB
COD
HACH 2720
Conductivity
Standard Method 2510
PH
Standard Method 4500-Fl+
TOC
EPA SW846, Method 9060
Sulfide
HACH 8131
Sulfate, Floride and Chloride
EPA SW846, Method 9056
Sodium
EPA SW846, Method 9060A
VFA
VFA analysis method using GC
(Innocente et al 2000)

27
Table 2-3. Comparison of initial and final characteristics of the aerobic lysimeters
Initial
Final (1 year)
Lys 1
Lys 2
Lys 1
Lys 2
Water quantity (mL)
19,000
19,000
15,137*
15,582(15,056*)
Dry waste quantity (g)
12,784
12,784
8,389*
8,740 (8,715*)
pH
5.7
5.7
8.5
8.5
COD (mg/L)
20,000
28,000
3,400
4,700
BOD
13,000
16,000
200
30
TOC
6,000
7,000
2,600
2,200
Ammonia
70
40
500
250
Fluoride
80
30
0
0
Chloride
200
130
1,200
1,700
Sodium
80
140
800
900
* predicted

28
Table 2-4. Comparison of initial and final characteristics of the anaerobic lysimeters
Initial
Fina
(2 years)
Lys 3
Lys 4
Lys 3
Lys 4
Water quantity (mL)
19,000
19,000
18,844*
18,833 (18,704*)
Dry waste quantity (g)
12,784
12,784
11,290
9,258 (8,997*)
PH
4.5
4.9
6.5
7.4
COD (mg/L)
65,000
67,000
42,000
24,000
BOD
48,000
62,000
14,000
6,500
TOC
26,000
27,000
12,000
5,600
Ammonia
120
100
1,000
800
Fluoride
1,500
1,400
460
200
Chloride
1,450
1,400
670
500
Sodium
2,000
2,000
4,800
3,800
* predicted

29
Table 2-5. Comparison of leachate parameters with other aerobic landfill studies
Parameters
(mg/L
except for
pH)
Compost
study3
Lysimeter
study lb
Lysimeter
study 2C
Lysimeter
study 3d
Lysimeter
study 4e
This study
Air flow
rate
20L/min
(once per a
week)
38L/min for
30min. at
12-h
intervals
8.4-
1300L/min
20mL/min
70-
120mL/min
COD
2434-31812
861-22026
130-23000
500-5000
2-1000
3400-47000
BODs
8-11571
100-10000
10-2000
30-45000
Ammonia
98-558
260-630
7-400
2-100
40-700
TDS
3300-11400
700-7700
EH
7.1-8.2
5.17-7.98
5.24-7.5
7-9
7.8
4.5-9.1
Krogmann and Woyczechowski, 2000; Agada and Sponza, 2004; cWarith and Takata, 2004; Stessel and
Murphy, 1992; eBorglin et al., 2004
Table 2-6. Comparison of leachate paramel
ers with other anaerobic landfi
1 studies
Parameters
(mg/L
except for
pH)
Conventional
landfill Ia
Conventional
landfill 2b
Bioreactor
landfills'
Bioreactor
landfill lab
scaled
This study
COD
140-152000
1000-40000
20-17000
100-88000
2000-80000
BODj
20-57000
50-25000
0-10000
6600-60000
TOC
30-29000
7000-19000
Ammonia
50-2200
50-1500
76-1850
100-1600
TDS
2000-60000
2000-25000
18000-50000
PH
4.5-9
3-7.5
5.4 -8.6
4-7.5
4.5-7.5
Ca
10-7200
300-4000
20-4000
aKjeldsen et a
., 2002; bPokhrel, 2004; cReinhart and Al-Yousfi, 1996;
dPohland and Kim, 1999

30
Back Front

31
CCA treated wood'
1%
Alumium
4%
Galvanized steel
4%
Plastic
15%
SYP
5%
CRT glass Mixed cullet
1% 6%
Paper office paper
27%
aper newsprint
6%
Food waste
15%
Paper cardboard
16%
Figure 2-2. The composition of fabricated municipal solid waste for this research.

32
Figure 2-3. Comparison of pH between aerobic and anaerobic lysimeters versus time

COD (mg/L) COD (mg/L)
33
Days
Figure 2-4. Changes in COD of aerobic and anaerobic lysimeters versus time

BOD (mg/L) BOD (mg/L)
34
Days
Figure 2-5. Changes in BOD of aerobic and anaerobic lysimeters versus time

30000
25000
20000
15000
10000
5000
0
30000
25000
20000
15000
10000
5000
0
35
200
400
Days
(A)
600
800

Volatile fatty acids (mg/L) Volatile fatty acids (mg/L)
36
30000
25000
20000
15000 -
10000 -
5000 -
lys 1 (aerobic)

Acetic acids
o
Propionic acids
--T
Butyric acids
-fW
50 100 150 200 250 300 350 400
(B)
Figure 2-6. Changes in VFAs of aerobic and anaerobic lysimeters versus time (A) acetic
acid only and (B) acetic acid, propionic acid and butyric acid

BOD/COD
37
Figure 2-7. Changes in the ratio BOD/COD of the aerobic and anaerobic lysimeters over
time

Ammonia (mg/L) Ammonia (mg/L)
38
Figure 2-8. Changes in ammonia concentrations versus time

60
50
40
30
20
10
0
60
50 -
40
30
20
10
0
re 2
>>
39
AEROBIC
OQ
100
200
300
400
i
100
I
200
i
600
300
400
Days
500
700
800
h Changes in TDS of the aerobic and anaerobic lysimeters versus time

20000
15000
10000
5000
0
20000
15000
10000
5000
0
e 2-10.
40
100
200
300
400
200
400
Days
600
800
hanges in alkalinity of the aerobic and anaerobic lysimeters versus time

Sulfide (mg/L)
41
0 100 200 300 400
Days
Figure 2-11. Changes in sulfide and pH versus time
600
800

Sulfate (mg/L) Sulfate (mg/L)
42
a
-o
3
C/3
10
- 8
- 6
L 0
Days
}
u
T3
3
cn
r 10
- 8
- 6
- 4
- 2
L 0
Days
Figure 2-12. The changes in sulfate and sulfide versus time in the presence of dissolved
oxygen
Dissolved oxygen (mg/L) X Dissolved oxygen (mg/L) -

43
Days
Figure 2-13. The changes in air injection rate and gas concentrations of aerobic lysimeter
Air injection rate (mL/min)

Gas concentrations (%)
44
Days
Figure 2-14. Changes in gas concentrations of anaerobic lysimeter 4

Cumulative Biogas (CH4 and C02), L
45
Figure 2-15. Cumulative biogas vs. days in aerobic and anaerobic lysimeters
800

60
50
40
30
20
10
0
7.0
6.5
6.0
46
Days
-16. Changes in gas concentrations, pH and gas generation rate after air injection
into lysimeter 3

CHAPTER 3
THE FATE OF HEAVY METALS IN SIMULATED LANDFILL BIOREACTORS
UNDER AEROBIC AND ANAEROBIC CONDITIONS
3.1 Introduction
An issue of current debate in the solid waste community is the fate of heavy metals
disposed in MSW landfills (SWANA, 2003). Heavy metals may be present as a result of
industrial residuals, but more importantly for MSW landfills, they result from
manufactured products. Examples include lead from electronic devices and copper,
chromate and arsenic from treated wood. This debate has taken on more immediate
concern as several US states have banned certain wastes containing heavy metals (e.g.,
leaded cathode ray tubes) from disposal in landfills (SWANA, 2003). These bans are in
part a result of fears regarding the fate of the disposal of metals in landfills.
For the most part, heavy metals have been thought to be relatively well contained in
typical anaerobic landfills. According to Kjeldsen et al. (2002), the amount of heavy
metals dissolved and contained in leachate is very low relative to those present in the
waste. Most metals are thought to be released during the initial stage of landfill
decomposition as a result of the lower pH. Once a landfill enters the methanogenic phase,
heavy metal concentrations in leachate dramatically decrease, and in many cases, their
levels decrease to lower than drinking water standards (Kjeldsen et al., 2002).
Bioreactor landfills are becoming a more common method of managing MSW, and the
impact of these systems on metal leachability should be evaluated. Since bioreactor
landfills involve exposing a much larger percentage of waste to moisture, the total mass
47

48
of metals released might be expected to be high relative to dry landfills. On the other
hand, given that bioreactor landfills recirculate leachate to the landfill and that traditional
bioreactors promote anaerobic waste decomposition (and thus enhance metal removal by
the mechanisms described previously), the impact to the environment may be limited.
An alternative bioreactor landfill technique is to add air in addition to moisture.
Taking into consideration that the leaching behavior of heavy metals is mainly controlled
by redox, pH and the presence of ligands (Benjamin, 2002), it may be that the fate of
heavy metals in aerobic systems will differ from anaerobic systems. The long term fate of
heavy metals in aerated landfills is a question yet to be satisfactorily addressed.
In this research, the fate of heavy metals in simulated aerobic and anaerobic
bioreactor landfills was studied. Four stainless-steel lysimeters were used, two each for
aerobic and anaerobic conditions. Heavy metal-containing wastes were included in the
waste stream added to each lysimeter. Leachate collected from each lysimeter was
analyzed for heavy metals over time. In order to evaluate the heavy metals adsorbed in
solid wastes, waste samples were removed from two of the lysimeters at the end of the
experimental period and analyzed for heavy metal content.
3.2 Materials and Methods
A detailed description of the lysimeters was presented in chapter 2. The methods as
described here focus on the metal-containing components in the fabricated waste and on
metal analysis in the leachate and waste samples.
3.2.1 Heavy Metal Sources in Synthetic Waste
Several MSW components were chosen as sources of heavy metals. Each
component, its corresponding heavy metals and the percentage of each component are
presented in Table 3-1. Aluminum and galvanized steel sheets were purchased from a

49
local hardware store and cut into 1.5 cm 1.5 cm square. Galvanized steel served as a
source of both Fe and Zn. CCA-treated wood was used as a source of Cr, Cu and As.
Crushed cathode ray tube (CRT) monitor glass was used as a Pb source. Total Cu, Cr,
and As concentrations were 2350 50, 2890 56, and 1330 10 mg/kg, respectively.
Crushed CRT monitor glass used in this research was a mixture of the funnel sections of
30 CRT color monitors. Jang and Townsend (2003) reported that 413 mg/L of Pb leached
from CRT funnel glass using the toxicity characteristics leaching procedure (TCLP).
3.2.2 Sampling Methods
Leachate samples were collected weekly via a sampling port located at the bottom
of each lysimeter. A portion of the leachate collected was used for analysis of general
water quality parameters and the remainder was injected back into the lysimeters. A 50
mL aliquot was preserved with concentrated nitric acid and used for heavy metal
analysis.
Lysimeter studies were conducted for 379 and 741 days for aerobic and anaerobic
lysimeters, respectively. After the lysimeter studies were completed, solid wastes were
removed from single aerobic and anaerobic lysimeter and analyzed for heavy metals. The
samples were divided by depth into 4 fractions. Details about these fractions are
summarized in Table A-2 in appendix A. Each fraction was then separated into 5
categories which include office paper, cardboard, newspaper, wood blocks and plastics.
The separated samples were ground using an Urschell mill (Fritsch, German).
3.2.3 Analytical Methods
Leachate samples were digested with nitric and hydrochloric acids following EPA
method 3050B and 3010A for solid and liquid digestion, respectively (USEPA, 2003).
Approximately 2 g of the ground samples were digested using nitric acid and 30%

50
hydrogen peroxide and then analyzed for heavy metals using ICP-AES following US
EPA, SW-846 Method 601 OB (USEPA, 2003). Digested samples were filtered using ash
free cellulose filters and analyzed for heavy metals and cations using Inductively Coupled
Plasma-Atomic Emission Spectrometry (ICP-AES) (Thermo Electronics, USA). Leachate
samples preserved with concentrated nitric acid were analyzed for a total of 8 metals (As,
Cu, Cr, Mn, Zn, Pb, Fe and Al).
3.3 Results and Discussions
3.3.1 Changes in Metal Concentrations versus Time and the Percentage of Mass
Loss
The following section presents the results (for each metal) of the aerobic and
anaerobic lysimeters as separated plots. The experimental time period for the aerobic and
anaerobic lysimeters differed. Because of their time scale difference, the cumulatitive
mass of metal leaching was plotted for all lysimeters as a function of waste mass loss.
The estimation of mass loss is described in appendix A. The total amounts of leachate
produced and used for the analysis are summarized in Table 3-3.
3.3.1.1 Aluminum
The changes in Al concentration in aerobic and anaerobic conditions over a period
of time are depicted in Figure 3-1. High Al concentrations were observed from both
aerobic and anaerobic lysimeters (18 and 20mg/L) for the first 10 to 20 days. Whereas the
Al concentrations of anaerobic lysimeters dramatically decreased to below 0.5mg/L
within 100 days, great changes in Al concentrations were not observed from the aerobic
lysimeters. The changes in Al concentrations in aerobic lysimeters were mainly
controlled by pH; high concentrations of Al were observed from both lysimeters 1 and 2
at pH < 6 and pH > 8, and lowest Al concentrations (< 1 mg/L) were observed at 6 < pH

51
< 8. Average Al concentrtaions (7.9 mg/L) of the aerobic lysimeters were significantly
higher than those of the anaerobic lysimeters (0.28 mg/L).
Generally, A1 leaching is not greatly affected by redox conditions but mainly
controlled by pH. Meima and Comans (1997) reported that A1 solubility was low in the
pH range 6 to 7, which corresponds to the A1 results in the aerobic condition presented in
Figure 3-1. Among many ligands forming A1 compelxation, hydroxide ion (OH ) is
known as a major ion to control the solubility of A1 in aquatic systems. The equilibrium
of A1 with gibbsite (Al(OH)3)), an Al-OH complex, is characterized by a U-shaped pH-
leaching curve (Eary, 1999). However, A1 concentrations in anaerobic lysimeters did not
appear to follow with the solubility of gibbsite. Besides pH and redox conditions, large
differences in leachate characterisitcs between aerobic and anaerobic lysimeters included
the high organic content and anions such as floride and sulfate in the leachate of the
anaerobic lysimeters. The most likely explanation of low A1 solubility in the anaerobic
conditions is complexation of A1 and organic matter. Tipping (2005) reported that A1
solubility is strongly associated with organic content. Skyllberg (2001) also reported that
high dissolved organic content made A1 solubility significantly decrease. Therefore, low
solubilitity of A1 during the first phase of the anaerobic lysimeters is the result of
complexation of A1 with high concentrations of organic matter.
3.3.1.2 Arsenic
Figure 3-2 depicts the change in As over time. High As concentrations were
observed from the anaerobic lysimeter before day 220. The greatest As concentration was
3.2 mg/L on the day 89. The As concentration then lowered below 1.5 mg/L after day
400. For lysimeter 4, As concentrations were continuously low, showing the lowest value,
0.27 mg/L on the last sampling day (day 741). In contrast to the anaerobic lysimeters, As

52
concentrations in the aerobic lysimeters decreased initially and increased with increasing
pH. The As leaching pattern of the aerobic lysimeters appears similar to A1 leachate
trends. The lowest As concentration of lysimeter 1 was 0.12 mg/L on the day 163. For
lysimeter 2, extremely low As concentrations were observed, with several samples below
the detection limit (0.011 mg/L) despite a pH < 6. Overall, As dissolved in the leachate of
the aerobic lysimeters was significantly lower than that of the anaerobic lysimeters (p <
0.05).
Figure 3-9 depicts the distribution of As concentrations observed from the aerobic
and anaerobic lysimeters at various pH conditions. It has been reported that As solubility
changes with pH and is characterized by a U-shaped curve in oxidizing conditions
(Drever, 1988). However, Carbonell-Barrachina et al (1999) reported that in the presence
of sulfide, Fe, and Mn, the solubility of As was dramatically lower and did not follow a
U-shaped solubility curve. However, As concentrations were not impacted by these
constituents in the anaerobic lysimeters. The most likely explantion is that the anaerobic
lysimeters had poor-anoxic conditions during the first phase. The low sulfide
concentrations at a pH < 6 confirm that the redox potential was not low enough for
sulfide to become involved in As precipitation. For the aerobic lysimeters, As
concentrations were low under acidic conditions and increased up to 1 mg/L at a pH of 9.
Masscheleyn et al (1991) found that As solubility decreased substantially as the
redox potential increased. The changes in As solubility are also associated with the
oxidation state of iron; Fe (III) has a strong affinity for arsenate. Therefore, low arsenic
concentrations are likely dictated by the low solubility of arsenate. Under oxidizing
conditions, As solubility may increase or decrease by pH changes. This is because of the

53
effect of pH on the total oxyaionic arsenate concentrations by pH. This changes in
solubility results in the relatively higher concentrations of As at alkaline conditions
observed in the aerobic lysimeters (Figure 3-2).
3.3.1.3 Chromium
The initial Cr concentrations of the anaerobic lysimeters were higher than those of
the aerobic lysimeters (Figure 3-3). However, Cr concentrations in the anaerobic
lysimeters gradually decreased to below 0.05mg/L by the day 453. As the pH of lysimeter
4 changed to moderately alkali (pH > 7.4) after day 464, minor increases in Cr
concentrations were observed. In contrast, clear Cr leaching trends were not exhibited by
the aerobic lysimeters before day 100, but an increase in Cr concentration did occur
following day 150. This increase in Cr concentration corresponds to an increase in pH.
Overall, the average Cr concentrations of the aerobic lysimeters were significantly greater
than those of the anaerobic lysimeters. This may be because thermodynamically Cr can
be present as an ionic form at alkaline pH under oxidized condition.
The toxicity of Cr is determined by its oxidation state. Among the various Cr
oxidation states, only trivalent and hexavalent forms are taken into consideration in
natural aquatic systems. Hexavalent Cr is considered more toxic than trivalent Cr due to
its high mobility and solubility. Cr (VI) may be reduced to Cr (III) at low ORP potential.
Cr (VI) becomes unstable and is reduced to Cr (III) at low pH. In order to maintain the
oxidation state of Cr as Cr (VI) at a low pH, it is necessary to keep highly oxidizing
conditions (Richard and Bourg, 1991). In contrast to other metals such as As and Cu, Cr
(III) is not likely to precipitate with sulfide. Chromium solubility is mainly controlled by
Cr(OH)3(s). Generally, Cr(OH)3(s) is formed in a pH range of 6.5 to 7 under moderately
oxidizing or reducing conditions.

54
According to the potential-pH diagram of Cr (Figure 3-10), total Cr obtained from
both aerobic and anaerobic lysimeters in an acidic environments is likely to be Cr (III) as
Cr(OH)2+. In contrast, dissolved Cr from the aerobic lysimeter at a pH 9 could be Cr (VI)
as CrC>42\ Since all Cr species presented on the potential-pH diagram are based upon
assuming thermodynamic equilibrium, all Cr obtained from the aerobic lysimeters at high
pH may not be Cr (VI).
It is noted that an increase in Cr was observed from lysimeter 4 around a neutral pH
(A in Figure 3-3). The most likely explanation for this is the lower Fe concentrations of
lysimeter 4 than of those of lysimeter 3 (Figure 3-6). In the presence of Fe, Cr may be
precipitated with Fe rather than OH' due to rapid kinetics. The complexation of Fe and Cr
decreases Cr solubility lower than the complexation of OH and Cr (Eary and Rai, 1987).
Therefore, an increase of Cr concentration at the end of lysimeter 4 would be the result of
a decrease of Fe concentrations.
3.3.1.4 Copper
Overall copper concentrations of the aerobic lysimeters were one to three orders of
magnitude higher than those of the anaerobic lysimeters (Figure 3-4). For the aerobic
lysimeters, clear Cu leaching patterns over time were not observed, but relatively large
changes in Cu concentrations were observed at lysimeter 1 from the day 140 to 190. This
period of time corresponded to a pH change from 5.5 to 9. For anaerobic lysimeters, Cu
concentrations gradually decreased for the first 450 days. The initial Cu concentrations of
lysimeter 3 and 4 were 0.082 and 0.234 mg/L, respectively. Although the concentrations
slightly increased after day 450, final concentrations of Cu remained lower than the initial
values.

55
Copper solubility is controlled by several Cu-containing minerals forming
complexation with Fe and sulfide. In addition, Cu sulfides may coexist with the sulfides
of other metals such as Zn, Pb and As (Faure, 1991). Representative mineral deposits
formed with OH', Fe and/or sulfide include chalcocite (CU2S), chalcopyrite (CuFeS2),
cuprite (Cu20) and malachite (Cu2(0H)2C03). These minerals are widely distributed over
a pe-pH diagram. Ionic Cu is present only at a pH less than neutral and under highly
oxidizing conditions (pe > 2.5). For these reasons, high concentrations of Cu may not be
found under landfill conditions where low ORP and neutral pH are predominant. These
concepts can be applied to the Cu distribution patterns displayed at a low pH in the pH-
Cu concentration chart shown in Figure 3-11.
However, there is a disparity in the Cu concentrations observed and those
thermodynamically predicted for the alkali conditions of the aerobic lysimeters; most of
the Cu precipitated by complexation with various Cu-containing mineral deposits at
alkali and oxidizing conditions. Edwards et al (2000) reported high Cu concentrations in
drinking water at alkali conditions, calling it the blue color phenomenon since water
color changed to blue with high concentrations of Cu. Critchley et al (2004) explained
this blue water was caused by microorganism-intermediated-Cu leaching from a part of
the water delivery system. In this research, blue water was also observed from
condensate passing through a copper tube which connected to a gas collection system of
an aerobic lysimeter. Further development of Cu corrosion caused a small hole on the
same copper tube and called for a replacement of the copper tube with plastic materials
(Figure B-7).

56
Another possibility of copper leaching from the aerobic lysimeters is the binding of
Cu with ammonium (NHj+). Since both Cu and ammonium are cations, their
complexations are present as an ionic form and can be dissolved in aquatic systems.
Arzutug et al (2004) reported that Cu leaching increased with ammonia concentrations.
However, complexation of Cu and ammonia may occur in a relatively narrow range of
ORP and pH (Hoar and Rothwell, 1970). When plotting Cu and ammonia data obtained
from the aerobic lysimeters, no clear evidence to prove the relationship between Cu
concentrations and ammonia was found (r2 = 0.021). Furthermore, since Cu complexes
with sulfide rather than ammonia in the presence of sulfide (Alymore, 2001), it may be
difficult to leach high concentrations of Cu under landfill conditions.
3.3.1.5 Lead
Figure 3-5 depicts the changes in lead concentrations over time. For aerobic
lysimeters, Pb concentrations dramatically increased to 1.7-2 mg/L within 30 days and
then gradually decreased. After the pH of the aerobic lysimeters stabilized, Pb
concentrations decreased to levels similar to the anaerobic lysimeters. In contrast, little
change in Pb concentrations was observed from the anaerobic lysimeters and low
concentrations of Pb were maintained over the test period. Generally, Pb concentrations
in landfill leachate have been reported to be very low (Charlatchka and Cambier, 2000;
Jang and Townsend, 2003). This is because lead may precipitate with various ligands
such as carbonate ions (CO3 ), sulfide, and volatile fatty acids (VFA).
Lead solubility is generally controlled by carbonate, or other Pb hydroxides and
phosphate in noncalcareous soils (Bradle, 2005). Charlatchka and Cambier (2000)
concluded that Pb solubility increased under oxidizing conditions at a pH of 6.2
However, Pb may precipitate as a form of PbS under reducing conditions in the presence

57
of sulfur. Lead is generally present in an ionic form at a pH < 6 under oxidizing
conditions (Drever, 1988).
As shown in Figure 3-5, Pb leached from the aerobic lysimeters significantly
greater than from the anaerobic lysimeters. Most samples with high concentrations were
distributed in the acidic phases. This leaching pattern corresponds to the characteristics of
Pb previously discussed. For both aerobic and anaerobic lysimeters, Pb concentrations
decreased with an increase in pH. Most Pb in alkali conditions may be precipitated as
forms of PbCC>3 or PbS depending upon redox potentials.
3.3.1.6 Iron
As shown on Figure 3-6, initial Fe concentrations of the aerobic lysimeters
(110mg/L for both aerobic lysimeters) were higher than those of the anaerobic lysimeters
(20-22mg/L). Iron concentrations of the aerobic lysimeters increased to 250mg/L on the
30th day and then gradually decreased. During changes in pH of the aerobic lysimeters to
alkali conditions, Fe concentrations lowered substantially to below 10mg/L. In contrast to
the aerobic lysimeters, Fe concentrations of the anaerobic lysimeters increased from 20 to
600mg/L for the first 450 days and then decreased with increasing pH. Although
lysimeters 3 and 4 are both anaerobic lysimeters, the final Fe concentrations were
substantially different (165mg/L and 5.6 mg/L for lysimeters 3 and 4, respectively).
Generally, free Fe concentration is strongly associated with the redox condition of
the system. Iron is present in aquatic systems in two oxidation states; Fe (III) and Fe (II).
Ferric (Fe3+) and ferrous (Fe2+) irons can be transformed to each other depending upon
the redox conditions. Iron (III) is precipitated as a mineral deposit such as Fe23 or
Fe(OH)3 at a pH > 5. Iron (III) is also involved in complexation with metals. Under
moderately oxidizing and reducing condition, Fe (II) ions are dominant in the pH range

58
of 5 to 9. In the presence of sulfur, Fe (II) is likely to be precipitated as pyrite (FeS2) at
pH > 5 under reducing conditions (Drever, 1988). Despite the many other reactions Fe is
involved with, Fe solubility is strongly controlled by sulfide concentrations. Thus, in
order to lower Fe concentration, the redox potential of the system needs to be low enough
to reduce sulfate to sulfide. Sulfide concentrations of lysimeter 3 were lower than those
of the other lysimeters, resulting in greater concentrations of Fe observed from lysimeter
3.
3.3.1.7 Manganese and Zinc
Leaching patterns of Mn and Zn look very similar for the aerobic and anaerobic
lysimeters (Figures 3-7 and 3-8) and are thus discussed together. Relatively high
concentrations of Mn and Zn were exhibited from the aerobic lysimeters for the first 150
days. The highest concentrations of Mn and Zn were 11 mg/L and 270 mg/L,
respectively. Manganese and zinc concentrations then substantially decreased to below
0.2 mg/L and 10 mg/L, respectively. In contrast, little change in Mn and Zn
concentrations was observed from the anaerobic lysimeters for 450 days, but decreased
following that period. This corresponds to when the pH of the anaerobic lysimeters
increased.
Since Mn and Zn are mainly precipitated by sulfide, differences of Mn and Zn
concentrations between lysimeters 3 and 4 are strongly associated with the sulfide
concentrations present in each lysimeter. Additionally, the solubility of Zn and Mn is
associated with organic matter. A decrease in Zn and Mn solubility accompanies an
increase of pH and could be accounted for by the generation of pH-dependent charge
sites on organic matter (McBride and Blasiak, 1979; Miyazawa et al, 1993).

59
3.3.2 Organic Wastes as Absorbents of Heavy Metals
Analytical results of the 8 metals absorbed on office paper (OP), cardboard (CB),
newspaper (NP), and wood blocks (WD) are shown on Figure 3-12. Concentrations of Al,
As, and Cu adsorbed on the lignocellulosic materials of the aerobic lysimeter appeared
greater than those of the anaerobic lysimeter.
Heavy metals adsorbed on solid wastes from the aerobic and anaerobic lysimeter
were statistically analyzed using the ANOVA test. Test results are presented in appendix
C. All metals, except for Pb, absorbed on CB in the aerobic lysimeter and were
significantly higher than those of the anaerobic lysimeter. Other significant differences
between the aerobic and anaerobic lysimeters were found with Al, As, Mn and Cu
adsorbed on NP, OP and WD.
Figure 3-13 depicts the differences of total mass of metals adsorbed on
lignocellulosic materials between the aerobic and anaerobic lysimeters. These values
were calculated by multiplying the metal concentrations by the mass of each waste
obtained from the garbage separation. Interestingly, the observed trends of adsorption of
some metals did not to correspond with their leaching trends. These trends can be found
from adsorption trends of As, Mn and Pb. The amounts of metals leached and adsorbed
between aerobic and anaerobic lysimeters are compared in Table 3-5. These results
indicate that metal adsorption may be influenced by environmental conditions such as
pH, redox, and the presence of other ligands. Ravat et al (2000) reported that the binding
of selected metals (Zn, Cu, and Pb) on lignocellulosic materials is strongly pH-dependent
in the absence of interference from other ligands. Adsorption of Fe on organic matter is
mainly controlled by the oxidation states of Fe; Fe (III) has greater affinity for organic
matter than Fe (II) does (Jansen et al., 2003), corresponding to the large differences of Fe

60
adsorbed between the aerobic and anaerobic lysimeters (Table 3-6). Adsorption of As
also mainly occurs when As is oxidized. As the pH rises, ionic forms of As changes
progressively (H2ASO4', HAs042' and As043') with each species showing different
adsorption properties (Drever, 1988).
It is noted that there were large differences between metals released through
leachate and metals that remained in the lysimeters by adsorption. These differences can
be numerically expressed using the ratios between metal leached (LC) and adsorbed
(AD). The smallest LC/AD ratio can be found from Al; only 0.06% of the amount of A1
adsorbed was released from the lysimeters. The LC/AD ratios of most metals fell into
around or below 2%. Relatively high LC/AD ratios were exhibited from a few metals in
the anaerobic lysimeter; LC/AD ratios of As, Mn and Zn were 13.8%, 16.5% and 8.6%,
respectively. However, if the amounts of metals precipitated as particulate forms without
adsorption are taken into consideration, the ratio of metals between leached and remained
would be much smaller than LC/AD ratios.
Lignocellulosic materials such as paper and wood products occupy as much as 45%
of MSW landfills (USEPA, 2003). Cellulose and lignin are reported as the major heavy
metal adsorbents (Basso et al, 2004). Lignin especially provides many chemical
functional groups such as carboxyl and phenolic groups. Babel and Kumiawan (2003)
concluded that lignin was considered as the best low-cost adsorbent for Pb and Zn. Basso
et al. (2002) also reported that maximum sorption capacity increases due to lignin
contents during Cd sorption research. Cellulose has also been heavily demonstrated to
remove heavy metals such as Cd, Cu, Ni, Zn and Pb (Sublet, 2003; Okieimen et al., 2005
and Shukla and Pai, 2005).

61
Figure 3-14 depicts metal concentrations adsorbed on selected lignocellulosic
materials (newspaper and cardboard) and plastic waste. Greater concentrations of most
metals except for Pb and Zn adsorbed on lignocellulosic materials were observed in the
aerobic lysimeter. Bradle (2005) explained that many metals tend to adsorb the organic
matter as the pH increases under oxidizing condition. Zhang and Itoh (2003) reported that
carbonized mixture of polyethylene terephthalate (PTE) and waste ash could be used as a
metal sorbent. However in this research, metal concentrations adsorbed on plastic waste
were substantially low in comparison with metal concentrations adsorbed on the
lignocellulosic materials.
3.4 Discussion
Among the various factors affecting heavy metal leaching under landfill conditions,
the redox and pH may play the most critical role. Under the given redox condition and
pH, the metal oxidation state, ligands, adsorption behavior can be determined. In many
cases, metal precipitation can be controlled by Fe (II) and sulfide concentrations in both
aerobic and anaerobic condition. Cr, Cu, Pb, Zn and As are reported to adsorb on hydrous
ferric oxide at pH > 6, and the precipitation of Cu, Fe, Pb, Mn and Zn is controlled by
sulfide in anaerobic condition. In addition to those ligands, hydroxide ion (OH ) also can
play an important role to precipitate A1 and Cr (Drever, 1988). The various chemical
interactions are depicted in Figure 3-15.
3.4.1 Overall Comparison of Metal Behavior
Figure 3-16 describes the overall trends of metal leaching in aerobic and anaerobic
lysimeters. Among 8 metals under consideration, (Al, As, Cr, Cu, Pb, Mn, Fe and Zn)
greater concentrations of Al, Cr, Cu and Pb were observed in the leachate of the aerobic
lysimeters, and As, Mn, Fe and Zn were observed in the leachate of the anaerobic

62
lysimeters. For the anaerobic lysimeters, the metal leaching occurred in the acid phase,
while occurrence of the metal leaching was relatively well distributed over the pH (5 <
pH < 9) for the aerobic lysimeters except for Pb.
Cumulative masses of metals were calculated by the multiplication of the amount
of leachate used for analysis by the concentration of the metals in the leachate sample
(Figure 3-17). The total amounts of leachate produced and used for the analysis are
summarized in Table 3-3.
Among the 8 metals under consideration, As, Fe, Pb, Mn and Zn increased
substantially for the first 10 to 15% of mass loss and reached a plateau. These leaching
patterns indicate that metal leaching mainly occurred at the initial phase, an acidic
environment, in both aerobic and anaerobic conditions. In contrast to these metals, only
minor change in cumulative masses of A1 and Cu was observed from the anaerobic
lysimeters whereas a consistent increase in these metals was exhibited from the aerobic
lysimeters. Relatively high concentrations of A1 were exhibited during the initial stage of
the anaerobic lysimeters, but no further changes were observed. It is notable that
cumulative concentrations of A1 and Cr increased more rapidly after 25% mass loss
occurred. An increase in the rate of accumulation of these metals corresponds to an
increase in pH to 9.
Overall metal leaching behavior is strongly associated with pH and redox
conditions. Since the anaerobic lysimeters remained in the acidic condition (pH < 6) for
more than 400 days, great amounts of metals such as As, Mn, Fe, Cr and Zn were
released through the leachate. In contrast to the anaerobic lysimeters, greater cumulative
concentrations of Al, Cu and Pb were observed from the aerobic lysimeters. Leaching of

63
these metals might be influenced by the different environment of the aerobic lysimeters
such as an alkaline pH and oxidizing conditions.
3.4.2 Comparison to Other Studies
Generally, heavy metal concentrations found in anaerobic landfills are reported low
(Kjeldson et al., 2002). The presence of sulfide and the low solubility of metals at neutral
pH may reduce metal concentrations in leachate. Metal concentrations of the aerobic and
anaerobic lysimeters along with MSW leachate summarized from the literature are
presented in Table 3-6. Metal concentrations of the aerobic lysimeters listed in Table 3-6
are the average value of the lysimeters 1 and 2. For the anaerobic lysimeters, since
lysimeter 3 remained in acidic condition for most of a test period, metal results of
lysimeter 3 during the methanogenic phase were not included in Table 3-6. For the
anaerobic lysimeters, As and Zn concentrations were substantially higher than those of
MSW leachate during acidic phase. They were reduced then during the methanogenic
phase and similar to those of MSW leachate despite the presence of CCA-treated wood.
Cu concentrations of the anaerobic lysimeters were extremely low, and they were lower
than even drinking water standards. Although most metal concentrations of the anaerobic
lysimeters were greater than drinking water standards, they were similar or lower than
those of general MSW leachate during the methanogenic phase.
For the aerobic lysimeters, As, Cr, Fe, Mn and Zn concentrations were lower than
those of MSW leachate and the anaerobic lysimeters during the first acidic phase.
Aluminum and copper concentrations were greater than those of the anaerobic lysimeters.
Only Pb concentration was greater than that of MSW landfill leachate and the anaerobic
lysimeters. However, different aspects of metal leaching were observed from the aerobic

64
lysimeters during the alkaline phase; the concentrations of most metals except for Fe
were greater than those of the anaerobic lysimeters.
The leaching behavior of CCA-treated wood of the aerobic and anaerobic
lysimeters was compared to a similar study (Jambeck, 2004). Jambeck (2004) researched
the leaching behavior of CCA-treated wood mixed with MSW through the 6.7-m high
PVC column tests. A total 2% of CCA-treated wood was included in the column and rain
water was used for Jambecks study while 1% of CCA-treated wood and DI water were
used for the present study (Table 3-7). In comparison metal leaching results from
Jambecks study proved similar to the anaerobic lysimeter here (Figure 3-18). Extremely
low Cu concentrations were observed in both studies. The range of Cr concentrations of
Jambecks study was higher than that of the anaerobic lysimeters during acid phase, but
the median of Cr concentrations of Jambecks study was bottom of the range. In contrast
to anaerobic condition, significantly different leaching trends of As and Cu were
exhibited from the aerobic lysimeters; overall As concentrations of the aerobic lysimeter
were lower than those of the anaerobic column studies during the acid phase. The 95th
percentile of As concentrations of the aerobic lysimeter was in the range of the anaerobic
system, but they were detected at the very beginning of the aerobic lysimeter operation.
After pH stabilized, the median As concentrations of both the aerobic and anaerobic
systems became identical. However, Cu concentrations were two orders of magnitude
higher than those of anaerobic lysimeters for both the acid and methane (alkaline) phase.
In comparison with other column study (Jambeck, 2004), it can be concluded that
leached As, Cu and Cr concentrations might not always be followed by the initial mass of
CCA-treated wood and metal concentrations contained. The differences between As and

65
Cu concentrations between the anaerobic and aerobic columns proved too high to be
comparable. Within the same anaerobic systems, As and Cr concentrations between
Jambeck and this study were not much different despite different initial As and Cr
masses. Similar leaching trends could be observed in the same anaerobic or aerobic
system, but overall concentrations of As and Cr per initial mass may depend more on the
chemistry of the system.
3.4.3 Implication for Disposal of Heavy Metals
In this research, CCA-treated wood and CRT monitor glass were used as metal
sources to represent treated wood and electronic waste. Though the land-disposal of these
wastes was banned in several states in the U.S. (SWANA, 2003), they can still be land-
disposed as a form of home appliances and ash. Since the greatest amount of As may be
leached during the first acid phase of anaerobic landfills, landfill owners need to monitor
the leachate quality during landfill construction. Though thermodynamically As may
precipitate with sulfide, a maximum of 70% of As dissolved in solution may combine
with sulfide at pH 8 (Carbonell-Barrachina, 1999). This adsorption ratio decreased as the
pH decreased. In addition to the sulfide, Fe, organic matters, and carbonate are also
known as As adsorbent. However, the adsorption efficiency of those ligands were low
relative to sulfide (Carbonell-Barrachina 1999). Thus, in the presence of an As source,
As can be found in landfill leachate for all operation periods. For aerobic landfills, As
concentrations may slightly increase during the alkaline phase. Overall As concentrations
found in aerobic landfills may be lower than those of anaerobic landfills.
Thermodynamically, Cr concentration in landfill condition during all phases may
be low. During the first acid phase, Cr may be combined with high concentrations of Fe
(II), and Cr may precipitate with sulfide during the methane phase. The lower

66
concentration of Fe (II) may increase the Cr solubility, but the change is very small. Cu
concentrations in anaerobic landfills are typically extremely low. However, high
concentrations of Cu may be found in both acid and alkaline phase of aerobic landfills.
Since the microorganisms which contribute Cu leaching may corrode all Cu-made
equipment connected to the landfill, it is important to avoid using any Cu-containing
equipment for gas and leachate collection systems. Pb has high solubility under oxidizing
conditions at pH < 6. Thus it would be recommendable to monitor leachate quality for the
first acid phase of aerobic landfills. However, air addition facilities are generally installed
after landfill closure, Pb concentrations may not be high over operation period. Pb
concentration in anaerobic landfill conditions is generally low.
As previously discussed, air addition into a current anaerobic bioreactor landfill
may enhance waste decomposition substantially. However, unlike anaerobic landfills,
concerns about high A1 and Cu concentration and a risk of Cr (VI) may arise. In order to
avoid these potential risks, it would be recommendable to inject air into the shallow well
rather than the deep well. Since Al, Cu and Cr are redox-sensitive, great amount of these
metals could be reduced after passing through the anaerobic zone. However, since air
injection into the shallow wells may reduce the air diffusion efficiency into a landfill,
further economical and efficiency of air distribution analysis are needed.
3.4.4 The Impact of Air on Metal Mobility
Although total amounts of metals leached were not considerably high, it is
necessary to pay great attention to certain metals due to changes of their toxicity by
different pH and redox conditions. For example, among Cr species dissolved in leachate,
Cr (III) can be dominant in current anaerobic sanitary landfills, however,
thermodynamically Cr (VI) becomes a major Cr component in the environment formed

67
upon air intrusion. Hexavalent chromium is a highly toxic metal causing decreased
pulmonary function and pneumonia (Bradle, 2005). On the contrary, toxicity of As can be
reduced by air injection. In oxidizing conditions, Arsenate can be oxidized to As (VI),
which is less toxic. Since As (VI) is less mobile and has a low solubility, the total amount
of As leached can be reduced. The influence of air injection on the amount of metals can
be more clearly understood by comparing the percentage of As, Cr and Cu (Table 3-4).
Under the scenario of high As content in leachate caused by co-disposal of fly ash or
ground CCA-treated wood, it would be recommended to inject air in order to reduce As
toxicity and the amount leached. However, Cr (VI) in an aerobic landfill can be present
as an ionic form under oxidizing conditions, and the cumulative mass may increase over
a period of time. For these reasons, it is necessary that landfill personnel develop a
strategy to optimize the influence of air injection.
Based on the metal results, the metal leaching behavior of old landfill metals can be
predictable. Kjeldsen et al. (2002) proposed that the condition of old landfills would be
aerobic due to air intrusion. Once air intruded in anaerobic landfills, metal leaching may
occur by the dissociation of the metals that adsorbed on decomposed organic matters.
Among 8 metals under consideration, Pb would be the most leachable metal. As Figure 3-
13 shows a large amount of Pb was adsorbed on organic matter. Bozkurt et al. (1999)
explained that the pH of the landfill would change to acid again after the long-term
process. This was due to the oxidation of sulfate and organic matter. Therefore, the
moderate oxidizing condition, and low pH, would accelerate Pb leaching.
3.5 Conclusions
Research on the fate of metals in simulated aerobic and anaerobic landfills was
conducted. The leaching behavior of selected metals was significantly different between

68
aerobic and anaerobic lysimeters. Among the 8 metals evaluated (Al, As, Cr, Cu, Pb, Mn,
Fe and Zn), the concentrations of Al, Cu, Cr and Pb in leachate of the aerobic lysimeters
were significantly greater than those of the anaerobic lysimeters, and the average
concentrations of As, Fe, Mn, and Zn in the anaerobic lysimeters were significantly
greater in concentration than observed in the anaerobic lysimeters.
After a test period, one each of the aerobic and anaerobic lysimeters was
dismantled and the metals adsorbed on decomposed lignocellulosic waste were analyzed.
Greater concentrations (mg metal/kg waste) of Fe, Mn, As, Al and Cu were found from
the aerobic lysimeters. All metals, except for Pb, adsorbed on the cardboard in the
aerobic lysimeter and were significantly higher than those of the anaerobic lysimeter.
Through this research, aerobic landfills were found to have a greater potential to release
several metals through leachate than that of anaerobic landfills; aerobic landfills have
greater Al, Cr, Cu and Pb leaching potential due to oxidizing condition. Experimental
results confirmed this notion. In the presence of CCA-treated wood and electronic waste
(CRT monitor glass), high As and Pb concentrations were observed from the anaerobic
lysimeter and the aerobic lysimeter during the acid phase, respectively.

69
Table 3-1.Heavy metal sources in fabricated waste Stream-
Waste components
Contained heavy metals
% of component in fabricated
waste
CCA treated wood
Copper, Chromium and
Arsenic
1%
Cathode-ray Tube (CRT) glass
Lead
1%
Aluminum sheet
Aluminum
4%
Galvanized steel sheet
Zinc, Manganese and Iron
4%
Table 3-2. Results of statistical analysis of metal leached between aerobic and anaerobic
Average concentrations (mg/L)
F
P-value
F-crit
Aerobic
Anaerobic
A1
7.89
1.28
206.89
8.3E-33
3.89
As
0.40
1.28
66.15
4. IE-14
3.89
Cr
0.19
0.10
40.67
1.19E-09
3.89
Cu
2.87
0.02
81.10
1.58E-16
3.89
Fe
35.06
167.68
53.09
6.92E-12
3.89
Mn
1.91
4.57
35.80
9.75E-09
3.89
Pb
0.22
0.03
31.32
7.07E-08
3.89
Zn
54.36
201.12
66.30
3.87E-14
3.89
Table 3-3. The amount of leachate produced and used for analysis
lys 1
lys 2
lys 3
lys 4
leachate produced (mL)
8,717
9,747
18,571
15,979
leachate released (mL)
(used for analysis)
4,024
4,081
6,135
6,111

70
Table 3-4. Leachability of As, Cr, and Cu
Initial cone, (mg/lys)
mg released
% released
aerobic
anaerobic
aerobic
anaerobic
As
1279.1
mg
Lys 1
Lys 3
1.53
8.21
0.12%
0.64%
Lys 2
Lys 4
1.17
9.30
0.09%
0.73%
Cr
1573.1
mg
Lys 1
Lys 3
0.72
0.50
0.05%
0.03%
Lys 2
Lys 4
0.56
1.24
0.04%
0.08%
Cu
723.9
mg
Lys 1
Lys 3
9.82
0.10
1.36%
0.01%
Lys 2
Lys 4
6.37
0.34
0.88%
0.05%
Table 3-5. Comparison of cumulative mass of metal dissolved in leachate and adsorbed
on lignocellulosic materials (unit: mg)
lys 1
lys 2
lys 3
lys 4
Aerobic
(lys 2)
Anaerobic
(lys 4)
LC/AD
of lys 2
LC/AD
of lys 4
A1
27.3
25.9
7.8
8.7
41900
15700
0.06%
0.06%
As
1.5
1.2
8.2
9.3
110
67.6
1.06%
13.76%
Cr
0.7
0.6
0.5
1.2
222.2
243
0.25%
0.51%
Cu
9.8
6.4
0.1
0.3
238.1
151.2
2.68%
0.22%
Fe
161.3
64.2
1200
506.4
31100
19200
0.21%
2.63%
Mn
10.5
2
31
29.3
494.2
177.3
0.40%
16.51%
Pb
0.6
0.5
0.2
0.1
47.8
156.7
1.13%
0.08%
Zn
292.9
64.2
1,100
1,100
9,400
12,500
0.69%
8.60%

71
Table 3-6. Comparison of average metal concentrations of the aerobic and anaerobic
lysimeters with MSW leachate and regulatory levels (SAIC, 2000; USEPA,
1996 and 2003; Kjeldsen et al., 2002) (unit: mg/L)
MSW
leachate
Aerobic lysimeters
Anaerobic lysimeters
TCLP TC
limits
Drinking
water
standards
Acid phase
Alkali
phase
Acid phase
Methane
phase
Al
15.05
4.56
9.07
0.31
0.17
-
0.2*
As
0.44
0.19
0.55
1.5
0.43
5
0.01
Cr
0.24
0.09
0.26
0.1
0.14
5
0.1
Cu
0.14
4.01
1.75
0.02
0.07
-
1.3
Fe
3.00
61.66
3.89
188.26
10.32
-
0.3*
Pb
0.13
0.37
0.03
0.03
0.01
5
0.015
Mn
6.08
3.35
0.09
5.63
0.07
-
-
Zn
5.1
96.32
8.26
250.52
4.58
-
5*
* secondary drinking water standards
Table 3-7. Comparison of characteristics of CCA-treated wood used for Jambeck (2004)
and this study
Jambeck (2004) | This study
The % of CCA-treated included in waste stream
1%
1%
1%
As
1390 20.0
1960 27.
1330 10
Cu
814 52.4
1340 54.0
2350 50
Cr
1450 68.3
2550 48.0
2890 56

T/§UI IV q/Sui [Y
72
too
10 -
AEROBIC
Lys 1
O Lys 2
CD
O
O O
o
o
o
o 8* cy.
8 0o oo*o
o
0.1
100
200
300
400
Figure 3-1. Changes of A1 concentrations over time

As, mg/L As, mg/L
73
1.0
0.8 -
0.6 -
0.4 -
0.2 -
AEROBIC
Lys 1
O Lys 2
M
O O o
Oo
tPo
o o



o
O


0.0
p
- Qd O CTO f -
Qo
100
200
300
400
ANAEROBIC
A Lys 3
A Lys 4
A

A
A
A M^*A
A A A A
ViA
AA a4£^-2AAa
i 1
400 600 800
Days
Figure 3-2. Changes of As concentrations over time

Cr, mg/L Cr, mg/L
74
0.5
0.4 -
0.3
0.2 -
0.1 -
Lys 1
O Lys 2
AEROBIC
cP
0.0
i
o


e
8 % o
o
o
o *
cO
o o
n r
I I
0 50 100 150 200 250 300 350
Figure 3-3 Changes of Cr concentrations over time

Cu, mg/L
75
100
~Sb
B
10 -
kO
o
o
AEROBIC
Lys 1
O Lys 2

$
8
o o
o O o o Q
O O
0.1
%
0.1 -
0.01
0.001
50
aaa^
A AA
100
S/***t A
V a.
4fc A
Below Detection Limit
150
200
250
I
300
350
ANAEROBIC
&A
Aaa A
A
A A
A A
A
A A
A AMA A
A A
AA
A Lys 3
A Lys 4
200
Days
400
600
Figure 3-4. Changes of Cu concentrations over time

Pb, mg/L Pb> m§/L
76
10
1 -
0.1 -
6
0.01 -
AEROBIC
Lys 1
O Lys 2
\

O
0
o*o
O
o o
o
0*0 ,
o
0 o
0.001
50
100
150
200
250
i
300
350
A Lys 3
A Lys 4
ANAEROBIC
0.1 -
ASAVa
a^aaa^
a Aa A
* aA
AaAAA aa
A .A
aa a
0.01 -
"VA
AA -A ^
A \
Aa a a
A
Below detection limit of ICP for Pb
0.001
~r
100
200
300
400
Days
500
T
600
700
Figure 3-5. Changes of Pb concentrations over time

Fe, mg/L Fe, mg/L
77
1000
0 50 100 150 200 250 300 350
Figure 3-6. Changes of Fe concentrations over time

Mn, mg/L Mn, mg/L
78
o.oi
50
100
150
200
250
300
350
Figure 3-7. Changes of Mn concentrations over time

Zn, mg/L Zn, mg/L
79
300
250
200 -
150
100
50 r


AEROBIC
Lys 1
O Lys 2
O
9 o<£
o
CP
r>Q0ooft 8cgoo 8
BDL
50
100
150 200
Days
250
300
350
700
Figure 3-8. Changes of Zn concentrations over time

80
pH
Figure 3-9. Distribution of As over a C-pH diagram

81
pH
Figure 3-10. Potential- pH diagram of Cr (Richard and Bourg, 1991)

Metal concentrations (mg/L)
82
Figrue 3-11. Distribution of Cu over a C-pH diagram

y.
oq
s
H;
a>
U)
i
to
>
Q-
C/3
O
3
*K
5'
3
O
*t-
3
o
o
3
C/3
O^
SI
<*>
a
Adsorption capacity (mg/g)
o
-i
f-
70
b
*
-'.'.-'.'.'. 'A
- :'WN VN VN
77/7777/7727
lili
& O z O
' -TI -T-! m
O O O O O
O O O O O
O to -^. On OC
800
Adsorption capacity (mg/g)
to
to
I
to
ro
j
to
-U
u
-u
to
)
-U
70
pa
*
to
to
i
to
to
J
to
i,
to
P
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-fc.
-U
77
ta
*
oo
OJ

2!
ero'
c
-!
OI
U>
K)
rT
O
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5'
c
n>
D-
Adsorption capacity (mg/g)
o
o
p
La O La
to to u> p
O La O La
£
N
3
i.
r
¡III
^ o z o
ti ti m
800
Adsorption capacity (mg/g)
ooooooppp
ooooo *r*r
ON)4^0s00OK)^0v
00
-P^

Total mass of metals adsorbed on lignocellulosic materials (mg)
85
120
100
80
60
40
20
0
300
250
200
150
100
50
0
Figure 3-13. The comparison of aerobic and anaerobic lysimeters in respect of total mass
of metals adsorbed on lignocellulosic materials

Total mass of metals adsorbed on lignocellulosic materials (mg)
86
600
500
400
300
200
100
0
14000
12000
10000
8000
6000
4000
2000
0
Figure 3-13. (continued)

Metal concentrations (mg/kg)
87
Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic) (organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic) (organic) (organic) (platic) (plastic)
Figure 3-14. The comparison of metal concentrations adsorbed on organic (newspaper
and cardboard) and plastic waste

Metal concentrations (mg/kg)
88
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Aerobic Anaerobic Aerobic Anaerobic
(organic) (organic) (platic) (plastic)
Figure 3-14. (continued)

89
OXIDIZED REDUCED
Methylation
Q : precipitated
o : precipitated in the presence of sulfur
Figure 3-15. Fate of heavy metals thermodynamically occurred in aerobic (oxidizing) and
anaerobic (reducing) conditions (Bridle, 2004; Drever, 1988; Sadiq, 1997;
Masscheleyn et al, 1991; Richard and Bourg, 1991; Benjamin, 2000; McBride
and Blasiak, 1979)

Metal concentrations (mg/L) Metal concentrations (mg/L)
90
3.5
3.0
2.5
2.0
1.5
1.0
0.5
0.0
As
0.5
0.4
0.3
0.2
0.1
0.0
1
4
Lysimeters
Lysimeters
Figure 3-16. Comparison of concentrations of metal leached between aerobic and
anaerobic lysimeters.

Figure 3-16 (continued)
Metal concentrations (mg/L)
ro u
U <-/
8 8 8 8 8 8
r k>
*<
C/
3'
O
HT
55 <-
r
v;
c>
3
o
o
55
500
Metal concentrations (mg/L)
o
k> bo N>
N) -U 00 N>
so

Cumulative metals (Al) released out of the lysimeters
(mg)
92
Mass loss, %
Figure 3-17. Changes in cumulative mass of meta released over a mass loss, %

Cumulative metals released out of the lysimeters (mg)
93
Mass loss, %
Figure 3-17 (continued). Changes in cumulative mass of metal released over a mass loss,
%

Metal concentration (mg/L)
94
3
1
0.01 -
0.001 -
0.0001
10
0.01 -
0.001 n 1 1 1 1 1 1
Jambeck Jambeck Aerobic Aerobic Anaerobic Anaerobic
(acid) (methane) (acid) (alkali) (acid) (methane)
Figure 3-18. Comparison of As, Cu and Cr leaching trend of the lysimeters to other study
(Jambeck, 2004)

CHAPTER 4
THE EVALUATION OF LIGNOCELLULOSIC WASTE DECOMPOSITION OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILLS
4.1 Introduction
According to an EPA report, 153 million tons of lingocellulosic materials were
generated as part of the MSW stream in 2003 (USEPA, 2005). In addition to MSW
components such as paper and wood products, lignocellulosic wastes also include forest
industry, pulp and paper industry, agricultural and food processing wastes (Source:
http://www.utoronto.ca/forest/termite/lig-mat.htm, Last accessed November 17th, 2005).
Since lignocellulosic materials occupy approximately 65% of the U.S. MSW stream by
weight (USEPA, 2005), their biological decomposition in landfills results in a large
volume of gas production and leads to the settlement of the landfill surface. An
understanding of this decomposition process is thus helpful to a landfill lysimeter.
The term lignocellulosic implies, materials consisting primarily of lignin and
cellulose. In addition to macro-cellulose (glucan), many different sugar residues
(hemicellulose) may be present. Cellulose consists of numerous glucoses linked by P-1-4
linkages, while lignin is composed of benzene-ring-containing monomers with net
structures. Lignin is present extensively with cellulose within and between distinctive
morphological regions protecting the cellulose and other cell structures (Pagarwal, 2005;
Brown, 1985). Chandler et al. (1980) showed the biodegradability of lignocellulosic
materials to be strongly associated with lignin content. Stinson (1995) showed that
methanogenic activity was highly affected by the degree of delignification of wood
95

96
samples. Generally, wood wastes are classified as recalcitrant materials in anaerobic
landfill conditions because of the high content of lignin (Chandler et al., 1980; Holt and
Jones, 1983; Stinson and Ham, 1995). The lignin content in wood differs according to
species, but typically, 27-33% of softwood species such as Douglas fir and Southern Pine
is composed of lignin, and 18-20% of hardwood species is composed of lignin (White,
1987).
Aerobic bioreactor landfills, which introduce air into a landfill to more rapidly
stabilize lignocellulosic wastes, are reported to have same advantages over conventional
anaerobic landfill (Reinhart et al, 2002). It is reported that lignin can be degraded through
the pretreatment or composting process of lignocellulosic materials (Fox and Noike,
2004; Vikman et al., 2002). Stinson and Ham (1995) reported that the degradation of total
cellulose contained in high lignin-containing materials such as newspaper dramatically
increased as lignin content decreased. If it is possible to decompose and/or depolymerize
lignin in aerobic landfills, greater decomposition of lignocellulosic materials can be
expected, which will lead to a reduction in the ultimate disposal capacity needed for the
landfill.
In this research, the biodegradation of lignocellulosic waste under aerobic and
anaerobic landfill conditions was evaluated. As described in chapter 2, four stainless steel
lysimeters were operated, two each aerobic and anaerobic bioreactors. Two of these (one
aerobic and one anaerobic) were excavated at the end of the experiment. The methane
yields of various lignocellulosic wastes contained in the lysimeters were measured. In
addition to the methane yields, the cellulose and lignin content of the wood waste

97
component was determined so that any differences in the lignin content between aerobic
and anaerobic landfill conditions could be observed.
4.2 Materials and Methods
4.2.1 Composition of Fabricated Waste
The waste stream fabricated for this research was based on the typical MSW
composition previously reported for the U. S. and Florida (see Figures D-4 and 5). For
simplification purposes, several minor components, such as textiles and tires were
excluded from the fabricated waste stream. A greater portion of commingled paper was
allotted as a substitute for those excluded materials. The ratio of office paper, cardboard
and newsprint in commingled paper (4.6:2.6:1) was estimated from previous published
data (FDEP, 2003 and USEPA, 2005). Southern yellow pine (SYP) was selected for the
wood waste used in this study and comprised 5% of the total waste stream. A detailed
classification of wood species that occur in MSW may not be available, but SYP is
known as one of the most widely used species for manufactured wood products (USDA,
1999).
4.2.2 Excavation and Processing of Decomposed Solid Waste
After 1 and 2 years of operation, one aerobic lysimeter and one anaerobic lysimeter
were dismantled and the decomposed waste was removed. When the carriage system was
removed, some amount of leachate remained pounded above the top waste layer in each
lysimeter. This leachate was removed using a vacuum pump; the volume was recorded.
The waste in each lysimeter was divided into four fractions during the removal process.
The wet weight of each fraction of solid waste was recorded. The wood blocks were
separated from the excavated samples which were then dried. The dried samples were
separated by each lignocellulosic component.

98
Separation was performed using a vibratory shaker assembly. Materials greater
than 0.475 cm were sorted into office paper, cardboard, newspaper, and identifiable non-
biodegradable materials. Materials less than 0.475 cm in size were weighed and ground
to less than 0.25 mm using an Urschell mill (Fritsch, German). VS content was
determined by comparing weight loss after igniting the ground samples at 550C for 2
hours. Since the size of ground wood samples was found to be irregular, samples were
screened again using No 20 mesh (mesh size = 0.8mm).
4.2.3 Methane Yield Determination
The methane yield of both the new and excavated samples was measured. A
synthetic media containing buffers, nutrients and trace metals was prepared using ASTM
method El 196-92 (ASTM, 1992). Anaerobically digested sludge obtained from a
laboratory-scale anaerobic reactor was added as an inoculum to the prepared media while
flushing with nitrogen gas. Three serum bottles were used for each type of waste sample.
A 100-mL portion of inoculated media was then transferred into the prepared serum
bottles along with solid waste samples under anaerobic conditions. Pure cellulose was
utilized as primary positive controls. Sludge without waste sample was employed as a
negative control. A gas sample was collected from each serum bottle once per week over
a 50-day period. At each sampling, a 50-mL capacity syringe was used to measure the
biogas volumes. The collected biogas was analyzed for methane and carbon dioxide
using a gas chromatograph (Model 5890, Hewlett Packard, USA) equipped with a
thermal conductivity detector and GS-Carbon PLOT capillary column (30 m X 0.32 mm
ID, Agilent Technology, Palo Alto, CA, USA).
In this research, only biodegradable materials were included to estimate the average
BMP value. Biodegradable components of the new waste included office paper,

99
newspaper, cardboard, SYP and dog food. For decomposed waste, however, the BMP
assay of the decomposed dog food was excluded due to difficulties in separation. Instead,
the fine fraction that passed through the screen (0.475 cm) obtained from the garbage
separation process was used for BMP assay. The overall BMP was determined by the
method used for the evaluation of the overall BMP of waste excavated from a landfill
(Lee, 1996; Townsend et al., 1996). The overall BMP values of the waste samples were
determined as the sum of the multiplication of the VS fraction and BMP of each
biodegradable waste divided by the sum of the VS fraction. This calculation can be
summarized in the following expression:
([dry mass fraction,g]i x [VS%\i x [BMP(L / gVS)]i)
[Overall BMP (L/g VS was,e)] =
^ ([dry mass fraction, g]i x [VS%]i)
i=i
(1)
where n is the number of different kinds of biodegradable materials under consideration.
The overall BMP value obtained from equation (1) represents the average methane
potential per mass of dry volatile solids in the waste. In order to correct it to the methane
potential based on the total mass of waste, the average volatile solids percentage was
multiplied by the average total solids percentage and the total BMP value obtained from
equation (1):
[BMP (L/g waste)] = [overall BMP (L/g VS waste)] x [F5%]average x [7Y%]average
Where,
(2)
Average VS (%) = ^([dry mass fraction,%]/ x [VS%]i)
1=1
(3)

100
Average TS (%) = ^ ([dry mass fraction,%]/ x [TS%]i) (4)
1=1
Based on the methane yield of the new and decomposed waste, the biodegradable
fraction contained in volatile solids was determined as:
, ,,, . .... a mass of waste (g) converted into biogas ...
Biodegradable fraction (%) = ><100
Initial dry mass (g)
Biodegradable volatile solids (BVS, %) = (VS %) x (Biodegradable fraction %)
Details about all calculations are described in appendix A.
4.2.4 Cellulose and Lignin Determination
In order to determine the cellulose content of selected solid waste samples (SYP),
lg of dried sample was placed in a 250 mL Erlenmeyer flask containing 6 mL of
deionized water, 24 mL of glacial acetic acid and 2mL of concentrated nitric acid. After
boiling for 20 minutes, 50 mL of toluene was added and swirled for 2 minutes. Waiting
until solids were settled, toluene was decanted through the Gooch filter apparatus. This
same procedure was repeated using 50 mL of ethyl ether. All solid materials were then
transferred from the flask to a Gooch crucible with glass-fiber filters using acetone.
Solids contained in the crucibles were washed by pouring 75 mL of hot toluene, hot
methanol and ether through the crucibles successively. Washed solid samples were dried
at 103C for 1 day and their dry weights were recorded. After removal of the organic
portion at 550C for 1 hour, weights of the residual materials were recorded again.
Cellulose content was then determined by multiplying 100 by the ratio of the organic
portion ignited to the initial weight of the sample.
To determine the lignin content, 1 g of sample was placed in a Gooch crucible with
a glass-fiber filter and washed with 150 mL of toluene and ethanol mixed in the 2:1 ratio.

101
After drying at 75C for at least one hour, 300 mg of sample was taken from the washed
sample and 3mL of 72% H2SO4 was added. Samples with 72% H2SO4 were then placed
in a shaking incubator at 30C for one hour. After reaction with H2SO4, samples were
transferred to a 250 mL Erlenmeyer flask and diluted with 84 mL of deionized (DI)
water. The flasks were covered with aluminum foil, and autoclaved for an hour at 121C
and 15 psi. The residues were filtered through Gooch crucibles with glass fiber filters and
washed with 100 mL of hot DI water. Washed residues were dried at 105C, weighed and
ignited at 550C for an hour. The lignin content was determined by dividing the
difference in sample weights before and after ignition by the initial sample weight.
4.2.5 Data Analysis
In order to evaluate the impact of air addition on waste decomposition, all methane
yield results were statically compared through the ANOVA test at the 0.05 level of
significance. ANOVA tests were conducted to compare the differences of the methane
yields of the lignocellulosic materials, and the cellulose and lignin analysis results of the
ground SYP blocks.
4.3 Results
4.3.1 Methane Yield of Raw Waste
By combining methane yields of all biodegradable organic fractions, the overall
methane yield of the raw waste was estimated. The methane yield of each component is
presented in Table 4-1. The overall methane yield of the organic fraction of the raw waste
calculated was 0.337 L/g VS (0.191 L/g total waste, dry). In comparison with the
methane yields of the organic fraction of MSW reported, the methane yield of the
fabricated waste here is higher than mechanical-sorted MSW but close to other hand-
sorted or source-sorted MSW (Table 4-2). Mata-Alvarez et al. (1990) pointed out that the

102
methane yield of mechanical-sorted waste could be low because great amounts of
inorganic waste may be included in sorted organic fractions. The default methane
potentials of Clean Air Act (CAA) and AP-42 are 0.170 L/g and 0.10 L/g of waste,
respectively (USEPA, 1997), but these values included all organic and inorganic fractions
of solid wastes disposed of in landfills. These values also include the moisture of waste.
When applying the field capacity (58%), methane yield of the raw waste could be 0.086
g/L total waste, wet.
Based on the biodegradable volatile solids (BVS) of the organic fraction of the
fabricated waste, the percentage of ultimate decomposition of fabricated waste was
estimated. A total of 48% of the fabricated waste can be ultimately decomposed based on
the methane yield data. BVS of the organic was calculated using the methane yields of
those components; BVS values calculated are summarized in Table 4-3. The detailed
procedure used for the estimation of BVS is described in appendix A.
Forty eight percent of the fabricated waste into mass corresponds to approximately
6500 g in each lysimeter. The total mass of lignocellulosic materials (office paper,
newspaper, cardboard, wood and dog food) loaded in each lysimeter was 9600 g. Thus,
ultimately 67% of total lignocellulosic materials were biodegradable. When the mass of
the biodegradable fraction was compared to the mass loss measured from the excavated
waste, 42% and 37% of the total lignocellulosic waste of the aerobic and anaerobic
lysimeters were decomposed during a test period, respectively. This corresponded to 62%
and 54 % of the lignocellulosic wastes having decomposed.
4.3,2 Solid Waste Excavation
The characteristics of solid waste excavated are summarized in Table 4-4. The total
amounts of water contained in each lysimeter were 15,600 mL and 18,800 mL for

103
lysimeters 2 and 4, respectively. The water in the column was approximately 3,500 mL
less than that initially added. This difference is thought to be a result of water lost by air
stripping for the aerobic lysimeter. The water loss (mL) was calculated assuming gas
emitted from the aerobic lysimeter was fully saturated. The calculated amounts of water
removed by gas-striping were 3,900 mL and 200 mL for lysimeters 2 and 4.
The dry weight of the solid waste excavated was compared with dry weight
calculated by volume of CH4 and CO2 and TOC concentration in leachate (see appendix
A). Dry weights measured and calculated were 8,715 g and 8,741 g for lysimeter 2 and
8,997 g and 9,258 g for lysimeter 4. The measured weight loss and the weight loss
calculated from gas production and leachate solute were comparable. They were less than
2% different for the aerobic column and less than 7% different for the anaerobic column
(Figure 4-1).
4.3.2 Mass Loss for Individual Components
The mass loss for individual components was calculated on the basis of the results
of waste separation procedure. Among four biodegradable wastes (office paper,
cardboard, newsprint and wood blocks), the largest difference of mass loss between the
raw and decomposed waste was observed from the office paper fraction for both
lysimeters (Figure 4-2). Theses differences were reduced in the following order: office
paper (OP) > cardboard (CB) > newspaper (NP) > wood (WD). The percentages of each
component of the excavated waste sample also changed in comparison to the raw waste
(Figure 4-3). The decomposition trends of these wastes were related to the
biodegradability of each component, as described next.

104
4.3.3 Biodegradability of Excavated Wastes
The biodegradability of each component was determined using the methane yield
data from the BMP assays. Figure 4-4 presents the changes in cumulative CH4 volume of
each component of waste layer 2-3 over time. The cumulative CH4 volume of all of the
components fell between cellulose (positive control) and the blank (sludge only; negative
control). Among the four biodegradable wastes, the greatest volume of methane was
produced from office paper fraction and the least methane volume was observed from the
wood fraction. Biodegradability of the lignocellulosic waste can be arranged as follows
(from largest to lowest): office paper > cardboard > newspaper > wood. Figure 4-5 shows
the differences of the methane yields of biodegradable wastes excavated from lysimeters
2 and 4 relative to those of the raw waste. Overall the methane yields of the solid wastes
excavated from the anaerobic lysimeter were statistically greater in comparison with
those of the aerobic lysimeter (p < 0.05). The higher the methane yields the waste
samples showed, the less decomposed they were. It is noted that higher methane yields of
newspaper were observed from all paper fractions of anaerobic lysimeter (Figures C-18
through C-20).
Table 4-4 presents the overall methane yields of each waste layer. These values
were calculated based on the methane yield of each component and the waste fraction
obtained from the garbage separation process. The greatest degree of waste
decomposition was observed from the middle layers (layer 2-2 and 2-3), and a relatively
low decomposition was observed with the top and bottom layers. For lysimeter 4,
however, though the overall methane yield was higher than that of the lysimeter 2, similar
methane yields were observed from most layers except for the layer 4-4. Among all waste
layers excavated from lysimeter 2 and 4, layer 4-4 showed the least waste decomposition.

105
This was because the pH of the layer 4-4 remained in the acid phase for more than 400
days. As previously described in chapter 2, a buffer (sodium bicarbonate) was added
from the top of the lysimeter. After the buffer addition, the pH of the waste layers was
measured on occasion using the front ports (see Figure 2-1 in chapter 2). While the pH of
the other layers changed to neutral, the pH of the layer 4-4 remained acidic. The pH of
layer 4-4 then increased to 7.3 at day 550.
Based on the methane yields of each layer, the mass losses of lysimeter 2 and 4
were calculated. The total potential gas volume generated from each layer was estimated
on the basis of methane yield of the raw waste. Calculated dry mass loss for lysimeter 2
using the methane yield of the all layers was 4,285g (wd, measured = 4,044g). The loss for
lysimeter 4 was 3088g (wdimeasured= 3,524g). These values were equivalent to 65.8 % and
46.9% of the biodegradable fraction of lignocellulosic materials (Figure 4-6). Though a
small difference between estimated and measured values was observed, the BMP assay
results of the excavated waste were shown to be a good parameter for the evaluation of
waste decomposition in landfills.
4.3.4 Biodegradability of Wood Waste
The excavated SYP blocks were analyzed for cellulose and lignin to access the
performance of the aerobic and anaerobic simulated landfills with respect to the
decomposition of waste with a high lignin content. Table 4-6 summarized the cellulose
and lignin analysis results of raw and decomposed ground SYP blocks. According to the
statistical analysis results, overall lignin and cellulose concentrations of SYP excavated
from the aerobic and anaerobic lysimeters were not significantly different from those of
raw SYP blocks (p > 0.05). However, a relatively large difference in cellulose content of
the SYP blocks from aerobic lysimeter samples 2-1 (42.3 %) and 2-2 (46.1 %) were

106
observed in comparison to the raw (50.5 %) and the anaerobic lysimeter blocks. A
comparison of SYP block biodegradability by methane yields was determined more
valuable.
In contrast to the cellulose and lignin results of the wood samples, the methane
yields of SYP excavated from the aerobic lysimeters were statistically lower than those of
raw and the anaerobic lysimeters (p < 0.05) (Figure 4-5). Average BMP assay results of
wood from the anaerobic lysimeters were statistically the same as the raw SYP.
There was no evidence to find if the lignin component was degraded in both the
aerobic and anaerobic lysimeters. Overall, the lignin concentrations of raw SYP blocks
were not significantly different from those of the both lysimeters.
4.4 Discussion
Through this research, waste decomposition was evaluated by conducting the BMP
assay on solid waste samples. The BMP assay results indicated that the methane yield of
the aerobic lysimeter was lower than that of the anaerobic lysimeter despite different test
periods (1 year and 2 years for the aerobic and anaerobic lysimeters, respectively). The
largest difference of methane yield between the raw and decomposed waste was observed
with office paper. These differences decreased following the order: office paper >
cardboard > newspaper > wood. This result indicated that waste may be decomposed in a
landfill in the same order. It was confirmed by the changes in the percentage of organic
fraction as waste decomposed (Figure 4-3).
The methane yield results of the newspaper and SYP blocks indicated that
lignocellulosic materials with high lignin contents can be more decomposable in aerobic
condition relative to anaerobic condition. It is reported that the volume reduction of
general lignocellulosic materials was not much different between aerobic and anaerobic

107
processes (Tchobanoglous et al., 1993). However, the effect of lignin on biodegradation
of lignocellulosic decomposition is reported to be less in aerobic conditions (Komilis and
Ham, 2003). The biodegradable fraction of lignocellulosic materials in terms of lignin
content can be written as:
Anaerobic: B = 0.83 (0.028) X (r2 = 0.94) (Chandler et al., 1980)
Aerobic: B = 0.85 (0.01) X (r2 = 0.50) (Komilis and Ham, 2003)
where X is the initial lignin contents (as %VS). The effect of lignin on lignocellulosic
decomposition in aerobic condition appeared to be more variable than that in the
anaerobic condition.
The BMP assay and lignin analysis results found no evidence that lignin could be
decomposed in either aerobic and anaerobic landfill conditions. If lignin components
were decomposed or depolymerized, methane yields of SYP blocks could be higher than
those of raw SYP. The most likely explanation may be because of particular ecological
demands of lignin scavengers such as White rot fungi and actinomycetes (Akhtar et al.,
1997); oxygen demands of these microorganisms are reported as 15% in atmosphere
(Reid and Seifert, 1982) while average oxygen content in the aerobic lysimeter was
below 6.5%. Furthermore, overburden pressure applied may limit oxygen transfer.
This research suggests that aerobic landfills may have some advantage for
decomposing high lignin-containing waste in comparison with anaerobic landfills.
However, the percentage of woody waste in MSW stream is relatively small because
Class III MSW landfills do not accept construction and demolition (C&D) debris.
Moreover, according to the United States Environmental Protection Agency (USEPA)
report, the recycling ratio of newspaper substantially increased in 2003 (USEPA, 2005).

108
For these reasons, it is proposed to conduct research on decomposition of woody
materials under air addition into the C&D debris landfill as a future research.
4.5 Conclusions
Decomposed waste was excavated from lysimeters 2 (aerobic) and 4 (anaerobic)
after a test period (1 and 2 years for aerobic and anaerobic lysimeters, respectively). The
mass of waste lost by waste decomposition of the aerobic lysimeter was greater than from
the anaerobic lysimeter. The mass losses measured and predicted by gas generation were
very similar. Due to the overburden pressure applied to the top layer of the lysimeters, the
greatest moisture content was observed from the top layer of both aerobic and anaerobic
lysimeters.
The overall methane yield (a measure of the degree of waste decomposition
occuring) of the aerobic lysimeter was lower than that of the anaerobic lysimeter despite
a shorter test period. The greatest difference of methane yields between raw and
decomposed waste was observed from office paper. This difference decreased following
the order: office paper > cardboard > newspaper > wood. Though cellulose
concentrations of decomposed SYP between the aerobic and anaerobic lysimeters were
statistically the same, the methane yields of SYP blocks excavated from the aerobic
lysimeter were significantly lower than those of the anaerobic lysimeter. The lower
methane yields of newspaper were also observed from the aerobic lysimeters. However,
lignin degradation was not observed from either aerobic or anaerobic lysimeters.

109
Table 4-1. Methane yields, VS and mass fraction of the lignocellulosic materials in raw
waste
Office paper
Cardboard
Newspaper
Dog food
SYP
VS, %
86.8%
98.5%
95.3%
90.0%
77.0%
Mass fraction,
%
27%
16%
6%
5%
15%
Methane
yield
(L/gvs)
0.401
0.265
0.094
0.027
0.540
Table 4-2. Comparison of methane yields of MSW with other studies
Sorting method
Fermenter
CH4 yield
(L/g VS added)
References
Mechanical -sorted
Lab plant
0.260
Baere, 1984
Mechanical -sorted
Pilot plant
0.187
Baere and Verstraete, 1984
Mechanical -sorted
Pilot plant
0.230
Valorga, 1985
Hand-sorted
CSTR
0.390
Pauss et al., 1984
Hand-sorted
BMP assay
0.205
Owens et al., 1993
Source-sorted
CSTR
0.399
Mata-Alvarez et al., 1990
Mechanical -sorted
(pre-composted)
CSTR
0.145
Mata-Alvarez et al., 1990
-
Landfill
0.17*
CAA
-
Landfill
0.10*
AP-42
Hand-sorted
BMP assay
0.337
This study
* L/g of refuse
Table 4-3. Biodegradable volatile solid (BVS) of organic fraction of t
ie raw waste
VS
Biodegradable
fraction of VS
(BVS)
Biodegradable
fraction of dry waste
office paper
87%
74%
86%
cardboard
99%
65%
66%
newspaper
95%
21%
23%
dog food
77%
77%
77%
wood
90%
6%
7%

Table 4-
. The physical characteristics of excavated waste
layers
depth (inches)
Wet wt
(lb)
vol (cf)
Density
(pcf)
dry density
(pcf)
Moisture
content
2-1
21-33
15.4
0.20
78.4
42
45.70%
2-2
33-45
11.5
0.2
58.46
21.7
62.9%
2-3
45-57
10
0.2
51.03
16.2
68.2%
2-4
57-66
8.9
0.15
60.46
23.7
60.7%
Lys 2
21-66
45.8
0.74
62.2
26.2
57.9%
4-1
22 28.5
8
0.11
75.2
37
50.20%
4-2
28.5-41
15
0.20
73.3
31
57.40%
4-3
41-53
12.4
0.2
63.31
25.3
59.9%
4-4
53-66
14.5
0.21
67.95
24.5
64.0%
Lys 4
28.5 66
41.2
0.61
67.1
33.5
58.8%

Ill
Table 4-5. Overall met
iane yields of waste la
yers of the lysimeters 2 and 4.
Waste layer
L/g-VS
vs
L/g
2-1
0.148
83.4%
0.1236
2-2
0.092
83.3%
0.0763
2-3
0.093
81.4%
0.0760
2-4
0.133
85.7%
0.1136
4-1
0.167
84.0%
0.1399
4-2
0.156
87.9%
0.1373
4-3
0.163
86.5%
0.1414
4-4
0.225
86.5%
0.1943
Raw waste
0.337
88.4%
0.2974

112
Table 4-6. Summary of cellulose and lignin content of the wood samples
Sample ID
Cellulose
Lignin
Measurement
Average
Measurement
Average
raw
53.2%
50.47%
29.2%
28.80%
47.7%
28.4%
2-1
44.4%
42.34%
31.3%
30.46%
40.3%
29.6%
2-2
46.3%
46.09%
30.9%
30.88%
45.9%
30.8%
2-3
49.0%
47.26%
32.1%
31.80%
45.5%
31.5%
2-4
48.7%
48.94%
29.3%
30.40%
49.1%
31.5%
4-1
47.4%
46.75%
27.6%
28.62%
46.2%
29.7%
4-2
45.4%
45.38%
28.5%
27.96%
45.4%
27.4%
4-3
50.8%
50.22%
27.4%
28.27%
49.6%
29.2%
4-4
49.1%
49.87%
27.9%
28.26%
50.7%
28.6%

Figure 4-1. The dry weight differences between predicted and measured remaining mass
Lysimeter 2 Lysimeter 4
(Aerobic) (Anaerobic)
Dry weight (g)
K>
O
o
o
4^
o
o
o
Os
o
o
o
oo
o
o
o
o
o
o
o
K>
o
o
o
Initial
Calculated
Remaining mass
Predicted
lost from gas
Measured
Remaining mass
Initial
Calculated
Remaining mass
Predicted
lost from gas
Measured
Remaining mass
14000

114
Figure 4-2. Comparison of dry weights between raw and decomposed lignocellulosic
wastes

115
wood NP
5% 6%
(A)
OP
6% CB
(B)
Figure 4-3. The changes in the percentage of waste components after decomposition; (A)
raw waste components and (B) decomposed waste (aerobic)

116
O 10 20 30 40 50
Days
Figure 4-4. Changes in cumulative methane volume of lignocellulosic materials over time

Methane yields (L Q\{^lg VS)
117
0.5
0.4
0.3 -
0.2 -
0.1
0.0
*
i
I I Raw waste
V////A Anaerobic
K888881 Aerobic
ii
ni
idfe
officepaper cardboard newspaper wood
(A)
0.035
0.030
> 0.025
SC
0.020
2
13
0.015
C
C3
-g
3a 0.010
0.005
0.000
Figure 4-5. Methane yields and weight differences of lignocellulosic materials among
raw and two lysimeters (A) all lignocellulosic materials; (B) wood only

118
80
-a
-a
ed
Ui
60
u
-o
C/
>
m
60
40 -
20 -
Based on mass
Y////A Based on gas produced
Based on methane yield
Anaerobic
t = 2 years
Figure 4-6. The comparison of dry masses measured and predicted by gas generated and
BMP assay

CHAPTER 5
LANDFILL SETTLEMENT BEHAVIOR WITH WASTE DECOMPOSITION
5.1 Introduction
Since Sowers (1973) proposed the equation to describe landfill settlement using
established soil consolidation models, various methods have been developed to interpret
and predict landfill settlement. Gibson and Lo (1961) applied a damping system to
account for landfill settlement and Ling et al. (1998) attempted to find best-matched
curves for landfill settlement behavior using various mathematical functions. In the past,
landfill settlement behavior has been interpreted similarly to soil consolidation models.
These types of interpretation were successfully applied to primary settlement, which
readily occurs as a result of overburden pressure. However, unlike soil consolidation,
landfill settlement is primarily determined by secondary settlement, which is controlled
for the most part by waste decomposition. Among numerous models developed so far,
only a few include waste decomposition to account for landfill settlement (Park and Lee,
1997). Secondary settlement is even more important for bioreactor landfills which are
designed to enhance waste decomposition by the addition of moisture and/or air. There is
a need to develop and to validate a model that can reliably predict secondary settlement.
It is proposed that methods for predicting the secondary settlement at MSW
landfills can be developed by relating waste decomposition over time with landfill
settlement over time. The most common approach for modeling, landfill gas production is
to model waste decomposition as a First-order function. A similar approach would be to
relate settlement (volume loss) to waste decomposition (mass loss) and to thus predict
119

120
settlement as a function of site specific waste decomposition. However, no data are
available to characterize waste volume loss with respect to mass loss caused by
decomposition in landfills.
The objective of this research was to collect data on mass loss versus volume loss
for future use in settlement model development. In order to correlate mass loss and
volume loss, a lab-scale experiment was designed where waste was decomposed in
simulated landfills in the laboratory, with both mass loss and volume loss being measured.
A difficulty with using lab experiments to simulate landfill settlement is that it is hard to
simulate true landfill conditions, especially, the large overburden pressure. For this
reason, in this research, the experiments conducted in this research included the
application of overburden pressure to make the laboratory condition closer to the field.
5.2 Materials and Methods
5.2.1 Lysimeters
Four simulated landfill columns (lysimeters) were used in this research, They each
consisted of a primary stainless steel column and a carriage system component. Two
lysimeters were operated as aerobic lysimeters (lysimeter 1 and 2) and two were operated
as anaerobic lysimeters (lysimeter 3 and 4). Three parameters (temperature, air addition,
and overburden pressure) were controlled in an effort to simulate actual aerobic or
anaerobic bioreactor landfills. Details about the lysimeters and the carriage system are
described in chapter 2 and in Figures B-l through B-3.
Mixed fabricated waste samples were created and loaded into the columns as four
fractions to prevent waste component stratification in a particular place in the column
(composition of the fabricated waste fractions and their weight are summarized in
appendix C). Prior to loading, 6 inches (15.3 cm) of river rock was placed at the bottom

121
of each lysimeter and a geotextile was placed between rock and waste. Each waste
fraction was then loaded and compacted until it occupied 25 % of the depth of the
lysimeter. Two liters of DI water was added along with the compaction of each waste
fraction. After loading, 11L of additional water was added from the top of each lysimeter.
The waste was compacted to a density of 30 lb/ft dry (480.6 kg/m dry)-
5.2.2 Application of Overburden Pressure
Overburden pressure was applied to the fabricated waste in order to characterize the
decomposition of waste disposed in a deep landfill. An Enerpac hydraulic cylinder and
hand pump was utilized to generate overburden pressure. The hydraulic cylinder was
mounted on the top of the carriage system. Overburden pressure was transferred through
a shaft attached to the cylinder to the waste in the lysimeter. As settlement proceeded, the
length of the shaft was extended by exchanging and connecting other pieces of shaft of
different lengths.
A load of 2040 lb/ft2 (98 kPa) was used as the overburden pressure to simulate a lift
of waste overlain by 40 ft (12 m) of waste. This was for the simulation of the pressure
applied under 4 ft (1.2 m) cover soil and 36ft (11 m)-depth of waste. It was assumed that
the cover soil occupied 10 % of the volume of the landfill, the density of the cover soil
was 110 lb/ft3 (1760 kg/m3), and the compacted waste density was 44.4 lb/ft3 (710
kg/m3). In the field, waste can be compacted by modem technology to 60 lb/ft3 (960
kg/m3). The density of waste used for this research falls into the good compacted range
for MSW (Oweis and Khera, 1998)
The measurement of the initial depth of each lysimeter was conducted after the first
application of overburden pressure at day 1. Thus the percentage of settlement addressed
here is the ratio between the depth of waste deformed and the initial depth of waste. The

122
daily difference of settlement was recorded by measuring the movement of the length of
the shaft visible above the flange after adjusting the pressure to 510 psi (3516 kPa).
5.2.3 Compression Index and Phase Separate Method
The measured settlement data were fit to several different relationships that had
been previously proposed to model landfill settlement in the secondary phase (without
consideration of condition with decomposition rate). These relationships included the
modified secondary index proposed by Sowers (1973). Bjangard and Edgers (1990)
developed the modified secondary index and explained landfill settlement mechanisms by
separating the landfill settlement curve by different phases (phase separate method).
Solving for parameter described on part of the relationship allowed comparison with
other studies.
A modified secondary index (Ca) was used to describe settlement behavior of the
aerobic and anaerobic lysimeters. Originally, the secondary compression index (Ca) was
used to describe the secondary settlement for soil tests, but Sowers (1973) first applied
this concept to landfill settlement. Since it is hard to estimate void volume in the field, the
secondary compression index was modified. The modified compression index (Ca) is
used to estimate the settlement that occurred after the first mechanical settlement. The Ca
and Ca can be expressed as follows:
C =
log(t2/t,)
(1)
C. =
AH
G
H-\og(t2/ti) 1 + e0
(2)

123
where Ae is the change in void volume; AH is the change in thickness of the waste layer;
H0 is the original thickness of the waste layer; ti is the starting time of secondary
settlement; and t2 is the ending time of secondary settlement.
Bjamgard and Edgers (1990) proposed the phase separate method to describe the
major causes of landfill settlement. They separated landfill settlement curve into two
phases, (Ca)min and (Ca)max by slope. The first phase, (Ca)m¡n, indicates that settlement
occurs by mechanical interactions such as delayed compression of the refuse. Landfill
settlement that occurs in the second phase, (Ca)max, is caused by both mechanical
interactions and waste decomposition (Figure 5-3)
5.2.4 Estimation of Mass Loss
The mass loss in each lysimeter was estimated by gas and leachate quality. The gas
volume generated was coupled with the biochemical reaction models for cellulose
biodegradation were used to calculate the mass of waste required for the moles of gas
obtained. Total organic carbon (TOC) concentrations were used to calculate the amount
of mass degraded, but remained in leachate before gas conversion. Details are described
in appendix A.
The settlement, gas and leachate data were collected for 1 and 2 years from the
aerobic and anaerobic lysimeters, respectively. At the end of the test period, lysimeters 2
and 4 were dismantled and partially decomposed waste was excavated. The waste loss
was then compared with the mearsured mass loss. The mass losses predicted of
lysimeters 2 and 4 were 4,069 g and 3,787 g, respectively, while the actual mass losses
were 4,044 g and 3,525 g, respectively.

124
5.2.5 Volume Loss versus Mass Loss
In this research, it was hypothesized that the mass loss resulting from waste
decomposition is the primary cause of volume loss. Microscopically, wastes consist of
small particles, and these particles may support each other to maintain their structure
against outer forces. A reduction in the number of particles disrupts this balance and
ultimately leads to loss of volume under the application of pressure. If the cross sectional
area is constant, such as the inside of lysimeters, the volume loss can be directly related
with a decrease of waste depth. Park and Lee (1997) explained the relationship between
landfill settlement and waste decomposition using an artificial model consisting of ice
and inert particles. As the volume of ice particles is reduced by melting, the overall
volume is reduced, a concept that can be applied to landfill settlement.
In order to evaluate the relationship between settlement and mass loss, the
percentage of settlement and overall mass loss obtained from each lysimeter at the end of
a test period were plotted. In order to find a better relation coefficient, linear and semi
logarithm correlations between the two parameters were conducted. This relationship can
be written as:
[settlement, %] = (AH, %) = a x/[mass loss, %] + P (3)
where, a and P were parameters obtained from the relationship. When deriving the
equation (3), several data points which deviated from this relation were excluded. These
data points were obtained during the first 30 days of the lysimeter studies. Thus, these
data points may have been influenced by primary mechanical settlement rather than
biological decomposition.

125
5.3 Results
5.3.1 Settlement Behavior over Time
Figures 5-1 and 5-2 show the settlement of the aerobic and anaerobic lysimeters
with respect to gas generated and pH. Though the aerobic lysimeters remained in an acid
phase for 200 days, CO2 gas was constantly generated over time. This trend corresponded
with the settlement behavior of the aerobic lysimeters. As previously discussed in chapter
2, the accumulation of organic acids might be localized at the bottom of the aerobic
lysimeters due to improper air addition. Thus, it is concluded that the pH of leachate in
the aerobic lysimeters during the acidic phase might not be representative of the pH of
the entire lysimeter and that waste was decomposed aerobically regardless of the pH of
leachate. For the anaerobic lysimeters, however, little change in settlement was observed
while the anaerobic lysimeters remained in the acid phase. For the first 100 days,
approximately 5 7% of the settlement was observed along with CO2 gas increases, but
the settlement was retarded until pH of the anaerobic lysimeter started to change at day
400. Most of the settlement observed from the aerobic and anaerobic lysimeters took
place in a similar trend as the increase in gas. These results indicate that long-term
landfill settlement is mainly affected by waste decomposition. This could be confirmed
by comparing different phase of landfill settlement behavior.
Figure 5-3 depicts the changes in settlement over time on a logarithm scale. The
settlement trends were separated into two phases, (Ca)mjn and (Ca)max by the changes in
slope. For the lysimeters 1 and 2, time differences (At) for (Ca)min were only 50 and 30
days, while At for (Ca)mm were 210 and 410 days for the lysimeters 3 and 4 (Table 5-3).
These results indicate that the settlement of the aerobic lysimeter mainly occurred by

126
waste decomposition ((Ca)max phase). The occurrence of biological activity during the
settlement can be also confirmed by an increase in cumulative gas over time (Figure 5-3).
For lysimeter 4, a rapid settlement curve was observed after a long lag period. It is noted
that the depth difference (AH) of lysimeter 4 during the (Ca)max phase is close to AH of
lysimeter 1 during the same phase. This result indicates that the aerobic landfill may
show the better settlement rate as opposed to the anaerobic landfill within a very short
period of time. However, great settlement performance was shown in both aerobic and
anaerobic lysimeters during the (Ca)max phase rather than (Ca)min phase. Therefore, it is
concluded that waste decomposition plays an important role for landfill settlement under
both aerobic and anaerobic conditions.
5.3.2 The Relationship between The Settlement and Mass Loss
Figure 5-4 depicts the relationship between the percentage of overall settlement and
mass loss. Since lysimeter 3 remained in the acid phase for 600 days, the percentage of
mass loss and its settlement at the end of a test period were substantially lower than other
three lysimeters. It is noted that the relationship between settlement and the overall mass
loss of each lysimeter correlated well (r = 0.88) despite their different overall mass loss
percentage. This can be confirmed by plotting all mass loss and settlement data of the
four lysimeters.
Figure 5-5 depicts the changes in the lysimeter depth over percent of mass loss
correlated (a) linearly and (b) semi-logarithmically. Comparing the relation coefficient
(r ) between two plots, semi-logarithmic correlation showed better fit between mass loss
and settlement (r2 = 0.89). This relation can be mathematically expressed as follows:
[settlement, %] = (AH, %) = 16.90 log [mass loss, %] 6.24 (4)

127
It is noted that this relationship is applicable to both aerobic and anaerobic
lysimeters. It also shows that the degree of waste decomposition is associated with
volume loss irrespective of the decomposition rate.
5.3.3 Ultimate Settlement
Prior to applying the settlement model as a function of mass loss, it is important to
consider the limitation of the waste decomposition. As previously discussed in chapter 4,
waste can be degraded up to the point BVS. Biodegradable volatile solid of the raw waste
was determined as 67% (see chapter 4). Therefore, any predicted values beyond that
point are necessarily deemed to be overestimation and need to be replaced with the
maximum mass loss. The time to reach the BVS may be variable depending on the
landfill conditions such as percentage of moisture, temperature and waste compositions.
In addition to the natural conditions, the time to reach BVS can be advanced by moisture
addition and air injection.
5.4 Discussion
5.4.1 Compression Index
Table 5-3 represents the modified compression indices of other studies in
comparison with results of this lysimeter study. In is noted that (Ca)'min values of
different case studies are similar while (Ca)'max values appear to be quite different. In
many cases compression index value of settlement occurred by mechanical interaction
would be less than 0.1. Sowers (1973) reported that the compression indices might
increase as organic content increased. Sowers also suggested that the compression index
could be a function of void ratio. In order words, compression indices may increase as
void ratio increases. Since the void volume is defined as the ratio of the volume of voids
to the volume of solids (Das, 2002), various modem landfill technologies such as

128
moisture addition and air addition may increase the compression index. In contrast,
highly compacted waste may have low compression index due to relatively lower void
ratio. In this study, however, initial conditions of all lysimeters were the same and the
flow rate of air addition (70mL/min) would not be high enough to change the void ratio.
Thus, the changes in compression indices could be contributed by the influence of waste
decomposition.
The compression index could be important and useful tool to compare the landfill
performance in terms of settlement, but Holtz and Kovacs (1981) pointed out that the
compression index for the secondary long-term settlements were oversimplified to
describe more complicated real behavior. Moreover, the compression index may not be
useful to compare the lab-scale landfill settlement occurred within a short period of time.
For example, the compression index of lysimeter 4 appeared to be 2 times greater than
that of lysimeter 2. However, the differences of time and the change in strain of lysimeter
2 were more advanced than those of lysimeter 4 (Table 5-1). Since the date starting with
the second phase was too early, log (t2/fi), a denominator of compression index equation,
was not comparable with that of the anaerobic lysimeters.
5.4.2 Correlation of Mass Loss and Volume Loss
As Figures 5-3 and 5-4 shown, the percentage of mass loss may not correspond
with the same percentage of volume loss; small volume loss may require relatively larger
mass loss. This phenomenon is difficult to assess because waste is heterogeneous and less
understood. Wall and Zeiss (1995) reported that total settlement of a landfill would be
range from 25% to 50%. If it takes consideration that more than 40% of biodegradable
lignocellulosic materials still remained, the relationship between mass loss and volume
loss (equation (4)) may be changed, ultimately.

129
5.4.3 Application
The development of the landfill settlement model using the biological reactions
may require many case studies since biological waste decomposition is affected by many
parameters. Through the many research efforts, the first-order kinetics model was
reported as one of the most suitable to describe the biological decomposition (El-Fadel et
al., 1989). However, it may be difficult to express all landfill settlement using the aspect
of waste decomposition. Once waste is disposed in landfills, initial compression and
primary settlement may take place by its own elasticity and mechanical interactions
caused by overburden pressure. After a period of time, the waste component, which can
be readily decomposed, may be depleted, and mass loss of waste is mainly controlled by
hydrolysis of waste (Park and Lee, 1997). The settlement models using this hydrolysis of
waste are called a bioconsolidation model (El-Fadel and Khoury, 2000). Park and Lee
(1997) proposed the bioconsolidation model using the hydrolysis coefficient, k:
E (t) = E tot-dec (1 6 ) (5)
where, e tot-dec= total amount of compression due to the waste decomposition; k =
hydrolysis coefficient (or decay rate); and t = time (year)
Based upon the data obtained from this research, mass loss also can be expressed
by the equation (6):
Mass loss, % = ^ x 100 = 100(1 e~h) (6)
M o
If this equation (6) is substituted into equation (4), a mass loss-settlement
relationship, a similar equation can be derived to the bioconsolidation model.
= H, (0.7244 0.169 log(l e*))
(7)

130
El-Fadel and Khoury (2000) pointed out that the determination of the kinetic
coefficient (k) (or hydrolysis coefficient) would be at best a difficult task in landfills.
Hoeks (1983) and Ham (1988) also pointed out that this coefficient (k) would be changed
by landfill phase. The k values reported were 0.046, 0.028 0.139 and 0.462 1.386 yr'1
by slow, moderate, and rapid waste decomposition, respectively (Hoeks, 1983; Ham,
1988). From this point of view, settlement loss mass relationship obtained from this
research may be meaningful. The mass loss calculated from gas and leachate qualities
may provide the important tool to determine this kinetics coefficient. Using the
settlement data and mass loss, the kinetics coefficient can be determined by:
= \-e~h
M o
In
1-
M
M
= -kt
o y
(8)
Figure 5-6 shows the correlations of the equation (8) using the settlement data and
mass loss of the aerobic lysimeters. However, Hoeks (1983) and Ham (1988) mentioned,
for the anaerobic lysimeters, k values were variable by the different phases (Figure 5-7).
The results of the k value are summarized in Table 5-2. Figure 5-8 shows the example of
the application of the k values and settlement waste loss relationship. Park and Lee
(1997) pointed out that the bioconsolidation model might be useful to correlate landfill
settlement in terms of waste decomposition behavior, but landfill settlement behavior
would be return to the mechanical stage after biodegradable fraction is depleted.

131
5.5 Conclusions
The landfill settlement behavior occurring in aerobic and anaerobic simulated
landfills was mathematically analyzed. The logarithm of mass loss can be linearly
correlated with volume loss as follows:
[Settlement, %] = (AH, %) = 16.90 x log [mass loss, %] 6.24
Assuming the same cross-sectional areas, volume loss and the percentage of settlement
can be identical. With these relationships, the secondary settlement of aerobic and
anaerobic simulated landfills could be mathematically modeled. The first-order
exponential functions could be used to describe waste decomposition.
Settlement data obtained from this research could provide a useful tool in determining the
decay value, the most difficult derivative in the bioconsolidation model. Assuming that
waste decomposition follows a first-order kinetic model, the decay coefficients
determined based on the settlement data were 0.379 and 0.377 yr'1 for the aerobic
lysimeters. Decay coefficients for the anaerobic lysimeters changed by waste
decomposition phase. Decay coefficients calculated were 0.015 and 0.022 yr'1 and 0.194
and 0.246 yr"1 for the acid and methane phases of the lysimeters 3 and 4, respectively.

132
Table 5-1. (Ca)min and (Ca)max values of lys 1 through 4
(cy
min
(C
1 max
Todays
t2
AH, %
(Ca) mjn
t.
t2
AH, %
(Ca) max
lys 1
23
70
3.05
0.063
70
316
11.45
0.175
lys 2
33
60
1.88
0.072
60
360
17.64
0.227
lys 3
46
260
3.76
0.050
260
761
6.52
0.140
lys 4
29
440
3.75
0.032
440
718
11.41
0.536
Table 5-2. k values of aerobic and anaerobic lysimeters
Lys 1
Lys 2
Lys 3
Lys 4
Overall (yr1)
0.379
0.377
0.0202
0.1048
delayed phase
(lys 3: 0 ~ 650 days; lys 4: 0~ 400 days)


0.0145
0.0219
Loe phase
0.379
0.377
0.1944
0.2458

133
Table 5-3. Comparison of compress indices between current study and other studies
References
mjn
V'-'ctj max
24 landfill case (Bjargard and Edgers, 1990)
0.019
0.125
Laboratory large scale (Gandolla et al., 1992)
0.063
0.34
Laboratory large scale (Lee et al., 1995)
0.063
0.149
Lysimeter 1 (this study)
0.063
0.175
Lysimeter 2 (this study)
0.072
0.227
Lysimeter 3 (this study)
0.050
0.140
Lysimeter 4 (this study)
0.032
0.536
15-year-old landfills, Boston, MA
0.24
Old landfill, WV
0.30
10-year-old landfills, Elizabeth, NJ
0.02

134
Figure 5-1. The changes in settlement, cumulative gas (CO2) and pH over time

Cumulative gases,
135
Figure 5-2 The changes in settlement, cumulative gas (CO2 and CH4) and pH over time

Settlement, %
136
Days
Figure 5-3. Settlement behaviors and compression coefficients of aerobic and anaerobic
lysimeter over a period of time

137
Figure 5-4. Relationship between settlement and overall mass loss of the aerobic and
anaerobic lysimeters

Settlement, % Settlement, %
138
(a)
Figure 5-5. Relationship between percentage of settlement and mass loss

139
O 100 200 300 400
Time (days)
Figure 5-6.Correlation of logarithm of mass loss of the aerobic lysimeters over time
Figure 5-7. Different k values of anaerobic lysimeters at lag and log phases

140
Figure 5-8. Settlement prediction of the aerobic lysimeters

CHAPTER 6
SUMMARY AND CONCLUSIONS
6.1 Summary
In this research, the overall performance of aerobic and anaerobic landfills was
compared. Four stainless-steel lysimeters were used as simulated landfills. Based on
waste stream data published by the U.S. EPA and FDEP, fabricated garbage was made
and loaded in each lysimeter. Constant pressure (2040 psf) was applied to the fabricated
waste to simulate MSW placed at a depth of 40 feet in the landfill. Leachate produced
was collected from each lysimeter on a weekly basis. Leachate collected was injected
back to the lysimeters, with deionized water used as makeup water, after analyzing for
metals and conventional water quality parameters. In addition, gas quality and settlement
were monitored during the entire operation. The period of time spent researching the
aerobic and anaerobic lysimeters differed (365 and 719 days, respectively). After the
lysimeter studies were completed, one each of the aerobic and anaerobic lysimeters was
dissembled and the decomposed wastes were excavated. Lignocellulosic materials
contained in the wastes were separated and analyzed for metal content and the degree of
biodegradation.
More than 90% of the leachate COD, BOD and TOC was reduced within 100 days
in the aerobic lysimeters. The concentrations of ammonia remained the same during the
acidic phase in the anaerobic condition; however, ammonia concentrations increased to
an amount four times higher in the methane phase than the initial concentration. No such
dramatic change in ammonia was observed from the aerobic lysimeter. It was noted that
141

142
the pH of the aerobic lysimeters increased to a level more alkaline than the pH of the
anaerobic lysimeters. This resulted from the change in the carbonic system caused by
CO2 stripping by air injection. The highest recorded pH of the aerobic lysimeters was
9.17. Air injection was used in an attempt to recover one of the acid-stuck anaerobic
lysimeters. The initial pH was 6.2 and the methane content was 38%. After air injection
started, the pH increased to 7.3 within 8 days, and the methane content increased to 55%
within 11 days after air injection stopped.
Redox and pH changes in the aerobic and anaerobic lysimeters resulted in changes
in metal concentrations dissolved in leachate. Among the 8 metals under consideration,
the average concentrations of As, Fe, Mn, and Zn in the leachate of the anaerobic
lysimeters were significantly greater than those of the aerobic lysimeters. Most of these
metals were leached while the anaerobic lysimeters remained in an acidic condition. In
the presence of sulfide, all of these metals were potentially precipitated with sulfide at a
neutral pH. In contrast, Al, Cu, Cr, and Pb dissolved in leachate of the aerobic lysimeters
and exhibited significantly higher concentrations than observed in the anaerobic
lysimeters. Various redox, pH and biological reactions dictated the leaching patterns of
these metals.
The waste mass losses of aerobic and anaerobic lysimeters were estimated using
leachate and gas quality data. Mass removed from the wastes was primarily converted
into gas. After waste was removed from the lysimeters, the mass of waste excavated was
compared to the estimated mass loss. A good correlation (a 1% difference between the
two values) was found from the excavated aerobic lysimeter. For the anaerobic lysimeter,
the estimated mass loss was greater than the measured mass loss excavated by 7%. These

143
mass losses were also correlated with volume loss represented by settlement of the
lysimeters. This relationship was found to be:
[settlement, %] = (AH, %) = 16.90 x log [mass loss, %] 6.24
The performance of the aerobic and anaerobic lysimeters was also evaluated by
analyzing lignocellulosic waste samples for biochemical methane potential (BMP). For
wood waste, which is typically categorized as a non-biodegradable material in landfills,
no significant influence of air injection on SYP block decomposition was found using
cellulose/lignin analysis. However, the methane yields of the SYP blocks excavated from
the aerobic lysimeter were significantly lower than those of the anaerobic lysimeter. BMP
analysis of the other lignocellulosic materials resulted in great differences in
biodegradation between the aerobic and anaerobic lysimeters.
6.2 The Implication of This Research
It may not prove surprising that waste can rapidly decompose in an aerobic
condition. It could be also anticipated that methane concentrations decreased as a result
of air addition. These results can be found from other research, and these may not be the
novel contributions of this research. However, in addition to these common consequences,
other results of this research, which could possibly occur in large scale landfill, may be
helpful for landfill operators and engineers when choosing between aerobic and
anaerobic systems.
As discussed in Chapter 4, all biodegradable wastes were not decomposed within
the test period of the aerobic lysimeter. However, Figure 2-5 in Chapter 2 showed that
BOD concentrations of the aerobic lysimeters lowered below 50 mg/L before day 300. In
comparison with the percentage of mass loss (20%) at low BOD concentrations of the

144
aerobic lysimeters, BOD concentration of lysimeter 4 was still approximately 35,000
mg/L. For the aerobic lysimeters, organic carbon dissolved in the leachate degraded
rapidly, resulting in treatment of the leachate with respect to organic carbon. The
settlement data of the aerobic lysimeters (Chapter 5) showed that constant settlement was
observed even after the BOD concentration was below 50 mg/L. This observation
indicated that waste could be continuously decomposed maintaining low BOD
concentration. Consequently, air addition may contribute to lower leachate treatment cost
as well as more waste decomposition.
Another important finding of this research lies is that alkaline pH (pH > 9.0) and
more oxidizing redox conditions can be formed by air addition. Under the oxidizing and
alkaline conditions in the aerobic lysimeters, there were greater concentrations of Al, Cr,
Cu and Pb in leachate of the aerobic lysimeters relative to those in the anaerobic
lysimeters among the eight metals under considerations (Chapter 3). A greater impact of
high pH on an increase in the solubilities of Al and Cr was observed. The high pH was
not reported from all research studying waste decomposition under aerobic conditions.
However, air addition may increase the potential to raise the pH by CO2 stripping as
discussed in Chapter 2. Among the leaching behaviors of the metals under the relatively
high pH, it is notable that the oxidation state of Cr can be converted into the hexavalent
form in aerobic landfill condition in the presence of CCA-treated wood. Song et al.
(2005) reported that Cr leached from CCA-treated wood could be oxidized to hexavalent
form under alkaline conditions. It was reasonable for Cr species to oxidize to hexavalent
form under the oxidizing condition. However, although the manufacturing of CCA-
treated wood is banned currently for most uses, it has been recognized as a safe material

145
in terms of Cr release to the environment since Cr should be reduced to trivalent form
when fixed in the wood. It was also reported that the oxidation of Cr would not happen in
landfill conditions due to the neutral pH and reducing conditions (Townsend et al., 2004).
The relationship between mass loss and volume loss would be another important
finding obtained through this research. As previously mentioned in Chapter 5, this
relationship could apply to the development of landfill settlement model assuming the
constant cross sectional area. Moreover, if it is possible to estimate the percentage of
biodegradable fraction of waste, ultimate settlement and long-term settlement could be
evaluated. More observations would be required to develop the relationship between
mass loss and volume loss since both aerobic and anaerobic lysimeters did not reach their
ultimate settlement. However, the exploration of the relationship between these two
parameters may be used to develop the landfill settlement model to describe waste
settlement more effectively.
The research results may provide insight into the various possibilities that may
occur in aerobic and anaerobic landfills. These results can be used for life-cycle
assessment to compare aerobic and anaerobic landfills. Aerobic landfills are expected to
reduce the cost for monitoring landfills and leachate treatment due to rapid waste
decomposition and low organic carbon in aerobic landfills. However, additional cost may
be required to treat the leachate with high metal concentrations and to install and operate
the facilities needed for air addition.
6.3 Conclusions
Conclusions obtained through this research can be summarized as follows:
The pH of aerobic lysimeters could increase up to 9.1

146
Air addition into the lysimeters substantially enhanced waste decomposition.
Concentrations of oxygen demanding substrates and VFAs were reduced. More
than 90% of the dissolved organic carbon was decomposed within 100 days in the
aerobic lysimeters.
Air addition could be used to recover acid-stuck landfills.
Ammonia concentrations in anaerobic lysimeters during the methanogenic phase
increased by an amount four times greater than those in the acidic phase.
High VFA concentrations did not affect mass loss of the aerobic lysimeters.
Mass loss could be estimated by gas and leachate quality. Estimated mass loss
corresponded with actual mass loss; mass losses predicted for lysimeters 2 and 4
were 4,069 g and 3,787 g, while actual mass losses were 4,044 g and 3,525 g,
respectively.
Among 8 metals (Al, As, Cu, Cr, Fe, Pb, Mn and Zn), average concentrations of
As, Fe, Mn, and Zn in the anaerobic lysimeters were significantly greater than
those of the aerobic lysimeters. In contrast, greater concentrations of Al, Cu, Cr
and Pb were found in leachate from the aerobic lysimeters.
Toxicity of some metals potentially changed with their oxidation states at given
conditions such as redox potential and pH. Cr changed its oxidation states to Cr
(VI) at alkaline pH under oxidizing conditions.
In the presence of CCA-treated wood, As concentrations proved substantially
higher than Cr and Cu concentrations in anaerobic condition; Cr and Cu
concentrations were relatively very low in the anaerobic lysimeters.

147
In the presence of CRT monitor glass, the highest Pb concentration was
observed from the aerobic lysimeters during the acidic condition.
When evaluating the total masses of metals adsorbed on lignocellulosic wastes
(office paper, newspaper, cardboard and wood blocks), greater masses of Fe, Mn,
As, A1 and Cu were found from the aerobic lysimeters.
Periods of time required for 20% settlement for the aerobic and anaerobic
lysimeters were 1 and 2 years, respectively.
The relationship between percentage of settlement and mass loss based on this
research can be written as: [settlement, %] = (AH, %) = 16.90 log [mass loss,
%] 6.24
The settlement and mass loss data obtained from this research could be used for
the development of the settlement models.
There was no significant impact of air addition on wood waste decomposition as
indicated by cellulose and lignin analysis, but methane yields of the SYP blocks
of the aerobic lysimeter were significantly lower than those of the anaerobic
lysimeters.
There was no evidence that the lignin component of the SYP blocks degraded in
either the aerobic or anaerobic lysimeters.
6.4 Future Work
The influences of air added to a landfill on gas and leachate quality, metal leaching
behavior, settlement and waste decompositions were explored. Air injection providing
various results distinguished as different from those of anaerobic landfills. Except for the
metal leaching results, much of the data obtained from this research gives validity to the

148
benefits of aerobic landfills. However, in order to apply the air injection technique to the
field, much work still needs to be done in order to solve the problems that may
potentially occur.
For landfill gas, closure of the flare system due to air addition may allow landfill
gas to emit into the atmosphere without filtration. Thus it is important to identify these
gas constituents and to evaluate the potential impact of these gases on the environment.
Moreover, it is required to explore the potential risk of explosion of landfill gas by
mixing certain concentrations of oxygen and methane (Coward and Jones, 1952).
All landfills may not have the problem with clogging of leachate collection system
(LCS), but the failure of the LCS was reported due to biological and chemical reactions
(Fleming et al., 1999). There is also potential risk of biological clogging on the geotextile
and LCS of the aerobic landfills (Reinhart and Chopra, 2000). Further research is needed
to assess the chemical and biological clogging possibly occurred in aerobic landfills in
comparison with anaerobic landfills.
It may be also required to assess the possible impact of rapid settlement of aerobic
landfills on the stability of an entire landfill. The examples of the potential risks may
include 1) the collapse or breakage of gas collection system and geomembrane on the top
of a landfill, and 2) the failure of landfill slope due to unbalanced settlement which may
occur in air injection areas.

APPENDIX A
ADDITIONAL PROCEDURES AND CONCEPTS
A.l Prediction of Mass Loss by Gas and Leachate
Once solid wastes are subject to land disposal, decomposition processes start.
Cellulolytic bacteria convert biodegradable macromolecules into simpler substrates,
which are accessible to other bacteria that produce gases and/or other byproducts (Figure
A-l). In anaerobic systems, methane and carbon dioxide are the final products of
decomposition, while water and carbon dioxide are the major products in aerobic
systems. Various biochemical mechanisms are involved in these processes; they can be
expressed by simple chemical equations as follows:
C6H,o05 + 602 -Â¥ 6C02 + 5H20
(aerobic)
(1)
C6H,o05 + H20 -> 3CH4 + 3C02
(anaerobic)
(2)
According to the equations, (1) and (2), 6 moles of CO2 are produced from 1 mole
of glucose, a monomer of cellulose in aerobic decomposition. Three moles each of CH4
and CO2 are produced from 1 mole of glucose in anaerobic decomposition. At standard
temperature and pressure (STP, temperature = 273K and pressure = 1 atm), 1 mole of
gas occupies 22.4L. If gas expansion by temperature around a gas totalizer (20C for
average) is taken into consideration, the volume occupied by 1 mole of gas at 20C is
calculated using the Ideal Gas Law:
Vx
Tx
(3)
149

150
22AL V2
273K ~ (273 + 20)K
(4)
Such that,
V2 = 24.0 L (5)
Based on information obtained thus far, the mass loss of waste by gas generation
can be calculated as follows:
Aerobic condition:
1 L C02 generated
lmole C02
lmole C6H10O5
162 g
24.0 L C02
6 mole C02
1 mole C6H10O5
= 1.125 g mass loss / L C02 generated (6)
Anaerobic condition:
1 L biogas (CFI4
and C02) generated
lmole
biogas
lmole C6H10O5
162 g
24.0 L
biogas
3 mole biogas
1 mole C6H10O5
= 2.25 g mass loss / L biogas generated (7)
In order to predict mass loss more precisely at given conditions, organic carbons
remaining in leachate are considered as well. Since many different types of
microorganisms involved in biodegradation are ecologically linked, the biogas produced
by final gas-producers may not always reflect mass loss caused by the first attackers.
For example, wastes consumed by acid-forming bacteria may not convert into methane
gas immediately during the acid-forming stage of anaerobic decomposition due to the low
activity of methane-generators. Most microorganisms may not consume large wastes at
first. Once long and branched structures of cellulose are hydrolyzed by cellulolytic
bacteria, a first attacker, other bacteria may follow and consume the fragments of
cellulose. For this reason, mass loss may take place as cellulolytic bacteria hydrolyze
wastes, increasing dissolved organic carbons in leachate. Gas-generators consume part of

151
these organic carbons, with the rest of the organic carbons remaining in leachate, which
can be measured using bulk organic parameters, such as total organic carbon (TOC).
These relationships can be expressed as shown in equation 8.
[Total mass loss of wastes (g)] = [Mass loss (g) by gas generation]
+ [Changes in TOC (g/L)*leachate remained in lysimeters (L)] (8)
Figure A-2 depicts changes in TOC, mass loss by gas generation and total mass
loss in both aerobic and anaerobic lysimeters. Since aerobic bacteria generate gas and
degrade organics rapidly, the impacts of TOC on total mass loss in aerobic lysimeters
may not be critical. Although small gaps between total mass loss and mass loss by gas
generation were observed at high TOC levels, it can be concluded that gas generation is
the main path of total mass loss. In the anaerobic lysimeters, however, total mass loss was
highly affected by both TOC and gas generation, especially during the first acid-forming
phase. The effects of TOC were then reduced as TOC concentrations decreased.
Ultimately, total mass loss of both aerobic and anaerobic lysimeters may be expressed as
mass loss by gas generation and dissolved organic carbons.
After lysimeter studies were completed, total dry mass of the actual waste was
compared to mass loss predicted using equation (8). Actual mass loss of excavated
garbage and mass loss predicted are summarized in Table A-l. For the aerobic lysimeters,
both values were within 2%. However, some discrepancies between actual mass loss and
the predicted values in the anaerobic lysimeter were observed. The most likely
explanation of this difference would be because of 1) the high methane yield of dog food
and 2) some fraction of waste was used for energy generation rather than CH4 conversion.

152
A.2 Estimation of Biodegradable Volatile Solids (BVS)
Biodegradable volatile solids (BVS) can be estimated using methane potential as
follows:
BVS(%) =
a mass of waste (g) converted into CH 4
Initial dry mass (g)
x 100
(9)
In order to calculate a mass converted into CH4, total CH4 volume generated was
corrected to CH4 volume in standard temperature and pressure (STP, 0C and 1 atm)
using the ideal gas law:
P V P V
rl V1 r2 v2
(10)
Arranging equation (10) for V2, CH4 volume at STP, and proper values for V, P
and T were substituted into equation (2), as shown in the following expression:
[CH4 generated at STP (L)] = [CH. generated(L)]x ^ x (76 ~ *2)l"mHg
(273 + 35)K 760mmHg
(11)
where, 35 K is a set temperature of BMP; 760 mm Hg is equivalent to 1 atm; and
42 mm Hg is water vapor pressure at 35C. Since gas in the serum bottle was saturated by
water, pressure was corrected by subtracting the partial pressure of water vapor.
Corrected CH4 volume was converted into moles by dividing it by 22.4 L, which is
equivalent to 1 mole of gas at STP. According to the chemical reactions presented in
equation (2), 1 mole of glucose may convert into 3 moles of CH4. However, in reality, a
part of the 1 mole of glucose is used to generate energy, and the rest of it converts into
CH4. Rittman and McCarty (2001) reported that actual CH4 mole converted from 1 mole

153
of glucose would be 2.4 moles rather than 3 moles. Thus,
into CH4 can be written as follows:
[Mass converted into CH4] =
1PH ntSTPfl IT: lmolegaS :.162gC6H10Os
[ 4 generated 22.4 L, STp 2.4m0leCH4
the mass of waste converted
(12)
Therefore, BVS percentage can be determined by substituting the mass converted
into CH4 obtained from equation (12) and initial dry mass for parameters in equation (9).
A.3 Lysimeter Dismantlement
After lysimeter studies were completed, solid wastes were excavated from aerobic
and anaerobic lysimeters. Before excavation, temperature controllers and an air injection
pumps were shut down, and all tubes and cables connected to lysimeters were removed.
Excavation of solid wastes was conducted using 4-ft long narrow-tined fork used for
gardening. A lysimeter was laid down slightly and leaned on a support. In order to collect
all excavated solid waste, a large tub was prepared and placed under the mouth of the
lysimeter. Excavated samples were collected at various depths and measured for their wet
weight. All wood blocks were then collected from the excavated samples and divided into
two fractions for the purpose of compressibility testing and garbage separation. A
fraction of excavated samples for the compressibility test was put inside black garbage
bags and stored at 4C. The other fraction of garbage samples was measured for initial
wet weight and placed in a drying oven to measure moisture content. The mass of waste
and depth of each layer are summarized in Table A-2.
Figures B-8 through B-12 show the features of solid wastes excavated. Solid waste
excavated from the aerobic lysimeter was darker than the solid waste from the anaerobic

154
lysimeters due to iron oxidation found in the aerobic lysimeters. Paper samples, which
looked undegraded, were found in the center areas of the aerobic lysimeters; solid wastes
excavated from the anaerobic lysimeters looked uniformly degraded. Wood blocks were
separated immediately the excavation of the solid waste. It proved harder to separate
wood blocks from the solid waste taken from the aerobic lysimeters because the degraded
paper remained attached to the wood (Figure B-9). Almost all of the wood samples were
too dark to recognize identification numbers written with oil-based pen. Since all paper
and wood samples remained damp and attached, parts of the samples were dried at 103C
for two days in order to effectively separate each type.

155
Table A-l. Actual mass loss and predicted values of the aerobic and anaerobic lysimeter
Actual mass loss (g)
Mass loss predicted (g)
Lys 2 (aerobic)
3,989 g
4,069 g
Lys 4 (anaerobic)
3,397 g
3,787 g
Table A-2. Mass and density of wastes excavated by de
pth
layers
depth
(from)
depth (to),
(inches)
wt(g)
wt (lb)
vol (cf)
Density
(pcf)
Density
(pcy)
2-1
21
24
1829.2
4.0
0.05
82.15
2218
24
28
2519.1
5.6
0.07
84.85
2291
28
33
2614.9
5.8
0.08
70.46
1903
2-2
33
45
5206.2
11.5
0.20
58.46
1578
2-3
45
57
4544.9
10.0
0.20
51.03
1378
2-4
57
66
4038.5
8.9
0.15
60.46
1632
4-1
22
24
1151.2
2.5
0.03
77.55
2094
24
26
1137.2
2.5
0.03
76.61
2069
26
28.5
1342.9
3.0
0.04
72.38
1954
4-2
28.5
32
1763.8
3.9
0.06
67.90
1833
32
34.5
1382.7
3.0
0.04
74.52
2012
34.5
41
3673.1
8.1
0.11
76.14
2056
4-3
41
53
5638.7
12.4
0.20
63.31
1709
4-4
53
66
6556.1
14.5
0.21
67.95
1835

156
Figure A-l. Schematic of mass loss by waste decomposition

Mass loss (g) Mass loss (g)
157
Days
(A)
Days
(B)
o
100
400

Mass loss (g) Mass loss (g)
158
Days
(C)
Days
(D)
Figure A-2. Waste mass loss by TOC and gas generation
(j
'Si)
y
o
H

APPENDIX B
SUPPLEMENTAL S
1. Peristaltic pump
2. air purging
3. leachate collection port
4. Wet-tip gas totalizer
5. Temperature controller
6. Gas-sampling bag
7. Hydraulic cylinder
8. Heating tape
9. Hydraulic jack
10. Plunger
11. Fabricated wastes
AEROBIC ANAEROBIC
LYSIMETER LYSIMETER
Figure B-l. Schematics of aerobic and anaerobic lysimeters used for this research
159

160
Top fringe
Hydraulic pressurizing unit
Figure B-2. The carriage system

161
Figure B-3. A schematic of the temperature control system

162
Figure B-4. Schematic of gas volume measuring tool; before gas measurement, fill tap-
water up to the top scale (VI close and V2 and V3 open) and close V2.
After connecting an air-sampling bag to the top, open V1 and a valve on the
air-sampling bag to let water drain out. As water drains out by gravity force,
gas in air-sampling bag transfers into the pipe and replaces with water. When
no more gas is left in air-sampling bag, water draining stops by itself. Close
V1 and measure the water volume replaced with gas.

163
yard waste
8%
food waste
17%
paper
26%
office paper
15%
newspaper
1%
cardboard
10%
plastics
15%
glass
metals 6/o
7%
Figure B-5. The nation-wide composition of discarded municipal solid waste in 2003
(EPA, 2005)
miscellaneous
12%
textile
ferrous metal
13%
food waste
9%
plastics
8%
paper
43%
office paper
6%
newspaper
9%
cardboard
14%
other paper
14%
Figure B-6. The composition of municipal solid waste in Florida in 2000 (FDEP, 2002)

164
Figure B-7. (A) Blue water phenomenon observed from gas collection system of
aerobic lysimeters; (B) a hole on copper tube caused by corrosion of Cu

165
Figure B-8. Solid samples excavated from one of the aerobic lysimeter

166

167
Figure B-10. Not well degraded office paper (aerobic lysimeter)
Figure B-l 1 Wood blocks excavated from aerobic lysimeter

168
Figure B-12. Separated newspaper and office paper

APPENDIX C
LYSIMETER EXPERIMENT RAW DATA AND GRAPHS
C.l Graphs
Figure C-l. The change in COD of the lysimeters over the percentage of mass loss.
169

170
Figure C-2. The change in BOD5 of the aerobic and anaerobic lysimeters over the
percentage of mass loss

N-NH3+(mg/L)
171
Figure C-3. The change in ammonia of the aerobic and anaerobic lysimeters over time

Fluoride (mg/L) Fluoride (mg/L)
172
Figure C-4. The change in fluoride of the aerobic and anaerobic lysimeters over time.

Chloride (mg/L) Chloride (mg/L)
173
Figure C-5. The change in chloride (Cl) of the aerobic and anaerobic lysimeters over
time.

Sulfate (mg/L) Sulfate (mg/L)
174
Figure C-6. The change in sulfate of the aerobic and anaerobic lysimeters over time

175
100 200 300 400 500 600 700
Days
Figure C-7. The change in calcium (Ca) of the aerobic and anaerobic lysimeters over time

1600
1400
1200
1000
800
600
400
200
0
14000
12000
10000
8000
6000
4000
2000
0
:e C-8
176
Lys 1
O Lys 2
*
O

o
o


ko
o*
o
o o
o
o , Qz>P-
50
100
150
200
i
250
i
300
350
Lys 3
O Lys 4
Day
800
change in sodium (Na) of the aerobic and anaerobic lysimeters over time

40
30
20
10
0
40
30
20
10
0
177
140
Lys 1
l
I:
I
1
I ;
| r
ll
ll
11
ll
is
II
II
I I
! 1 1
I I
n
I I
l I
I I
I
l
L-
i
i
i i
'V
11
I
/
-O-J
I /
1/
- 120
- 100
- 80
60
- 40
20
100
200
300
400
140
120
- 100
- 80
- 60
40
- 20
Lys 2
"I r-
ll
H
H
|!
ll
n
I I
I I
I I
I
I
i r.
trj
u
a Lrv~V""/
-^Ij'l
l|
I j l /
U i\! V
i
100
i
300
200
Days
400
?. The change in biogas produced from the aerobic lysimeters
Air injection rate (mL/min) Air injection rate (mL/min)

Gas concentrations (%) Gas concentrations (%)
178
Figure C-10. The change in biogas produced from the anaerobic lysimeters

Al concentrations, mg/L
179
4 5 6 7 8 9 10
pH
Figure C-l 1. Al concentration versus pH in leachate from the lysimeters.

Cr concentrations (mg/L)
180
Figure C-12. Cr concentration versus pH in leachate from the lysimeters

Metal concentrations (mg/L)
181
pH
Figure C-13. Cu concentration versus pH in leachate from the lysimeters

Metal concentrations (mg/L)
182
Figure C-14. Mn concentration versus pH in leachate from the lysimeters

Metal concentrations (mg/L)
183
Figure C-15. Pb concentration versus pH in leachate from the lysimeters

Metal concentrations (mg/L)
184
pH
Figure C-16. Zn concentration versus pH in leachate from the lysimeters

100
80
60
40
20
0 <
100
80
60
40
20
0 i
re C
mim \ hhhi
185
2-2 CB
10 20 30 40
50
Days
. Change in methane yields of the waste layer 2-1 and 2-2

80
60
40
20
0
100
80
60
40
20
0
eC
IHHHf \ HHHI
186
2-3 CB
2-3 NP
2-3 OP
2-3 wood
cellulose
Blank
Newspaper
Jk 1
'i
10
- -rTU
0 ; $=&
20 30 40 50
2-4 CB
2-4 NP
2-4 OP
2-4 wood
cellulose
Blank
Newspaper
Col 1 vs Blank
^ 1
=?=g~T ^
10 20 30 40 50
Days
. Change in methane yields of the waste layer 2-3 and 2-4

80
60
40
20
0
100
80
60
40
20
0
eC
187
Days
. Change in methane yields of the waste layer 4-1 and 4-2

Cumulative methane volume (mL) Cumulative methane volume (mL)
188
Figure C-20. Change in methane yields of the waste layer 4-3 and 4-4

189
C.2 Raw Data
Table C-l.
dH of the aerobic and anaerobic
lysimeters
date
lys 1
lys 2
date
lys 3
lys 4
7/28/2004
5.669
5.695
8/8/2003
4.546
4.868
8/4/2004
5.29
5.373
8/13/2003
5.342
5.457
8/12/2004
5.14
5.07
8/15/2003
5.237
5.321
8/18/2004
5.264
4.862
8/19/2003
4.971
5.308
8/24/2004
5.45
4.562
8/22/2003
5.157
6.063
8/31/2004
5.448
4.747
8/26/2003
5.54
5.926
9/10/2004
5.389
4.747
9/9/2003
5.31
5.49
9/15/2004
5.25
4.569
9/12/2003
5.239
5.471
9/23/2004
5.316
4.395
9/16/2003
5.318
5.447
10/8/2004
5.66
4.871
9/19/2003
5.233
5.303
10/19/2004
5.663
4.9
9/24/2003
5.186
5.262
10/26/2004
5.471
4.66
9/26/2003
5.159
5.205
11/6/2004
5.525
4.995
9/29/2003
5.262
5.257
11/13/2004
5.442
5.05
10/3/2003
5.196
5.26
11/24/2004
5.396
5.329
10/8/2003
5.206
5.242
12/8/2004
5.156
5.509
10/10/2003
5.146
5.193
12/20/2004
5.006
5.42
10/15/2003
5.195
5.243
1/5/2005
5.373
5.855
10/17/2003
5.212
5.195
1/7/2005
5.4
5.94
10/21/2003
5.206
5.213
1/9/2005
5.296
5.932
10/26/2003
5.246
5.23
1/10/2005
5.494
5.946
10/28/2003
5.143
5.171
1/13/2005
5.758
6.149
10/31/2003
5.158
5.158
1/17/2005
7.474
6.16
11/5/2003
5.142
5.165
1/21/2005
7.14
6.45
11/7/2003
5.151
5.177
1/25/2005
7.404
6.962
11/12/2003
5.177
5.221
1/30/2005
8.279
8.646
11/15/2003
5.154
5.172
2/5/2005
8.71
8.82
11/19/2003
5.11
5.16
2/8/2005
8.82
8.87
11/26/2003
5.094
5.1
2/16/2005
8.88
8.86
11/28/2003
5.038
5.051
2/24/2005
9.165
9.024
12/3/2003
5.288
5.357
3/1/2005
8.973
9.045
12/10/2003
5.088
5.151
3/12/2005
8.974
9.015
12/17/2003
5.161
5.19
3/20/2005
8.85
8.89
12/19/2003
5.11
5.204
3/26/2005
8.908
8.926
12/23/2003
5.143
5.179
4/3/2005
8.972
8.883
1/2/2004
5.214
5.255
4/10/2005
8.998
8.87
1/7/2004
5.08
5.274
4/17/2005
8.982
8.851
1/14/2004
5.251
5.271
4/18/2005
9.035
8.774
1/16/2004
5.126
5.193

190
Table C-l (continued'
date
lys 1
lys 2
date
lys 3
lys 4
4/30/2005
9.01
8.814
1/22/2004
5.222
5.298
5/7/2005
9.161
8.752
1/29/2004
5.189
5.295
5/16/2005
8.855
8.563
2/12/2004
5.157
5.182
5/23/2005
8.839
8.665
2/25/2004
5.138
5.206
5/31/2005
8.497
8.573
3/16/2004
5.14
5.168
6/6/2005
8.564
8.549
3/19/2004
5.093
5.134
6/14/2005
8.59
8.55
5/9/2004
5.05
5.094
6/28/2005
8.4
8.555
6/1/2004
5.214
5.246
7/5/2005
8.586
8.518
7/28/2004
5.069
5.198
7/12/2005
8.587
8.552
8/4/2004
5.129
5.046
7/19/2005
8.77
8.78
8/12/2004
5.167
5.203
7/27/2005
8.545
8.621
8/18/2004
5.235
5.233
8/10/2005
8.518
8.533
8/24/2004
5.203
5.161
8/31/2004
5.239
5.182
9/10/2004
5.264
5.178
9/15/2004
5.15
5.078
9/23/2004
5.273
5.148
10/8/2004
5.582
5.43
10/19/2004
5.61
5.432
10/26/2004
5.343
5.206
11/6/2004
5.448
5.252
11/13/2004
5.442
5.276
11/24/2004
5.521
5.308
12/8/2004
5.527
5.352
12/13/2004
6.2
5.8
12/16/2004
6.2
5.7
12/20/2004
6.153
6.264
1/5/2005
5.8
5.8
1/13/2005
5.875
5.97
1/17/2005
5.532
5.54
1/21/2005
5.8
5.95
1/25/2005
5.892
5.982
1/30/2005
5.855
5.993
2/5/2005
5.93
6.24
2/8/2005
5.92
6.31
2/16/2005
5.89
6.471
2/24/2005
5.881
6.7
3/1/2005
5.897
6.761
3/12/2005
5.95
6.9
3/20/2005
6.04
7.18
3/26/2005
6.102
7.292
4/3/2005
6.071
7.25
4/10/2005
6.105
7.485
4/17/2005
6.116
7.338
4/24/2005
7.182
7.377
4/30/2005
6.873
7.392

191
Table C-l (continued
date
lys 1
lys 2
date
lys 3
lys 4
5/7/2005
6.822
7.436
5/16/2005
6.724
7.305
5/23/2005
6.658
7.363
5/31/2005
6.622
7.597
6/6/2005
6.544
7.357
6/14/2005
6.493
7.368
6/28/2005
6.501
7.431
7/12/2005
6.459
7.471
7/19/2005
6.57
7.46
7/27/2005
6.468
7.362
8/10/2005
6.459
7.372

192
Table C-2. Conductivity of the aerobic and anaerobic lysimeters (unit: fas)
date
lys 1
lys 2
date
lys 3
lys 4
10/8/2004
17800
1900
8/8/2003
12540
28000
10/26/2004
15200
1700
8/13/2003
17300
17900
11/6/2004
17500
2600
8/15/2003
16400
16500
11/13/2004
16100
5400
8/19/2003
14560
15380
11/24/2004
10500
7000
8/22/2003
13630
14890
12/8/2004
12000
2000
8/26/2003
16100
16200
12/20/2004
2500
4700
9/9/2003
17300
17600
1/5/2005
12200
15300
9/12/2003
17200
17600
1/17/2005
8600
12100
9/16/2003
15440
15300
1/25/2005
11200
7600
9/19/2003
19680
1302
1/30/2005
7600
3300
9/24/2003
18200
18000
2/5/2005
7400
2200
9/26/2003
21900
20200
2/8/2005
6800
5800
9/29/2003
20100
19600
2/16/2005
8100
6600
10/3/2003
22400
> 20000
2/24/2005
7600
9800
10/8/2003
16900
19300
3/1/2005
5300
9200
10/10/2003
19000
> 20000
3/12/2005
7100
8300
10/15/2003
19700
22800
3/20/2005
7500
7800
10/17/2003
18600
19400
3/26/2005
6500
7600
10/21/2003
19000
22300
4/3/2005
7500
7600
10/26/2003
14800
16600
4/10/2005
7600
7400
10/31/2003
17100
22600
4/17/2005
8500
6800
11/5/2003
16800
18600
4/18/2005
8300
6800
11/7/2003
17200
19000
4/30/2005
10000
7000
11/12/2003
19400
19000
5/7/2005
9400
7200
11/15/2003
17400
22400
5/16/2005
7800
6600
11/19/2003
17000
19500
5/23/2005
8000
6300
11/26/2003
18000
5/31/2005
8400
3900
11/28/2003
18000
22100
6/6/2005
8200
6800
12/3/2003
17600
18100
6/14/2005
8700
6500
12/10/2003
16000
18500
6/28/2005
8400
7600
12/17/2003
16000
17800
7/5/2005
8000
6100
12/23/2003
14230
15940
7/12/2005
7700
7200
1/2/2004
15400
16000
7/19/2005
6700
6200
1/7/2004
19200
1/14/2004
18600
17520
2/12/2004
12400
12400
2/25/2004
10910
11040
3/16/2004
11700
11400
3/19/2004
12200
14500
6/1/2004
13500
11100
10/8/2004
16300
13600
10/19/2004
16500
15100
10/26/2004
15900
15300
11/6/2004
17200
15800
11/13/2004
16500
17200
11/24/2004
19200
14000

193
Table C-2 (continued)
date
lys 1
lys 2
date
lys 3
lys 4
12/13/2004
> 20000
15500
12/20/2004
18200
1/5/2005
15700
13300
1/17/2005
14500
12600
1/25/2005
19200
18000
1/30/2005
17200
16100
2/5/2005
18400
14500
2/8/2005
14900
2/16/2005
18900
19000
2/24/2005
> 20000
18000
3/1/2005
19200
> 20000
3/12/2005
18800
> 20000
3/20/2005
19900
19900
3/26/2005
19400
> 20000
4/3/2005
> 20000
> 20000
4/10/2005
> 20000
> 20000
4/17/2005
> 20000
> 20000
4/24/2005
19700
> 20000
4/30/2005
> 20000
> 20000
5/7/2005
17200
> 20000
5/16/2005
23200
23200
5/23/2005
22400
22600
5/31/2005
21800
21700
6/6/2005
22000
21000
6/14/2005
24600
23800
6/28/2005
25900
22900
7/12/2005
25600
23600
7/19/2005
24800
22700

194
Table C-3. Alkalinity
of aerobic and anaerobic lysimeters (unit: mg/L as CaCCb)
date
lys 1
lys 2
date
lys 3
lys 4
7/28/2004
5000
3500
8/26/2003
12,600.00
8/4/2004
4500
2600
9/9/2003
15,000.00
8/12/2004
4250
3000
9/16/2003
10,500.00
8/18/2004
6000
2200
10/8/2003
18,333.33
12,857.14
8/24/2004
9500
80
10/21/2003
12,142.86
12,857.14
9/10/2004
10750
700
10/28/2003
10,000.00
14,285.71
10/8/2004
16050
750
11/5/2003
12,000.00
14,000.00
10/19/2004
16000
11/12/2003
12,000.00
14,000.00
10/26/2004
11250
500
11/19/2003
10,500.00
12,000.00
11/6/2004
14000
1000
11/26/2003
8,500.00
10,500.00
11/13/2004
12300
2000
12/3/2003
11200
15000
11/24/2004
10500
2200
12/10/2003
10200
11000
12/8/2004
7050
5750
12/17/2003
12500
13000
12/20/2004
2969
3750
12/24/2003
10000
10800
1/5/2005
6000
8000
1/2/2004
11000
12000
1/25/2005
3690
2340
1/7/2004
12000
14000
2/5/2005
570
456
1/14/2004
7000
12500
2/16/2005
720
900
1/22/2004
11500
13000
2/24/2005
880
2000
1/29/2004
6000
13500
3/1/2005
700
1750
2/25/2004
12000
3/12/2005
700
1250
3/12/2004
6000
3/20/2005
1000
1400
3/16/2004
8000
11500
3/26/2005
1000
1300
7/28/2004
11000
11500
4/3/2005
1000
1100
8/4/2004
10000
11000
4/10/2005
1000
1100
8/12/2004
10000
11500
4/17/2005
1880
1000
8/18/2004
10000
11000
4/24/2005
1640
1000
8/24/2004
7500
8000
4/30/2005
2120
1000
9/10/2004
8000
9500
5/7/2005
2200
1000
10/8/2004
14550
12300
5/16/2005
2200
1000
10/19/2004
13500
12000
5/23/2005
2200
1000
10/26/2004
12150
11700
5/31/2005
2280
960
11/6/2004
12000
10500
6/6/2005
2280
1000
11/13/2004
12000
12075
6/14/2005
2240
1040
11/24/2004
12750
11250
6/28/2005
2400
1000
12/8/2004
13500
10950
7/12/2005
2200
1000
12/20/2004
13500
12750
7/27/2005
2200
1000
1/5/2005
14000
10710
8/10/2005
2040
1000
1/25/2005
13500
10200
2/5/2005
12880
12880
2/16/2005
14620
14790
2/24/2005
15000
15000
3/1/2005
14700
13650
3/12/2005
14900
15750
3/20/2005
15150
15000
3/26/2005
16300
16200

195
Table C-3 (continued)
date
lys 1
lys 2
date
lys 3
lys 4
4/3/2005
16500
16500
4/10/2005
15500
15150
4/17/2005
15750
16050
4/24/2005
13650
15750
4/30/2005
14550
14850
5/7/2005
14700
15000
5/16/2005
13200
14250
5/23/2005
13050
13500
5/31/2005
12000
12750
6/6/2005
13200
13350
6/14/2005
13800
14250
6/28/2005
13800
12900
7/12/2005
13500
12750
7/27/2005
12600
12300
8/10/2005
13050
12300

196
Table C-4 Total dissolved solids (TPS) of aerobic and anaerobic lysimeters (unit: mg/L)
date
lys 1
lys 2
date
lys 3
lys 4
7/28/2004
4.00
4.00
8/13/2003
37.62
48.26
8/4/2004
16.44
12.72
8/15/2003
44.41
37.62
8/17/2004
15.32
6.04
8/22/2003
48.16
53.78
9/9/2004
43.84
2.98
8/26/2003
48.31
50.70
10/8/2004
39.92
0.86
9/9/2003
49.63
49.06
10/19/2004
46.02
0.70
9/16/2003
50.12
52.31
10/26/2004
41.12
11.02
9/24/2003
48.42
53.94
11/6/2004
49.26
7.88
10/21/2003
43.04
54.84
11/13/2004
39.42
6.28
10/28/2003
27.54
43.64
1/25/2005
15.54
6.20
7/28/2004
32.00
30.00
1/30/2005
2.26
8/4/2004
28.72
30.66
5-Feb
6.12
1.48
8/17/2004
30.64
29.06
16-Feb
7.16
4.64
9/9/2004
32.60
37.72
24-Feb
5.50
8.48
10/8/2004
27.46
35.42
1-Mar
4.38
6.84
10/19/2004
33.92
35.00
12-Mar
5.74
6.18
10/26/2004
41.08
36.26
20-Mar
6.42
6.10
11/6/2004
32.22
33.70
26-Mar
5.68
6.32
11/13/2004
38.10
36.28
3-Apr
6.46
6.14
1/25/2005
34.08
27.66
10-Apr
6.86
6.52
1/30/2005
25.74
17-Apr
8.32
5.90
5-Feb
33.04
25.48
24-Apr
8.14
4.88
16-Feb
33.34
27.14
30-Apr
8.1
6.26
24-Feb
35.30
28.54
7-May
8.88
6.70
1-Mar
31.46
23.34
16-May
8.14
6.48
12-Mar
34.04
29.14
23-May
8.18
6.42
20-Mar
35.82
26.30
31-May
7.9
6.28
26-Mar
33.78
29.32
6-Jun
8.82
7.08
3-Apr
35.20
25.88
14-Jun
7.76
6.82
10-Apr
32.42
25.8
28-Jun
7.34
6.32
17-Apr
33.98
27.98
5-Jul
6.82
7.28
24-Apr
30.86
24.72
12-Jul
5.275
5.8
30-Apr
30.84
25.7
10-Aug
6.5
7.68
7-May
33.44
26.12
16-May
32.92
24.48
23-May
30.66
23.72
31-May
30.42
22.52
6-Jun
31.62
23.48
14-Jun
29.84
22.42
28-Jun
28.22
21.94
5-Jul
29.54
20.96
12-Jul
28.60
19.78
10-Aug
27.84
17.28

197
Table C-5. Total organic contents (TOC) o
date
lys 1
lys 2
date
lys 3
8/9/2004
6288
0
8/8/2003
26266
21771
8/12/2004
6227
7268
8/13/2003
24492
22581
8/17/2004
8053
5219
8/15/2003
21679
25649
8/9/2004
6418
8/19/2003
28491
28005
8/24/2004
7238
5847
8/26/2003
16688
18640
9/15/2004
18738
4678
9/2/2003
18480
19881
10/5/2004
18881
3363
9/12/2003
18148
18088
10/9/2004
1447
9/19/2003
19209
24694
10/26/2004
17985
1614
10/3/2003
19431
23817
11/19/2004
5205
1496
10/10/2003
18166
21096
12/12/2004
7178
6634
10/17/2003
17519
22649
1/6/2005
12977
13166
10/26/2003
18357
22802
1/13/2005
5617
10720
10/31/2003
15581
21267
1/21/2005
6864
11/12/2003
14397
1/30/2005
630
1997
11/19/2003
14112
16858
2/5/2005
1671
845
11/26/2003
13627
16335
2/16/2005
1454
1392
12/3/2003
14428
16779
2/24/2005
1359
2102
12/10/2003
13523
18360
3/12/2005
1525
1833
12/17/2003
13362
16588
3/20/2005
1565
1614
12/23/2003
13337
17636
3/26/2005
1394
1771
1/2/2004
13158
16128
4/3/2005
1673
2337
1/14/2004
18156
20742
4/17/2005
2225
1562
1/22/2004
14387
7892
4/24/2005
1935
1203
1/29/2004
17721
12777
4/30/2005
2451
1160
2/12/2004
13306
17529
5/7/2005
2382
3548
2/26/2004
12293
11076
5/16/2005
2428
1671
3/12/2004
14975
17391
5/31/2005
3718
2993
3/16/2004
14464
18522
6/6/2005
4233
3186
6/1/2004
13041
16700
6/28/2005
2469
1419
8/9/2004
13961
17174
7/5/2005
3016
8/12/2004
16567
15254
7/11/2005
3432
2102
8/17/2004
12648
16307
7/19/2005
3894
2905
8/9/2004
12983
16356
7/27/2005
2296
8/24/2004
14460
16431
8/10/2005
2639
2228
9/15/2004
14762
16696
10/5/2004
14238
16406
10/26/2004
15047
17252
11/19/2004
15995
19933
12/12/2004
13760
13121
1/6/2005
14561
13351
1/13/2005
16461
12668
1/21/2005
12567
12499
1/30/2005
16806
11683
2/5/2005
16595
12062
2/16/2005
17044
11606
2/24/2005
11526
aerobic and anaero
3C lysimeter (unit: mg/L)

198
Table C-5
continued)
date
lys 1
lys 2
date
lys 3
lys 4
3/12/2005
14576
10648
3/20/2005
15017
9954
3/24/2005
9636
3/26/2005
15017
9203
4/3/2005
13539
4/17/2005
13484
9526
4/24/2005
12010
9368
4/30/2005
9303
4/30/2005
13470
5/7/2005
13302
8205
5/16/2005
12293
8637
5/23/2005
12580
8490
6/6/2005
12990
6/14/2005
12505
6/28/2005
12930
3231
7/5/2005
13160
6493
7/11/2005
11760
6775
7/19/2005
7990
7/27/2005
12230
5625
8/10/2005
7685

199
Table C-6 Chemical oxygen demand (COD) of aerobic and anaerobic lysimeter (unit:
r mg/L) ,
Date
Lys 1
Lys 2
Date
Lys 3
Lys 4
7/28/2004
6950
11250
8/8/2003
64745
66941
8/12/2004
21900
27850
8/13/2003
77801
62846
8/17/2004
34050
26600
8/15/2003
59820
58277
8/24/2004
22700
21200
8/19/2003
78988
82964
9/3/2004
64950
24400
8/26/2003
69671
77267
9/10/2004
69250
19600
9/2/2003
72876
74656
9/15/2004
66750
18900
9/12/2003
72876
77030
9/23/2004
66850
19250
9/19/2003
73593
83079
10/5/2004
66249
13464
10/3/2003
67983
81600
10/26/2004
61302
5814
10/10/2003
65943
79560
11/3/2004
70788
5406
10/17/2003
65892
84150
11/11/2004
66453
15402
10/26/2003
62934
81498
11/19/2004
70635
7293
10/31/2003
61710
80631
11/24/2004
50337
16371
11/12/2003
57936
75939
12/1/2004
42228
29631
11/19/2003
58956
74460
12/8/2004
34884
24939
11/26/2003
58497
74970
12/12/2004
30498
23766
12/3/2003
56559
73899
1/6/2005
38900
38850
12/10/2003
56253
73287
1/13/2005
18600
32700
12/17/2003
55233
70890
1/21/2005
13650
12/23/2003
58191
69870
2/5/2005
4100
1450
1/2/2004
49929
68085
2/8/2005
4000
3700
1/14/2004
56661
60996
2/24/2005
3400
6850
1/22/2004
51561
64821
3/12/2005
3650
4300
1/29/2004
54621
63648
3/20/2005
4600
4500
2/22/2004
58089
62577
3/26/2005
4600
4500
2/26/2004
52785
61302
4/3/2005
5050
5250
7/28/2004
61450
67850
4/17/2005
6700
3900
8/12/2004
50400
56600
4/26/2005
5890
4440
8/17/2004
52650
55850
4/30/2005
6700
4100
8/24/2004
54100
57650
5/7/2005
8550
5000
9/3/2004
53150
57850
5/16/2005
6800
4350
9/10/2004
53150
58150
5/23/2005
7350
4350
9/15/2004
53350
58050
5/31/2005
7290
5320
9/23/2004
51250
56450
6/6/2005
6730
5250
10/5/2004
50694
58701
6/14/2005
7230
5160
10/26/2004
52122
53652
6/28/2005
5880
4610
11/3/2004
41922
56610
11/11/2004
50337
52734
11/19/2004
52632
52326
11/24/2004
52632
52836
12/1/2004
52479
54519
12/8/2004
51408
51000
12/12/2004
48246
49572
1/6/2005
52000
39900
1/13/2005
52350
38850

200
Table C-6
continued)
1/21/2005
52350
37850
2/5/2005
48550
38650
2/8/2005
49600
36900
2/24/2005
51200
40150
3/12/2005
51500
36500
3/20/2005
52050
37000
3/26/2005
64850
33100
4/3/2005
55650
32550
4/17/2005
48650
32150
4/26/2005
40850
28650
4/30/2005
43000
29950
5/7/2005
45050
28650
5/16/2005
44250
27450
5/23/2005
45000
26850
5/31/2005
39250
22650
6/6/2005
41750
24400
6/14/2005
43800
23400
6/28/2005
42050
21400

201
Table C-7. NH3+-N concentrations of aerobic and anaerobic lysimeters
Date
Lys 1
Lys 2
Date
Lys 3
Lys 4
1
44.5737
26.0448
13
123.1292
101.564
13
52.794
19
129.5536
129.0838
16
43.9773
51.5659
29
129.0838
144.4716
21
98.4427
41.2616
40
158.783
174.5122
28
125.3132
43.1491
47
176.4246
179.6586
38
153.8117
83.6837
55
198.1744
232.5246
45
125.3132
62.3102
69
216.2289
250.9585
50
184.8154
112.8914
76
199.6195
262.141
58
166.1884
70.1399
82
196.0262
278.8412
70
146.9202
95.2447
89
201.0752
282.9228
74
126.157
144.5331
97
184.286
283.9525
91
114.8975
80.0651
104
173.8794
363.5242
107
144.0877
106.5606
118
191.2202
383.5107
115
187.1865
206.4054
125
201.2484
310.1469
120
189.5881
142.6142
132
151.6302
342.1821
138
125.5446
141.4073
139
190.4698
335.5207
159
179.0542
148.5193
145
198.8885
332.8926
160
338.2198
196.4142
155
196.5562
326.412
163
268.6915
395.4994
162
215.9736
221.551
178
334.7243
210.4829
225
178.0782
178.0782
187
102.4254
183.2859
363
156.0991
196.3619
193
166.1915
100.2522
375
171.3963
207.5146
204
256.9294
144.1649
378
166.373
207.5146
212
255.886
217.4478
383
204.8859
228
448.6748
230.1956
390
202.743
244.4261
236
152.6165
162.2223
400
354.8838
347.6131
242
168.9598
179.5944
407
228.8206
255.2256
250
194.031
181.062
412
451.3427
313.1798
257
303.5795
214.8094
420
328.1665
339.5144
264
348.6262
219.2248
432
364.9485
325.3893
271
289.1112
220.1187
436
211.072
212.8735
277
402.0369
266.8583
453
299.062
329.7673
286
404.2729
270.5288
469
543.3762
507.6597
293
342.8068
290.9748
477
236.5871
267.615
300
215.5594
259.967
482
605.1256
600.0046
308
356.8128
231.5037
500
664.4257
770.9723
314
350.5038
198.0456
521
209.6439
240.1807
322
319.1664
242.0624
522
608.4553
426.5548
336
301.1875
193.6781
525
923.2783
505.5622
343
506.625
324.996
540
919.5288
1068.944
349
496.8555
295.9966
549
912.0755
1348.001
357
502.6943
269.5848
555
868.6071
365
512.5787
265.4179
566
915.7946
1252.792
379
510.5864
253.2997
574
861.5665
1252.792
590
997.4957
1567.038
598
854.583
1432.845
604
1017.999
1612.317

202
Table C-7 (continuec
D
Date
Lys 1
Lys 2
Date
Lys 3
Lys 4
612
886.4613
1444.554
619
879.276
1310.144
626
993.4448
1468.26
633
872.1489
1409.711
639
837.3708
1257.9
648
669.5123
977.0133
655
702.8366
1046.08
662
771.0181
883.8794
670
654.4303
718.6858
676
799.8811
943.3915
684
716.8409
730.0174
698
699.709
886.2881
705
926.4855
880.748
711
970.8101
857.0626
719
1053.542
771.5203
727
1061.779
780.5869
741
1029.21
817.9315

203
Table C-8. Sulfide concentrations of aerobic and anaerobic lysimeters (uniti^g/L)
date
lys 1
lys 2
date
lys 3
lys 4
8/4/2004
335
65
8/13/2003
305
310
8/12/2004
165
25
8/15/2003
485
8/17/2004
230
40
8/26/2003
29
8/24/2004
110
0
8/29/2003
36
8/31/2004
175
5
9/9/2003
190
290
9/9/2004
145
45
9/16/2003
150
180
9/14/2004
110
0
9/24/2003
160
180
9/22/2004
40
0
9/29/2003
160
180
10/1/2004
60
0
10/8/2003
100
200
10/5/2004
65
20
10/15/2003
95
155
10/15/2004
120
30
10/21/2003
120
150
10/21/2004
230
10
10/28/2003
80
155
10/31/2004
125
45
11/5/2003
85
130
11/12/2004
95
15
11/12/2003
80
145
11/19/2004
65
5
11/19/2003
85
120
12/2/2004
45
55
11/26/2003
90
155
1/13/2005
345
125
12/3/2003
75
110
1/22/2005
820
645
12/10/2003
85
145
1/30/2005
1375
500
12/17/2003
65
135
2/8/2005
1063
563
12/19/2003
50
100
2/16/2005
425
575
1/2/2004
55
130
2/24/2005
125
1250
1/7/2004
65
125
3/1/2005
225
905
1/22/2004
80
110
3/12/2005
275
540
1/29/2004
70
85
3/20/2005
530
725
2/10/2004
88
136
3/25/2005
500
775
2/25/2004
80
120
4/3/2005
705
350
3/12/2004
120
108
4/13/2005
625
3/14/2004
46
117
4/25/2005
1450
900
6/1/2004
28
73
5/3/2005
2375
1225
7/28/2004
75
90
5/12/2005
2575
1025
8/4/2004
65
105
6/1/2005
2650
1075
8/12/2004
10
75
6/8/2005
2375
1075
8/17/2004
50
95
6/16/2005
2500
1150
8/24/2004
40
35
6/29/2005
2450
875
8/31/2004
45
75
7/11/2005
2275
875
9/9/2004
80
40
7/19/2005
2200
950
9/14/2004
35
75
7/27/2005
1950
900
9/22/2004
0
50
10/1/2004
15
60
10/5/2004
15
90
10/15/2004
40
90
10/21/2004
30
55
10/31/2004
115
140
11/12/2004
15
11/19/2004
35
105
12/2/2004
50
50

204
Table C-8 (continued)
date
lys 1
lys 2
date
lys 3
lys 4
12/20/2004
0
65
1/13/2005
15
100
1/22/2005
0
355
1/30/2005
45
340
2/8/2005
85
400
2/16/2005
35
445
2/24/2005
55
485
3/1/2005
60
545
3/12/2005
55
700
3/20/2005
95
740
3/25/2005
75
1205
4/3/2005
75
975
4/13/2005
0
1014
4/25/2005
80
1165
5/3/2005
60
1420
5/12/2005
100
1350
6/1/2005
115
2400
6/8/2005
95
2525
6/16/2005
135
2400
6/29/2005
155
2450
7/11/2005
180
2150
7/19/2005
250
2350
7/27/2005
325
2750

205
Table C-9. Volatile fa
tty acids (VFA) of lysimeter 1
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/18/2004
25045.1
695.9
2160.3
5020.9
9/22/2004
13103.4
1421.7
4984.8
9269.7
9/30/2004
10662.5
1166.9
4304.6
8284.1
10/21/2004
14481.1
1457.8
5599.2
9500.1
11/3/2004
13667.2
1310.8
5194.9
9088.5
12/12/2004
6856.0
773.9
2681.6
5399.9
12/20/2004
3581.9
419.9
1532.8
2810.4
1/22/2005
3303.1
353.9
2515.4
5576.7
1/30/2005
117.4
21.1
166.1
404.0
2/8/2005
14.8
0.0
10.3
30.2
2/16/2005
7.9
0.0
3.7
4.1
3/3/2005
43.3
41.3
19.5
29.8
3/12/2005
58.5
0.0
21.9
36.5
3/25/2005
31.0
42.9
18.7
27.5
4/3/2005
55.2
45.8
19.4
29.4
4/17/2005
35.0
0.0
1.5
3.8
4/24/2005
24.1
0.0
0.0
0.0
4/30/2005
12.6
2.0
0.0
3.7
5/7/2005
32.1
1.7
1.6
3.7
5/16/2005
9.9
0.0
0.0
3.8
5/23/2005
3.9
3.0
1.9
3.7
5/31/2005
6.3
2.8
1.5
3.7
6/6/2005
5.2
2.1
1.5
4.3
6/14/2005
9.9
2.5
0.0
4.6
6/24/2005
6.4
2.5
1.6
4.0
7/5/2005
0.4
0.0
2.2
4.9
7/11/2005
0.9
0.0
1.8
4.5
7/19/2005
2.0
0.0
2.4
6.1

206
Table C-10.
Volatile fatty acids
rVFA) of lysimeter 2
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/18/2004
11166.0
123.6
553.1
1801.7
9/22/2004
2935.9
92.1
506.9
1095.2
9/30/2004
2335.6
84.2
470.3
1024.8
10/21/2004
1432.7
74.0
432.5
1002.9
11/3/2004
1033.7
66.8
415.4
899.2
12/12/2004
3467.0
767.1
1257.8
6229.2
12/20/2004
2726.1
618.7
1120.2
5420.8
1/22/2005
693.0
99.1
1057.2
3258.2
1/30/2005
124.8
19.3
206.7
611.9
2/8/2005
2.7
0.0
3.3
10.8
2/16/2005
7.5
0.0
3.8
7.1
3/3/2005
42.5
0.0
19.2
29.0
3/12/2005
44.0
0.0
19.5
30.3
3/25/2005
36.5
43.2
0.0
27.6
4/3/2005
33.6
42.9
18.8
28.2
4/17/2005
7.9
0.0
2.1
3.7
4/24/2005
8.3
2.1
1.5
0.0
4/30/2005
13.7
2.5
2.8
6.6
5/7/2005
5.3
0.0
1.6
3.7
5/16/2005
8.4
2.7
1.8
4.6
5/23/2005
5.7
1.8
1.6
3.7
5/31/2005
2.7
2.8
1.7
3.7
6/6/2005
3.1
0.0
1.6
4.8
6/14/2005
2.6
2.3
1.5
3.7
6/24/2005
1.9
1.8
0.0
4.0
7/5/2005
0.7
0.0
1.7
4.6
7/11/2005
0.6
3.5
1.8
4.7
7/19/2005
0.9
0.0
2.9
5.8

207
Table C-ll.
Volatile fatty acids
rVFA) of lysimeter 3
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/7/2003
1554.1
245.3
0.0
25.2
8/13/2003
5827.3
250.0
1552.1
2356.0
8/15/2003
6130.6
217.2
1633.6
2420.7
8/19/2003
6053.6
215.0
1725.8
2642.0
8/22/2003
4858.8
190.3
1871.4
3027.2
8/26/2003
3132.3
189.0
1807.9
3108.4
9/9/2003
3662.6
185.9
1775.5
2794.8
10/8/2003
7937.9
766.0
3099.2
4090.7
10/15/2003
9727.6
814.7
3484.5
4468.6
10/22/2003
9846.3
770.7
3252.5
4111.8
11/5/2003
10064.4
3984.7
3676.7
4501.6
11/15/2003
11155.3
3995.7
3716.8
4409.8
11/19/2003
10719.7
3849.9
3630.4
4283.9
11/28/2003
11690.5
4238.9
3970.9
4558.8
12/10/2003
10418.4
3681.0
3603.5
4082.4
12/19/2003
8760.7
3471.0
4544.1
6219.2
1/5/2004
7523.3
2792.3
3649.2
4777.0
1/14/2043
8602.9
3084.2
4010.1
5133.5
1/8/2004
10141.2
3784.2
4975.7
6531.1
2/13/2004
9903.0
3553.1
4898.4
6098.9
3/16/2004
10625.4
3616.4
5025.0
6182.5
7/17/2004
13562.0
3304.9
3533.5
3583.1
8/18/2004
27827.8
4940.0
4545.7
6098.8
9/22/2004
17197.7
3640.1
5805.2
6827.0
9/30/2004
12572.9
2739.3
4210.0
5343.0
10/21/2004
11580.7
2639.4
3930.8
5135.3
11/3/2004
12671.8
2809.7
4037.1
5397.7
12/12/2004
8997.0
3211.7
4022.8
5578.8
12/20/2004
14545.6
3897.8
4411.4
6253.5
1/22/2005
15118.1
3327.8
3976.9
5180.8
1/30/2005
16654.3
4341.5
4405.8
5347.6
2/8/2005
18671.3
4486.5
4509.3
5667.4
2/16/2005
16358.8
3847.6
4212.3
5383.6
3/3/2005
14300.5
3576.5
3598.3
6033.2
3/12/2005
14220.9
3209.0
3108.9
5614.0
3/25/2005
14513.6
3534.5
3495.5
5950.6
4/3/2005
14191.5
3391.0
3240.1
5696.3
4/17/2005
13386.3
2435.1
2837.4
5459.3
4/24/2005
12372.5
2138.5
2408.6
4692.6
4/30/2005
11690.6
2100.7
2295.6
4397.6
5/7/2005
13138.9
2411.2
2569.5
4926.4
5/16/2005
13708.8
2693.4
2825.1
5358.1
5/23/2005
31195.0
5051.0
5536.2
9693.8
5/31/2005
12812.7
2619.3
2634.6
4941.7
6/6/2005
12600.7
2630.5
2579.5
4771.7
6/14/2005
14119.6
2921.4
2853.6
5442.2

208
Table C-ll
continuec
)
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
6/14/2005
14119.6
2921.4
2853.6
5442.2
6/24/2005
11563.2
2668.4
2571.0
4663.6
7/5/2005
7374.9
2194.9
1865.6
3339.8
7/11/2005
8675.3
2771.5
2149.7
3920.2
7/19/2005
7789.0
3316.3
2432.0
3744.9

209
Table C-12 Volatile fatty acids (VFA) of lysimeter 4 (unit: mg/L)
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/7/2003
2694.5
243.4
0.0
97.1
8/13/2003
5911.9
216.3
1323.3
2605.4
8/15/2003
5931.9
206.6
1346.7
2585.0
8/19/2003
4163.9
220.4
1317.0
2776.7
8/22/2003
1668.6
193.0
1476.3
3338.6
8/26/2003
1793.1
187.8
1473.3
3285.1
9/9/2003
2880.9
171.0
1686.9
3271.3
10/8/2003
5258.6
759.0
3006.2
4820.7
10/15/2003
6366.8
743.4
3018.9
4803.3
10/22/2003
7897.2
761.1
3205.4
4977.5
11/5/2003
7383.7
3775.5
3558.7
5437.6
11/15/2003
8051.3
3653.4
3452.6
5241.6
11/19/2003
8961.9
3837.7
3580.4
5374.4
11/28/2003
9574.0
3973.5
3703.3
5476.4
12/10/2003
10036.6
4004.2
3754.1
5496.6
12/19/2003
6580.8
3043.4
3907.1
10782.7
1/5/2003
8193.8
3405.5
4473.5
11310.0
1/8/2003
7685.9
3100.9
3937.7
10264.8
1/14/2003
7559.2
2617.0
3356.1
8451.9
2/13/2003
8839.9
2995.2
4048.9
9055.6
3/16/2004
9964.5
3469.7
4734.5
10038.3
7/17/2004
15855.6
3891.7
3845.4
4636.0
8/18/2004
22080.7
3290.6
3601.4
5822.7
9/22/2004
12112.4
2824.2
4058.2
6519.4
9/30/2004
11624.9
2706.9
3989.4
6413.2
10/21/2004
14603.2
3118.2
4725.8
7205.4
11/3/2004
12304.7
2711.0
3912.8
6141.4
12/8/2004
13600.4
3576.6
3977.6
5854.4
12/12/2004
11090.6
3256.8
3250.8
5184.9
12/20/2004
12577.5
3733.5
3017.6
5102.5
1/22/2005
13159.3
3351.2
2018.2
3570.5
1/30/2005
16013.2
4858.1
2757.5
4242.8
2/8/2005
16170.3
5019.1
2499.3
4130.4
2/16/2005
15610.6
5419.8
2553.6
4078.9
3/3/2005
11124.0
5202.6
1615.2
2950.1
3/12/2005
10597.2
4919.0
1298.3
2249.0
3/25/2005
8711.5
4910.3
1086.6
1672.3
4/3/2005
8461.1
5422.0
1171.0
1735.4
4/17/2005
9927.2
5086.0
1039.4
1864.3
4/24/2005
8648.7
4514.9
856.9
1490.3
4/30/2005
8194.4
4767.2
937.3
1453.1
5/7/2005
6624.6
4145.4
776.0
1144.7
5/16/2005
8062.8
5347.5
942.1
1201.3
5/23/2005
6288.2
4771.3
839.7
894.0
5/31/2005
4314.8
3642.3
630.1
498.2
6/6/2005
5017.1
4667.2
773.4
546.1

210
Table C-12. (continued)
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
6/14/2005
4462.3
4249.4
630.0
247.1
6/24/2005
4883.7
6142.8
782.9
87.4
7/5/2005
2960.0
5356.2
623.3
51.5
7/11/2005
1712.2
3791.7
423.2
32.9
7/19/2005
1729.3
4328.7
333.5
79.8

211
Table C-13. Fabricated waste in lysimeters (unit: g)
CB
NP
cullets
CRT
A1
Plastic
lys 1 fraction 1
544.4
205.5
170.1
34
136.4
510.6
lys 1- fraction 2
544.3
204.1
170.1
34
136.1
510.3
lys 1- fraction 3
544.3
204.1
170.1
34
136.2
510.4
lys 1- fraction 4
544.4
204.3
170.1
34
136.1
510.3
lys 2- fraction 1
544.3
204.1
170.1
34
136.2
510.3
lys 2- fraction 2
544.3
204.2
170.1
34
136.2
510.3
lys 2- fraction 3
544.3
204.1
170.1
34
136.1
510.3
lys 2- fraction 4
544.3
204.3
170.2
34
136.2
510.3
Lys 3 Fraction 1
544.5
204.1
188.2
34
136.7
510.4
Lys 3 Fraction 2
544.1
204.5
187.8
34
136.6
510.7
Lys 3 Fraction 3
544.5
204.4
186.6
34
136.3
510.6
Lys 3 Fraction 4
544.2
204.1
187.4
34
136.2
510.6
Lys 4 Fraction 1
544.7
204.8
187.7
34
136.5
510
Lys 4 Fraction 2
544.2
204.3
187.1
34
136.2
510
Lys 4 Fraction 3
544.4
204.6
187
34
136
510.7
Lys 4 Fraction 4
544.2
204
187.1
34
136
510
steel
OP
dogfood
SYP
CCA
Total
lys 1 fraction 1
136.1
952.5
510.8
171
34.1
3405.0
lys 1 fraction 2
136.2
952.5
510.4
170
34
3402.2
lys 1- fraction 3
136.2
952.5
510.4
170
34
3402.5
lys 1- fraction 4
136.2
952.5
510.4
171
34.1
3402.9
lys 2- fraction 1
136.1
952.5
510.3
170
34.1
3402.2
lys 2- fraction 2
136.1
952.5
510.3
170
34
3402.1
lys 2- fraction 3
136.2
952.5
510.3
170
34
3402.0
lys 2- fraction 4
136.2
952.5
510.3
171
34
3402.8
Lys 3 Fraction 1
136.2
953
510
170
34
3420.9
Lys 3 Fraction 2
136.5
953
510
170
34
3421.0
Lys 3 Fraction 3
136.6
953
510
170
34
3419.8
Lys 3 Fraction 4
136
952.6
510
170
34
3419.0
Lys 4 Fraction 1
136.4
953
510
170
34
3421.0
Lys 4 Fraction 2
136.4
953
510
170
34
3419.4
Lys 4 Fraction 3
136.3
953.5
510
170
34
3420.9
Lys 4 Fraction 4
136
953
510
170
34
3417.8

212
Table C-14. Metal concentrations of lysimeter 1 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
7/28/2004
2.27
0.08
0.02
0.32
110.94
3.64
0.03
28.58
8/29/2004
2.94
0.24
0.12
3.31
90.13
4.03
0.75
97.71
9/2/2004
1.28
0.45
0.18
0.11
7.47
2.30
0.03
95.27
9/3/2004
2.86
0.39
0.10
2.05
74.80
4.25
0.27
132.47
9/10/2004
4.45
0.59
0.16
8.81
115.63
6.24
0.79
209.81
9/15/2004
3.22
0.42
0.10
8.91
79.81
4.33
0.63
172.08
10/7/2004
3.96
0.43
0.10
82.61
5.15
0.55
214.32
10/14/2004
6.73
0.39
0.13
0.99
47.40
6.65
0.12
244.32
10/19/2004
4.84
0.25
0.09
6.50
29.69
5.34
0.58
214.90
11/6/2004
4.33
0.28
0.11
7.23
65.77
6.99
0.46
203.98
11/24/2004
2.93
0.19
0.07
6.45
50.69
5.45
0.50
137.60
12/19/2004
2.05
0.16
0.04
1.49
89.08
2.84
0.08
82.72
1/6/2005
1.17
0.11
0.04
0.24
45.10
4.12
0.03
69.49
1/21/2005
0.80
0.12
0.05
0.51
2.31
1.10
0.00
5.52
1/30/2005
3.45
0.26
0.11
1.32
1.68
0.08
0.02
2.55
2/5/2005
7.68
0.30
0.17
4.48
2.74
0.09
0.03
6.46
2/16/2005
5.47
0.32
0.11
4.79
1.08
0.04
0.01
4.09
2/24/2005
6.76
0.37
0.12
5.56
1.29
0.03
0.04
3.83
3/1/2005
6.91
0.31
0.10
3.42
0.90
0.03
0.01
3.52
3/12/2005
6.44
0.35
0.11
2.90
1.13
0.03
0.01
4.09
3/20/2005
7.49
0.37
0.13
2.14
1.71
0.04
0.01
5.00
3/26/2005
10.28
0.40
0.20
3.64
2.21
0.08
0.06
4.96
4/3/2005
11.50
0.50
0.22
1.73
2.09
0.06
0.02
6.83
4/17/2005
11.80
0.71
0.33
1.79
4.85
0.12
0.06
10.83
4/24/2005
11.77
0.59
0.27
1.65
3.58
0.08
0.05
9.43
4/30/2005
12.19
0.74
0.43
1.79
6.03
0.14
0.06
14.88
5/7/2005
14.05
0.80
0.43
2.17
5.31
0.11
0.05
15.03
5/16/2005
13.82
0.77
0.45
1.85
6.56
0.14
0.05
16.79
5/31/2005
10.74
0.72
0.45
0.91
8.14
0.19
0.03
13.70
6/6/2005
10.79
0.72
0.44
0.60
7.12
0.18
0.02
14.27
6/14/2005
10.81
0.55
0.37
0.61
6.94
0.19
0.02
13.13
6/28/2005
9.52
0.38
0.32
1.13
7.67
0.17
0.03
11.85
7/5/2005
8.40
0.30
0.26
1.31
7.42
0.18
0.04
11.31
7/11/2005
9.02
0.30
0.23
0.95
6.62
0.18
0.02
10.70
7/27/2005
8.12
0.25
0.26
0.60
6.17
0.14
0.02
10.31
8/10/2005
9.37
0.29
0.28
0.79
6.82
0.14
0.02
10.56

213
Table C-15. Metal concentrations of lysimeter 2 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
7/28/2004
1.32
0.08
0.02
0.70
111.65
5.37
0.02
56.49
8/9/2004
4.54
0.03
0.03
0.36
88.62
2.67
0.12
93.43
8/29/2004
10.66
0.02
0.12
7.46
98.22
1.43
1.68
109.50
9/3/2004
10.35
0.01
0.08
2.52
32.56
0.76
0.25
35.76
9/17/2004
0.46
0.12
0.08
0.44
1.50
1.31
0.04
39.90
10/7/2004
8.69
0.01
0.06
3.43
23.45
0.56
0.31
31.35
10/14/2004
10.86
0.01
0.08
1.78
35.85
0.69
0.16
35.11
10/19/2004
8.97
0.01
0.06
2.22
20.94
0.49
0.25
37.84
11/6/2004
2.63
0.02
0.02
1.45
22.36
0.45
0.26
15.66
11/11/2004
6.22
0.04
0.09
5.07
40.55
1.32
0.63
17.47
12/8/2004
6.33
0.09
0.13
4.58
63.06
3.56
0.59
33.93
12/19/2004
2.24
0.03
0.04
1.26
51.02
0.95
0.22
12.90
1/6/2005
4.79
0.12
0.15
2.25
65.76
3.18
0.23
35.85
1/13/2005
1.44
0.11
0.13
1.76
37.96
2.36
0.14
22.78
2/16/2005
7.80
0.41
0.19
1.42
2.39
0.03
0.01
3.02
2/24/2005
11.96
0.54
0.31
1.66
6.63
0.07
0.04
6.82
3/1/2005
12.42
0.80
0.32
4.07
3.93
0.05
0.06
7.45
3/12/2005
9.08
0.51
0.21
2.45
2.45
0.03
0.02
4.71
3/20/2005
6.59
0.50
0.20
1.09
2.49
0.03
0.05
4.98
3/26/2005
9.28
0.70
0.23
1.52
2.70
0.05
0.03
7.36
4/3/2005
11.21
0.84
0.25
1.99
4.15
0.07
0.03
7.40
4/5/2005
10.34
0.13
0.11
0.80
3.78
0.03
0.10
3.60
4/17/2005
5.99
0.51
0.19
1.06
2.09
0.05
0.01
5.19
4/24/2005
16.58
0.82
0.24
0.79
3.38
0.09
0.03
7.50
4/30/2005
8.28
0.60
0.25
1.03
2.70
0.06
0.02
5.93
5/7/2005
6.37
0.59
0.25
1.42
2.70
0.06
0.01
6.25
5/16/2005
7.62
0.71
0.29
1.40
3.26
0.08
0.03
7.55
5/31/2005
9.30
0.79
0.33
1.06
3.66
0.10
0.01
8.50
6/6/2005
6.44
0.66
0.27
1.63
2.87
0.07
0.01
8.56
6/14/2005
6.65
0.66
0.28
1.27
2.99
0.07
0.01
8.64
6/28/2005
9.30
0.63
0.24
0.76
2.94
0.09
0.01
8.21
7/5/2005
4.96
0.59
0.24
1.09
2.93
0.07
0.03
8.80
7/11/2005
7.38
0.69
0.28
1.44
3.31
0.10
0.05
10.72
7/27/2005
5.56
0.61
0.24
0.84
2.72
0.07
0.01
8.46
8/10/2005
7.13
0.85
0.35
1.16
5.20
0.10
0.02
12.40

214
Table C-16. Metal concentrations of lysimeter 3 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
8/13/2003
5.04
2.14
0.42
0.08
20.59
6.45
0.02
77.06
8/15/2003
10.57
2.10
0.40
0.06
21.03
7.83
0.01
87.07
8/19/2003
20.35
1.89
0.42
0.05
31.37
7.86
0.06
87.99
8/22/2003
20.27
2.04
0.46
0.04
40.74
9.05
0.03
97.37
8/26/2003
9.26
2.09
0.38
0.05
40.01
9.18
0.02
103.86
9/9/2003
3.66
2.35
0.27
0.03
32.39
8.40
0.02
124.89
9/17/2003
2.61
2.72
0.24
0.03
31.84
8.80
0.03
155.53
9/24/2003
1.87
2.57
0.18
0.02
27.68
7.87
0.02
167.62
10/9/2003
1.35
2.62
0.15
0.03
26.36
7.80
0.04
181.27
10/15/2003
1.23
2.65
0.14
0.02
26.85
7.95
0.03
183.65
10/21/2003
1.23
2.72
0.14
0.03
27.91
8.03
0.03
182.82
10/28/2003
0.55
2.98
0.15
0.03
30.76
8.39
0.02
399.65
11/5/2003
0.53
2.91
0.13
0.06
30.86
8.23
0.03
398.20
11/12/2003
0.47
2.75
0.12
0.02
30.80
7.90
0.02
389.54
11/19/2003
0.44
2.74
0.11
0.02
33.28
7.83
0.03
386.99
11/26/2003
0.43
2.75
0.10
0.03
34.45
7.97
0.02
390.47
12/3/2003
0.44
2.80
0.10
0.02
38.23
8.13
0.03
389.78
12/10/2003
0.32
1.97
0.09
0.02
31.30
8.00
0.02
387.52
12/17/2003
0.38
2.64
0.09
0.02
41.92
7.78
0.04
378.75
12/23/2003
0.42
2.80
0.09
0.02
49.61
8.17
0.03
388.76
1/2/2004
0.40
2.67
0.08
0.01
53.79
7.99
0.03
384.01
1/9/2004
0.42
2.86
0.08
0.02
60.94
8.55
0.03
396.12
1/14/2004
0.41
2.66
0.08
0.01
62.34
8.02
0.03
380.51
1/22/2004
0.38
2.59
0.07
0.01
72.81
8.06
0.03
381.30
2/12/2004
0.36
2.22
0.07
0.01
85.17
7.76
0.01
403.33
3/16/2004
0.40
2.58
0.07
0.01
114.72
8.70
0.03
391.24
3/26/2004
0.41
2.74
0.07
0.01
103.28
8.91
0.04
397.52
6/1/2004
0.25
1.13
0.03
0.00
120.84
5.40
0.03
290.20
7/28/2004
0.31
0.91
0.03
0.00
192.79
5.37
0.03
289.10
8/9/2004
0.07
0.53
0.02
0.00
138.54
3.52
0.03
194.27
8/12/2004
0.20
0.86
0.03
0.01
253.34
5.70
0.05
269.86
8/17/2004
0.39
0.81
0.03
0.00
278.07
5.61
0.05
287.40
8/24/2004
0.19
0.78
0.04
0.01
301.47
5.66
0.09
262.24
9/3/2004
0.26
0.70
0.02
0.00
293.19
5.13
0.05
259.62
9/10/2004
0.11
0.60
0.02
0.01
269.31
4.50
0.04
217.89
9/15/2004
0.30
0.75
0.02
0.00
346.35
5.55
0.06
269.70
9/23/2004
0.60
1.26
0.04
0.02
534.21
8.08
0.09
263.15
10/7/2004
0.20
0.57
0.02
0.00
306.63
4.20
0.05
203.07
10/14/2004
0.00
0.37
0.01
0.00
217.10
2.91
0.03
140.47
10/19/2004
0.12
0.44
0.01
0.00
244.93
3.18
0.03
157.73

215
Table C-16 (continued)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
10/26/2004
0.48
1.11
0.03
0.02
539.75
6.61
0.09
227.72
11/6/2004
0.12
0.48
0.01
0.00
287.44
3.41
0.04
160.68
11/11/2004
0.45
1.13
0.03
0.02
597.42
6.50
0.10
220.40
11/24/2004
0.10
0.47
0.01
0.00
312.82
3.25
0.04
145.12
12/8/2004
0.39
1.02
0.03
0.02
585.70
5.77
0.09
193.25
1/6/2005
0.07
0.34
0.02
0.00
292.58
2.21
0.03
84.95
1/13/2005
0.30
0.80
0.05
0.02
591.98
4.27
0.08
135.46
1/21/2005
0.06
0.35
0.02
0.00
300.24
2.13
0.03
79.30
2/5/2005
0.16
0.72
0.04
0.01
539.54
3.69
0.06
121.26
2/16/2005
0.19
0.83
0.04
0.02
517.02
3.41
0.06
100.37
2/24/2005
0.26
0.83
0.04
0.02
486.62
3.17
0.06
94.80
3/1/2005
0.34
0.94
0.04
0.02
519.69
3.29
0.06
94.12
3/12/2005
0.15
0.89
0.04
0.02
468.42
2.94
0.05
85.92
3/20/2005
0.23
0.97
0.04
0.02
465.67
2.84
0.05
79.82
3/26/2005
0.14
0.90
0.04
0.02
441.47
2.69
0.04
80.03
4/3/2005
0.15
0.88
0.04
0.02
407.69
2.44
0.04
69.79
4/17/2005
0.12
0.84
0.04
0.02
356.34
2.11
0.04
60.49
4/24/2005
0.17
0.33
0.02
0.01
58.06
3.29
0.03
57.00
4/30/2005
0.18
0.29
0.02
0.02
89.49
3.22
0.03
67.41
5/7/2005
0.07
0.26
0.02
0.02
179.21
2.66
0.02
63.49
5/31/2005
0.24
0.49
0.04
0.02
335.52
2.47
0.04
61.55
6/6/2005
0.03
0.46
0.04
0.01
253.10
1.78
0.02
54.65
6/14/2005
0.00
0.41
0.03
0.00
211.36
1.37
0.02
45.09
6/28/2005
0.00
0.51
0.03
0.00
208.26
1.24
0.01
39.54
7/5/2005
0.04
0.67
0.03
0.00
220.61
1.29
0.03
37.57
7/11/2005
0.05
0.70
0.04
0.00
208.28
1.21
0.02
33.97
7/27/2005
0.06
0.72
0.04
0.00
165.58
0.88
0.01
23.89

216
Table C-17. Metal concentrations of lysimeter 4 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
8/3/2003
9.36
2.47
0.64
0.23
20.55
6.75
0.02
78.37
8/13/2003
10.82
2.77
0.72
0.26
23.00
7.61
0.01
86.55
8/15/2003
12.97
2.47
0.64
0.22
20.80
7.25
0.01
81.52
8/19/2003
18.42
2.53
0.65
0.19
29.50
7.18
0.04
83.08
8/22/2003
10.57
2.61
0.58
0.16
30.88
7.26
0.02
79.50
8/26/2003
5.79
2.83
0.52
0.14
34.59
7.69
0.02
84.37
9/9/2003
2.70
2.17
0.37
0.08
32.60
6.66
0.03
85.14
9/17/2003
3.15
2.67
0.42
0.18
38.39
7.96
0.04
103.91
9/24/2003
3.70
2.99
0.41
0.08
32.52
8.33
0.02
138.19
10/9/2003
3.11
3.01
0.43
0.09
34.50
8.84
0.04
152.48
10/15/2003
2.57
2.79
0.39
0.06
31.39
8.43
0.06
156.50
10/21/2003
2.80
3.24
0.43
0.09
36.51
9.69
0.09
166.70
10/28/2003
1.11
3.23
0.45
0.08
35.60
9.47
0.03
332.44
11/5/2003
0.99
3.19
0.42
0.06
34.97
9.74
0.03
365.16
11/12/2003
0.88
3.12
0.40
0.07
35.35
9.80
0.02
376.90
11/19/2003
0.71
2.89
0.34
0.06
33.88
9.40
0.02
389.22
11/26/2003
0.71
3.04
0.34
0.05
36.16
9.97
0.03
402.26
12/3/2003
0.65
3.18
0.34
0.05
38.87
10.60
0.02
417.96
12/10/2003
0.40
2.18
0.29
0.04
29.52
9.35
0.02
392.13
12/17/2003
0.46
2.89
0.29
0.03
39.36
9.85
0.03
406.29
12/23/2003
0.54
2.97
0.27
0.04
44.25
10.11
0.02
418.35
1/2/2004
0.49
2.77
0.26
0.03
47.92
9.61
0.02
407.58
1/9/2004
0.48
2.80
0.24
0.03
52.53
9.85
0.02
415.38
1/14/2004
0.45
2.68
0.22
0.03
56.31
9.38
0.02
406.26
1/22/2004
0.42
2.60
0.21
0.03
63.99
9.20
0.02
407.18
2/12/2004
0.43
2.50
0.20
0.03
85.93
9.59
0.01
455.51
3/16/2004
0.36
2.20
0.17
0.02
77.91
8.14
0.02
380.16
3/26/2004
0.44
2.66
0.21
0.03
89.91
9.87
0.02
422.14
6/1/2004
0.42
1.45
0.10
0.01
89.47
6.59
0.02
353.83
7/28/2004
0.38
1.27
0.08
0.01
111.89
5.99
0.02
328.98
8/9/2004
0.00
0.28
0.02
0.00
244.59
1.70
0.03
60.70
8/12/2004
0.23
1.12
0.10
0.02
129.38
5.66
0.12
290.65
8/17/2004
0.29
0.88
0.06
0.02
116.11
4.58
0.01
263.51
8/24/2004
0.22
1.11
0.10
0.02
158.65
5.84
0.02
288.44
9/3/2004
0.23
0.73
0.05
0.00
117.19
3.99
0.01
234.78
9/10/2004
0.21
1.09
0.08
0.01
194.40
6.17
0.02
300.74
9/15/2004
0.21
0.68
0.05
0.00
127.40
3.91
0.01
228.76
9/23/2004
0.16
0.98
0.07
0.01
197.70
5.77
0.02
286.15
10/7/2004
0.17
0.60
0.04
0.00
150.94
3.75
0.01
212.72
10/14/2004
0.23
1.06
0.08
0.01
288.81
6.66
0.03
309.69
10/19/2004
0.17
0.58
0.04
0.00
165.30
3.71
0.01
211.07
10/26/2004
0.04
0.65
0.05
0.00
200.54
4.26
0.02
227.80
10/30/2004
0.00
0.70
0.02
0.00
82.41
0.71
0.01
31.18
11/6/2004
0.22
0.60
0.04
0.00
203.20
3.99
0.02
220.41
11/11/2004
0.08
0.66
0.04
0.01
281.63
4.81
0.03
239.09
11/24/2004
0.20
0.63
0.04
0.00
363.98
4.46
0.03
210.94

217
Table C-17 (continued)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
12/8/2004
0.10
0.78
0.05
0.01
432.43
4.82
0.04
207.50
1/6/2005
0.06
0.71
0.03
0.00
216.03
1.89
0.02
84.38
1/13/2005
0.03
1.11
0.03
0.01
247.93
2.14
0.04
90.68
1/21/2005
0.03
1.10
0.03
0.01
151.48
1.29
0.01
56.25
2/5/2005
0.00
1.31
0.03
0.02
123.73
1.07
0.01
43.28
2/16/2005
0.00
1.26
0.04
0.02
80.33
0.67
0.01
23.90
2/24/2005
0.00
1.25
0.05
0.03
66.61
0.55
0.02
18.83
3/1/2005
0.12
1.47
0.07
0.07
57.44
0.47
0.01
14.51
3/12/2005
0.00
0.72
0.06
0.04
22.78
0.18
0.00
6.42
3/20/2005
0.00
0.65
0.09
0.05
17.27
0.13
0.00
5.33
3/26/2005
0.03
0.58
0.11
0.06
15.20
0.12
0.01
4.81
4/3/2005
0.03
0.55
0.11
0.06
13.75
0.09
0.00
4.11
4/17/2005
0.02
0.47
0.10
0.05
11.66
0.08
0.05
3.78
4/24/2005
0.24
0.72
0.14
0.09
21.29
0.10
0.02
5.12
4/30/2005
0.09
0.52
0.13
0.07
12.10
0.08
0.01
4.61
5/7/2005
0.01
0.31
0.09
0.04
7.63
0.05
0.00
2.89
5/16/2005
0.13
0.51
0.15
0.08
11.25
0.07
0.01
4.43
5/31/2005
0.16
0.38
0.14
0.09
9.31
0.05
0.00
4.07
6/6/2005
0.24
0.41
0.19
0.08
8.27
0.05
0.00
4.95
6/14/2005
0.31
0.44
0.17
0.09
8.46
0.05
0.00
5.59
6/28/2005
0.21
0.33
0.15
0.07
5.83
0.03
0.00
4.37
7/5/2005
0.25
0.32
0.17
0.08
6.14
0.03
0.00
4.64
7/11/2005
0.27
0.32
0.18
0.09
6.53
0.08
0.00
4.90
7/27/2005
0.45
0.31
0.20
0.11
6.80
0.03
0.01
5.74
8/10/2005
0.39
0.27
0.19
0.11
5.62
0.03
0.01
5.74

218
Table C-18. ANOVA results of metals and organic absorbence
A1 (mg/L)
As
Cr
Cu
CB
aerobic
Avg
108.627
0.266
0.576
0.603
anaerobic
Avg
18.536
0.092
0.351
0.216
P
2.6E-4
8.22E-08
4.7E-4
2.72E-09
NP
aerobic
Avg
91.152
0.223
0.412
0.486
anaerobic
Avg
33.307
0.112
0.427
0.317
P
0.012
0.001
0.744
0.023
OP
aerobic
Avg
52.442
0.128
0.254
0.222
anaerobic
Avg
22.368
0.086
0.252
0.174
P
0.025
0.007
0.925
0.040
PL
aerobic
Avg
7.584
0.019
0.051
0.026
anaerobic
Avg
2.010
0.023
0.071
0.053
P
0.003
0.443
0.076
0.015
WD
aerobic
Avg
3.326
0.047
0.068
0.097
anaerobic
Avg
1.555
0.026
0.159
0.031
P
0.150
0.048
0.001
0.004
Fe
Mn
Pb
Zn
CB
aerobic
Avg
66.453
1.175
0.125
21.936
anaerobic
Avg
21.938
0.206
0.117
15.392
P
7.52E-05
1.13E-08
0.701
0.031
NP
aerobic
Avg
42.885
0.801
0.070
14.747
anaerobic
Avg
22.349
0.227
0.180
20.363
P
0.044
0.002
0.060
0.068
OP
aerobic
Avg
41.314
0.648
0.061
13.558
anaerobic
Avg
33.843
0.301
0.348
17.405
P
0.242
0.001
0.008
0.064
PL
aerobic
Avg
2.196
0.071
0.010
1.173
anaerobic
Avg
8.273
0.073
0.035
5.074
P
0.008
0.942
0.001
0.001
WD
aerobic
Avg
26.832
0.206
0.017
4.261
anaerobic
Avg
10.274
0.050
0.022
5.027
P
0.131
0.031
0.487
0.556

LIST OF REFERENCES
Agdag, O. N. and Sponza D. T., 2004, Effect of aeration on the performance of a
simulated landfilling reactor stabilizing municipal solid wastes, Journal of
Environmental Science and Health Part a-Toxic/Hazardous Substances &
Environmental Engineering 39(11-12), 2955-2972.
Akhtar, M., Blanchette, R. A. and Kirk, T. K., 1997, Fungal Delignification and
Biochemical pulping of wood. Advances in Biochemical Engineering &
Biotechnology 57, 159-195.
Alymore, M. G., 2001, Treatment of a refractory gold-copper sulfide concentrate by
copper ammoniacal thiosulfate leaching. Minerals Engineering 14(6), 615-637.
American Public Health Association (APHA), American Water Works Association
(AWWA), Water Environment Federation (WEF), 1995, Standard Methods for the
examination of water and wastewater, American Public Health Association,
Washington D.C., USA.
American Society for Testing and Materials (ASTM), 1992, Standard Test Method for
Determining the Anaerobic Biodegradation Potential of Organic Chemicals,
Annual Book of ASTM standard Vol. 11.01, 897-901.
Arzutug, M. E., Kocakerim, M. M. and Copur, M., 2004, Leaching of malachite ore in
NH3-saturated water. Industrial & Engineering Chemistry Research 43(15), 4118-
4123.
Audsley, E., Alber, S., Clift, R., Cowell, S., Crettaz, P., Gaillard, Hausheer, J., Jolliet, O.,
Kleijn, R., Mortensen, B., D., Roger, E., Teulon, H., Weidema, B., van Zeijts, H.,
1997, Harmonisation of environmental life cycle assessment for agriculture, final
report, Concerted Action AIR3-CT94-2028, European Commission DG VI,
Brussels, Belgium.
Babel, S. and Kumiawan, T. A., 2003, Low-cost adsorbents for heavy metals uptake from
contaminated water: a review. Journal of Hazardous Materials 97(1-3), 219-243.
Barlaz, M. A., Ham, R. K. and Schaefer, D. M., 1992, Microbial, chemical and methane
production characteristics of anaerobically decomposed refuse with and without
leachate recycling. Waste Management & Research 10(3), 257-267.
Barlaz, M.A., Ph.D. Dissertation, 1988, University of Wisconsin at Madison.
219

220
Basso, M. C., Cerrella, E. G. and Cukierman, A. L., 2002, Lignocellulosic materials as
potential biosorbents of trace toxic metals from wastewater. Industrial &
Engineering Chemistry Research, 41(15), 3580-3585.
Benjamin, M. M., 2002, Water chemistry. Boston, McGraw-Hill.
Berge, N. D., Reinhart, D. R. and Townsend, T. G., 2005, The fate of nitrogen in
bioreactor landfills. Critical Reviews in Environmental Science and Technology
35(4), 365-399.
Bissen, M. and Frimmel, F. H., 2003, Arsenic a review. part 1: occurrence, toxicity,
speciation, mobility. Acta Hydrochimica Et Hydrobiologica 31(1), 9-18.
Bjamgard, A. B. and Edgers, L., 1990, Settlement of municipal solid waste landfills,
Proceedings of the 13th Annual Madison Waste Conference, University of
Wisconsin, Madison, WI, 192-205.
Bleiker, D. E., Farquhar, G. and Mcbean, E., 1995, Landfill Settlement and the Impact on
Site Capacity and Refuse Hydraulic Conductivity, Waste Management & Research,
13(6), 533-554.
Borglin, S. E., Hazen, T. C., Oldenburg, C. M. and Zawislanski, P. T., 2004, Comparison
of aerobic and anaerobic biotreatment of municipal solid waste, Journal of the Air
& Waste Management Association, 54(7), 815-822.
Bozkurt, S., Moreno, L. and Neretnieks, I., 1999, Long-term fate of organics in waste
deposits and its effect on metal release, The Science of the Total Environment, 228,
135-152.
Bradl, H. B., 2005, Heavy metals in the environment: [origin, interaction and
remediation]. Amsterdam ; Boston, Elsevier Academic Press.
Brown, A., 1985, Review of lignin in biomass. Journal of Applied biochemistry 7, 371-
387.
Calli, B., Mertoglu, B., Inane, B. and Yenigun, O., 2005, Community changes during
start-up in methanogenic bioreactors exposed to increasing levels of ammonia,
Environmental Technology, 26(1), 85-91.
Carbonell-Barrachina, J., DeLaune, A., Patrick, R. D., Burlo, W. H., Sirisukhodom, F.,
and Anurakpongsatom, P., 1999, The influence of redox chemistry and pH on
chemically active forms of arsenic in sewage sludge-amended soil. Environment
International 25(5), 613-618.
Chandler, J. A., Jewell, W. J., Gossett, J. M., Vansoest, P. J. and Robertson, J. B., 1980,
Predicting Methane Fermentation Biodegradability, Biotechnology and
Bioengineering 22, 93-107.

221
Charlatchka, R. and Cambier, P., 2000, Influence of reducing conditions on solubility of
trace metals in contaminated soils. Water Air and Soil Pollution 118(1-2), 143-167.
Chynoweth, D. P. and Isaacson, R., 1987, Anaerobic digestion of biomass. London ; New
York, Elsevier Applied Science.
Chynoweth, D. P., Jerger, D. E. and Srivastava, V. J., 1985, Biological gasification of
woody biomass, Proceedings of the 20th intersociety Energy Conversion
Engineering Conference, Warrendale, PA, Society of Automotive Engineers, Inc. 1,
573-579.
Coward, H. F. and Jones, G. W., 1952, Limits of flammability of gases and vapors.
Bulletin 503, U.S. Bureau of Mines.
Critchley, M. M., Pasetto, R. and O'Halloran, R. J., 2004, Microbiological influences in
'blue water' copper corrosion, Journal of Applied Microbiology 97(3), 590-597.
Cummings, S. P. and Stewart, C. S., 1994, Newspaper as a Substrate for Cellulolytic
Landfill Bacteria, Journal of Applied Bacteriology, 76(2), 196-202.
Das, B. M., 2002, Principles of geotechnical engineering, Pacific Grove, CA, Brooks
Cole/Thompson Learning.
De Baere, L. and Verstrate, W., 1984, High rate anaerobic composting with biogas
recovery, Biocycle, 25, 30-31.
De Boer, I. J. M., 2003, Environmental impact assessment of conventional and organic
milk production. Livestock Production Science 80(1-2), 69-77.
De. Baere, L., 1984, High rate dry anaerobic composting process for the organic fraction
of solid waste, 7th Symposium on Biotechnology for Fuel and Chemicals,
Gatlinburg, Tennessee, 1984.
Drever, J. I., 1988, The geochemistry of natural waters. Englewood Cliffs, N.J., Prentice
Hall.
Dubey, B. and Townsend, T., 2004, Arsenic and lead leaching from the waste derived
fertilizer ironite. Environmental Science & Technology 38(20), 5400-5404.
Eary, L. E., 1999, Geochemical and equilibrium trends in mine pit lakes, Applied
Geochemistry, 14(8), 963-987.
Edwards, M., Jacobs, S. and Taylor, R. J., 2000, The blue water phenomenon, Journal
American Water Works Association, 92(7), 72-82.
El-Fadel, M. and Khoury, R., 2000, Modeling settlement in MSW landfills: a critical
review, Critical Reviews in Environmental Science and Technology, 30(3), 327-
361.

222
El-Fadel, M, Findikakis, A. N. and Leckie, J. O., 1989, A numerical-model for methane
production in managed sanitary landfills, Waste Management & Research, 7(1), 31-
42.
Fang, M., J. Wong, W. C, Ma, K. K. and Wong, M. H., 1999, Co-composting of sewage
sludge and coal fly ash: nutrient transformations, Bioresource Technology, 67(1),
19-24.
Fannin, K., 1987, Anaerobic digestion of biomass, edited by Chynoweth, P. D., and
Roinsaacson, Elsevier applied science, New York, USA
Faure, G., 1991, Principles and applications of inorganic geochemistry : a comprehensive
textbook for geology students, McMillan Pub. Co., New York
Fleming, I. R., Rowe, R. K. and Cullimore, D. R., 1999, Field observations of clogging in
a landfill leachate collection system, Canadian Geotechnical Journal, 36(4), 685-
707.
Florida Department of Environmental Protection (FDEP), 2002, Solid Waste Mangement
Annual Report, Tallahassee, FL, USA.
Fox, M. and Noike, T., 2004, Wet oxidation pretreatment for the increase in anaerobic
biodegradability of newspaper waste, Bioresource Technology, 91(3): 273-281.
Gerritse, R., Vriesema, G., R., Dalenberg, J. W. and Deroos, FI. P., 1982, Effect of
sewage-sludge on trace-element mobility in soils, Journal of Environmental Quality,
11(3), 359-363.
Gibson, R. E. and Lo, K. Y., 1961, A theory of consolidation for soils exhibiting
secondary compression, ACTA Polytechnic Scandianavica, 10, 296
Grima, S., Bellon-Maurel, V., Feuilloley, P. and Silvestre, F., 2000, Aerobic
biodegradation of polymers in solid-state conditions: A review of environmental
and physicochemical parameter settings in laboratory simulations, Journal of
Polymers and the Environment, 8(4), 183-195.
Gujer, W. and Jenkins, D., 1975, Nitrification Model for Contact Stabilization Activated-
Sludge Process, Water Research, 9(5-6), 561-566.
Gunaseelan, V. N., 1997, Anaerobic digestion of biomass for methane production: a
review, Biomass & Bioenergy, 13(1-2), 83-114.
Hoar, T. P. and Rothwell, G. P., 1970, Potential/Ph Diagram for a Copper-Water-
Ammonia System Its Significance in Stress-Corrosion Cracking of Brass in
Ammoniacal Solutions, Electrochimica Acta 15(6), 1037.

223
Holt, D. M. and E. B. G. Jones, 1983, Bacterial Degradation of Lignified wood cell wall
in anaerobic aquatic habitats, Applied and Environmental Microbiology, 46(3),
722-727.
Holtz, R. D. and Kovacs, W. D., 1981, An introduction to geotechnical engineering.
Englewood Cliffs, N.J., Prentice-Hall.
Jain P., Kim H., and Towsend, T. G., 2005, Heavy metal content in soil reclaimed from a
municipal solid waste landfill, Waste Management 25(1), 25-35.
Jambeck, J., 2004, The disposal of CCA-treated wood in simulated landfills: potential
impacts, Doctoral dissertation, University of Florida
Jang, Y. C. and Townsend, T. G., 2003, Leaching of lead from computer printed wire
boards and cathode ray tubes by municipal solid waste landfill leachates,
Environmental Science & Technology, 37(20), 4778-4784.
Jang, Y. C. and Townsend, T. G., 2003, Effect of waste depth on leachate quality from
laboratory construction and demolition debris landfills, Environmental Engineering
Science, 20(3), 183-196.
Jansen, B., Nierop, K. G. J. and Verstraten, J. M., 2003, Mobility of Fe(II), Fe(III) and A1
in acidic forest soils mediated by dissolved organic matter: influence of solution pH
and metal/organic carbon ratios, Geoderma 113(3-4), 323-340.
Jerger. D., and Tsao, T. G, 1987, Feed composition in anaerobic digestion of biomass,
Chynoweth, D. P. and R. Isaacson edited. London, New York, Elsevier Applied
Science.
Kayhanian, M., 1995, Biodegradability of the organic fraction of municipal solid-waste
in a high-solids anaerobic digester, Waste Management & Research, 13(2), 123-
136.
Kjeldsen, P.,Barlaz, M. A., Rooker, A. P., Baun, A., Ledin, A. and Christensen, T. H,
2002, Present and long-term composition of MSW landfill leachate: A review.
Critical Reviews in Environmental Science and Technology 32(4), 297-336.
Komilis, D. P. and Ham, R. K., 2003, The effect of lignin and sugars to the aerobic
decomposition of solid wastes, Waste Management, 23(5), 419-423.
Komilis, D. P., Ham, R. K. and Stegmann, R., 1999, The effect of municipal solid waste
pretreatment on landfill behavior: a literature review, Waste Management Research,
17, 10-19.
Krogmann, U. and Woyczechowski H., 2000, Selected characteristics of leachate,
condensate and runoff released during composting of biogenic waste, Waste
Management & Research 18(3), 235-248.

224
Eary, L. E and Rai, D., 1987, Kinetics of Chromium(IlI) oxidation to chromium(VI) by
reaction with manganese-dioxide, Environmental Science & Technology, 21(12),
1187-1193.
Lee, G., Bigham, J. M. and Faure, G., 2002, Removal of trace metals by coprecipitation
with Fe, A1 and Mn from natural waters contaminated with acid mine drainage in
the Ducktown Mining District, Tennessee, Applied Geochemistry, 17(5), 569-581.
Lee, FL, 1996, Waste composition and characteristics as predictors of landfill
stabilization, doctoral dissertation, University of Florida
Lee, J. Y., Lee, C. H. and Lee, K. K., 2002, Evaluation of air injection and extraction
tests in a landfill site in Korea: implications for landfill management,
Environmental Geology, 42(8), 945-954.
Liao, S. Y., Cheng, Q., Jiang, D. M., and Gao, J., 2005, Experimental study of
flammability limits of natural gas-air mixture, Journal of Hazardous Materials,
119(1-3), 81-84.
Ling, H. L, Leshchinsky, D., Mohri, Y., and Kawabata, T., 1998, Estimation of municipal
solid waste landfill settlement, Journal of Geotechnical and Geoenvironmental
Engineering, 124(1), 21-28.
Marques, A. C. M., Filz, G. M. and Vilar, O. M., 2003, Composite compressibility model
for municipal solid waste, Journal of Geotechnical and Geoenvironmental
Engineering, 129(4), 372-378.
Masscheleyn, P. H., Delaune, R. D. and Patrick, W. H., 1991, Effect of redox potential
and pH on arsenic speciation and solubility in a contaminated soil, Environmental
Science & Technology, 25(8), 1414-1419.
Mata-Alvarez, J., Cecchi, F., Pavan, P. and Llabres, P., 1990, The performance of
digesters treating the organic fraction of municipal solid wastes differently sorted,
Biological Wastes, 33, 181-199
McBean, E. A., Rovers, F. A. and Farquhar, G. J., 1995, Solid waste landfill engineering
and design, Englewood Cliffs, N.J., Prentice Hall PTR.
McBride, M. B. and Blasiak, J. J., 1979, Zinc and copper solubility as a function of pH in
an acid soil, Soil Science Society of America Journal, 43(5), 866-870.
Meima, J. A. and Comans, R. N. J., 1997, Geochemical modeling of weathering reactions
in municipal solid waste incinerator bottom ash, Environmental Science &
Technology, 31(5), 1269-1276.
Miyazawa, M., Pavan, M. A. and Neto, L. M., 1993, A possible mechanism for
manganese release from acid soil, Pesquisa Agropecuaria Brasileira, 28(6), 725-
731.

225
Myles, T. G. http://www.utoronto.ca/forest/termite/lig-mat.htm
Innocente, N., Moret, S., Corradini, C. and Conte, L. S., 2000, A rapid method for the
quantitative determination of short-chain free volatile fatty acids from cheese
Journal of Agricultural and Food Chemistry, 48(8), 3321-3323.
O'Keefe, D. M. and Chynoweth, D. P. 2000, Influence of phase separation, leachate
recycle and aeration on treatment of municipal solid waste in simulated landfill
cells, Bioresource Technology, 72, 55-66.
Okieimen, F. E., Sogbaike, C. E. and Ebhoaye, J. E., 2005, Removal of cadmium and
copper ions from aqueous solution with cellulose graft copolymers, Separation and
Purification Technology 44(1), 85-89.
Oweis, I. S. and Khera, R. P., 1998, Geotechnology of waste management, Boston, PWS
Publishing.
Owen W, Stuckey D, Healy J, Young L, McCarty P, 1979, Bioassay for monitoring
biochemical methane potential and anaerobic toxicity, Water Research, 13, 485-492.
Owens, J. M. and Chynoweth, D. P., 1993, Biochemical methane potential of municipal
solid-waste (MSW) components, Water Science and Technology, 27(2), 1-14.
Pagarwal. U, 2005, Raman imaging of lignin and cellulose distribution in black spruce
wood Picea mariana cell walls Proceedings of the 59th APPITA Annual
Conference and Exhibition incorporating the 13th ISWFPC, Carlton, Victoria,
Australia
Parawira, W., M. Murto, J. S. Read and B. Mattiasson, 2004, Volatile fatty acid
production during anaerobic mesophilic digestion of solid potato waste. Journal of
Chemical Technology and Biotechnology, 79(7), 673-677.
Park H. I. And Lee, S. R., 1997, Long-term settlement behavior of landfills with refuse
decomposition, Journal of Solid Waste Technology and Management, 24(4), 159-
165
Pauss, A., Nyns, E. J. and Naveau, FL, 1984, Production of methane by anaerobic
digesting of domestic refuse, EEC Conference on Anaerobic and Carbohydrate
Hydrolysis of Waste, Luxembourg, 8-10
Pohland, F. G., 1980, Leachate Recycle as Landfill Management Option, Journal of the
Environmental Engineering Division-ASCE, 106(6), 1057-1069.
Pohland, F. G. and Kim, J. C., 1999, In situ anaerobic treatment of leachate in landfill
bioreactors, Water Science and Technology, 40(8), 200-210.
Pokhrel, D. and Viraraghavan, T., 2004, Leachate generation and treatment A review,
Fresenius Environmental Bulletin, 13(3B), 223-232.

226
Qian, X., Koemer, R. M. and Gray, D. H., 2002, Geotechnical aspects of landfill design
and construction, Upper Saddle River, N.J., Prentice Hall.
Rai, D., Sass, B. M. and Moore, D. A., 1987, Chromium(III) hydrolysis constants and
solubility of chromium(III) hydroxide, Inorganic Chemistry, 26(3), 345-349.
Rai, D., Eary, L. E., and Zachara, J. M., 1989, Environmental chemistry of chromium,
Science of the Total Environment, 86(1-2), 15-23.
Rajwanshi, P., Singh, V., Gupta, M. K. and Dass, S., 1997, Leaching of aluminium from
cookwares a review, Environmental Geochemistry and Health, 19(1), 1-18
Ravat, C., Monteil-Rivera, F. and Dumonceau, J., 2000, Metal ions binding to natural
organic matter extracted from wheat bran: application of the surface complexation
model, Journal of Colloid and Interface Science, 225(2), 329-339.
Read, A. D., Hudgins, M., Harper, S., Phillips, P. and Morris, J., 2001, The successful
demonstration of aerobic landfilling: The potential for a more sustainable solid
waste management approach?, Resource, Conservation and Recycling, 32, 115-
146.
Reid, I. D. and Seifert, K. A., 1982, Effect of an atmosphere of oxygen on growth,
respiration, and lignin degradation by white-rot fungi, Canadian Journal of Botany-
Revue Canadienne De Botanique, 60(3), 252-260.
Reinhart, D. R. and AlYousfi, A. B., 1996, The impact of leachate recirculation on
municipal solid waste landfill operating characteristics, Waste Management &
Research, 14(4), 337-346.
Reinhart, D. R. and Chopra, M. B., 2000, MSW landfill leachate collection systems for
the new millennium, report 00-13, Florida center for solid and hazardous waste
management, Gainesville, Florida, USA
Reinhart, D. R., McCreanor, P. T. and Townsend, T., 2002, The bioreactor landfill: its
status and future, Waste Management Research 20, 172-186.
Richard, F. C. and Bourg, A. C. M., 1991, Aqueous geochemistry of chromium a review,
Water Research, 25(7), 807-816.
Rittmann, B. E. and McCarty, P. L., 2001, Environmental biotechnology : principles and
applications, Boston, McGraw-Hill.
Rowell, R., 2005, Handbook of wood chemistry and wood composites, Boca Raton, FL
CRC press,
Sadiq, M., 1997, Arsenic chemistry in soils: an overview of thermodynamic predictions
and field observations, Water Air and Soil Pollution, 93(1-4), 117-136.

227
Science Applications International Corporation (SAIC), 2000, Characterization and
evaluation of landfill leachate.,Draft, ,EPA Contract 68-W6-0068,, Arlington, VA,
USA.
Sheridan, S. (2003), Modeling solid waste settlement as a function of mass loss, Masters
thesis, University of Florida, Gainesville, FL, USA.
Shukla, S. R. and Pai, R. S., 2005, Adsorption of Cu(II), Ni(II) and Zn(II) on dye loaded
groundnut shells and sawdust, Separation and Purification Technology, 43(1), 1-8.
Skyllberg, U., Raulund-Rasmussen, K. and Borggaard, O. K., 2001, pH buffering in
acidic soils developed under Picea abies and Quercus robur effects of soil organic
matter, adsorbed cations and soil solution ionic strength, Biogeochemistry, 56(1),
51-74.
Snoeyink, V. L. and Jenkins, D., 1980, Water chemistry. New York, Wiley.
Solid Waste Association of North America (SWANA), 2003, The environmental
consequences of disposing of products containing heavy metals in municipal solid
waste landfills, SWANA, Silver Springs, MD, USA.
Sowers, G. F., 1973, Settlement of waste disposal fills. Soil mechanics and foundation
engineering conference, Minneapolis, MN., USA.
Stegmann, R., 1983, New Aspects on Enhancing Biological Processes in Sanitary
Landfill, Waste Management & Research, 1, 201-211.
Stessel, R. I. and Murphy R. J., 1992, A lysimeter study of the aerobic landfill concept
Waste Management Research, 10,485-503.
Stinson, J. A. and Ham, R. K., 1995, Effect of lignin on the anaerobic decomposition of
cellulose as determined through the use of a biochemical methane potential method,
Environmental Science & Technology, 29(9), 2305-2310.
Sublet, R., Simonnot, M. O., Boireau, A. and Sardin, M., 2003, Selection of an adsorbent
for lead removal from drinking water by a point-of-use treatment device, Water
Research 37(20), 4904-4912.
Summerfelt, S. T., Davidson, J. and Waldrop, T., 2003, Evaluation of full-scale carbon
dioxide stripping columns in a coldwater recirculating system, Aquacultural
Engineering, 28(3-4), 155-169.
Tchobanoglous, G., Theisen, H. and Vigil, S. A., 1993, Integrated solid waste
management: engineering principles and management issues, New York, McGraw-
Hill.

228
Tipping, E., 2005, Modelling A1 competition for heavy metal binding by dissolved
organic matter in soil and surface waters of acid and neutral pH, Geoderma, 127(3-
4), 293-304.
Townsend, T. G., Miller, W. L., Lee, H. J. and Earle, J. F. K., 1996, Acceleration of
landfill stabilization using leachate recycle, Journal of Environmental Engineering-
ASCE, 122(4), 263-268.
Townsend, T., Tolaymat, T., Solo-Gabriele, H., Dubey, B., Stook, K. and Wadanambi, L.,
2004, Leaching of CCA-treated wood: implications for waste disposal, Journal of
Hazardous Materials, 114(1-3), 75-91.
Ugwuanyi, J. O., Harvey, L. M. and McNeil, B., 1999, Effect of process temperature, pH
and suspended solids content upon pasteurization of a model agricultural waste
during thermophilic aerobic digestion, Journal of Applied Microbiology, 87(3),
387-395.
United States Department of Agriculture (USDA), 1999, Wood handbook, FPL-GTR-
113, Madison, WI, USA.
United States Environmental Protection Agency (USEPA), 1996, Test Methods for
Evaluating Solid Waste, SW-846, 3r ed.; Office of Solid Waste: Washington, D. C.,
USA.
United States Environmental Protection Agency (USEPA), 1997, Emission factor
documentation for AP-42 section 2.4 municipal solid waste landfills, Office of
Air Quality Planning and Standards and Office of Air and Radiation, North
Carolina, USA
United States Environmental Protection Agency (USEPA), 1999, Water quality criteria;
notice of availability; 1999 update of ambient water quality criteria for ammonia,
Federal Register vol. 64 no.245, Washington D.C., USA.
United States Environmental Protection Agency (USEPA), 2002, Municipal solid waste
in the United States: 2000 facts and s, EPA530-R-02-001, Office of solid waste and
emergency response, Wahington D. C., USA.
United States Environmental Protection Agency (USEPA), 2003, Municipal solid waste
in the United States: 2003 facts and s, EPA 530-R-03-011, Office of solid waste
and emergency response, Wahington D. C., USA.
United States Environmental Protection Agency (USEPA), 2003, Municipal Solid Waste
Generation, Recycling, and Disposal in the United States Facts and s for 2003,
U.S.EPA, Washington D.C., USA.
United States Environmental Protection Agency (USEPA), 2003, National Primary
Drinking Water Standards, EPA epa816-f-03-016, Office of Water, Washington D. C.,
USA.

229
United States Environmental Protection Agency (USEPA), 2004, Monitoring approaches
for landfill bioreactors, EPA-600-R-04-31, Cincinnati, OH, USA.
United States Environmental Protection Agency (USEPA), 2005, Landfill gas emission
model ,LandGEM, version 3.02 users guide, EPA-600/R-05/047, Washington D.C.,
USA.
United States Environmental Protection Agency (USEPA), 2005, Municipal solid waste
in the United States: 2003 facts and s, EPA530-F-05-003, Office of solid waste and
emergency response, Wahington D. C., USA.
Valorga, 1985, Waste recovery as a source of methane and fertilizer The Valorga
process, 2nd Annual Internal Symposium on Industrial Resource Management,
Philadelphia, USA.
Vikman, M., Karjomaa, S., Kapanen, A., Wallenius, K. and Itavaara, M., 2002, The
influence of lignin content and temperature on the biodegradation of lignocellulose
in composting conditions, Applied Microbiology and Biotechnology, 59(4-5), 591-
598.
Vikman, M., Karjomaa, S., Kapanen, A., Wallenius, K. and Itavaara, M., 2002, The
influence of lignin content and temperature on the biodegradation of lignocellulose
in composting conditions, Applied Microbiology and Biotechnology, 59(4-5), 591 -
598.
Wall, D. K. and Zeiss, C., 1995, Municipal Landfill Biodegradation and Settlement,
Journal of Environmental Engineering-ASCE, 121(3), 214-224.
Wang, Q. H., Kuninobu, M., Ogawa, H. L, and Kato, Y., 1999, Degradation of volatile
fatty acids in highly efficient anaerobic digestion, Biomass & Bioenergy, 16(6),
407-416.
Warith M. A. and Takata, G. J., 2004, Effect of aeration on fresh and aged municipal
solid waste in a simulated landfill bioreactor, Water Quality Research Journal of
Canada, 39(3), 223-229.
Watsoncraik, I. A., James, A. G. and Senior, E., 1994, Use of multistage continuous-
culture systems to investigate the effects of temperature on the methanogenic
fermentation of cellulose-degradation intermediates, Water Science and
Technology, 30(12), 153-159.
White, R. H., 1987, Effect of lignin content and extractives on the higher heating value of
wood. Wood and Fiber Science, 19(4), 446-452.
Yazdani, R., Kieffer, J., Akau, H., and Augenstein, D., 2003, Monitoring the performance
of anaerobic landfill cells with fluids recirculation final report, Yolo County,
California, USA.

230
Zhang, F. S. and Itoh, H., 2003, Adsorbents made from waste ashes and post-consumer
PET and their potential utilization in wastewater treatment, Journal of Hazardous
Materials 101(3), 323-337.
Zysset, M., Blaser, P., Luster, J. and Gehring, A. U., 1999, Aluminum solubility control
in different horizons of a Podzol, Soil Science Society of America Journal, 63(5),
1106-1115.

BIOGRAPHICAL SKETCH
Hwidong Kim was bom on Dec 17, 1970 in Wonju, South Korea. He graduated
with a Bachelor of Engineering degree from the Department of Biochemical Engineering
in 1992 from Inha University, Inchon, Korea. He enrolled the graduate school of Inha
University in the spring of 1993. After 1 academic year of graduate study, he enlisted in
military service for two years from 1994 through 1996. He received Masters degree in
Biochemical Engineering from Inha University in 1997.
He enrolled in the graduate program in the Department of Agricultural and
Biological Engineering at the University of Florida in August, 1999 and transferred to the
Department of Environmental Engineering Sciences to study solid and hazardous waste.
He earned a graduate research assistantship and teaching assistantship to complete his
study.
231

I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosopl
Tirfrth^G. Townsend, Chairman
Associate Professor of Environmental
Engineering Sciences
I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosophy.
Assistant Professor of Environmental
Engineering Sciences
I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosophy.
(Zar
Frank C. Townsend
Professor of Civil and Coastal Engineering
I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosophy.
Rogef A. Nordstedt
Professor of Agricultural and Biological
Engineering
This dissertation was submitted to the Graduate Faculty of the School of Forest
Resources and Conservation in the College of Agricultural and Life Sciences and to the
Graduate School and was accepted as partial fulfillment of the requirements for the
degree of Doctor of Philosophy. ^
Pramod P. Khargonekar
Dean, College of Engineering
December, 2005
Kenneth Gerhardt
Dean, Graduate School



I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosopl
Tirfrth^G. Townsend, Chairman
Associate Professor of Environmental
Engineering Sciences
I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosophy.
Assistant Professor of Environmental
Engineering Sciences
I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosophy.
(Zar
Frank C. Townsend
Professor of Civil and Coastal Engineering
I certify that I have read this study and that in my opinion it conforms to acceptable
standards of scholarly presentation and is fully adequate, in scope and quality, as a
dissertation for the degree of Doctor of Philosophy.
Rogef A. Nordstedt
Professor of Agricultural and Biological
Engineering
This dissertation was submitted to the Graduate Faculty of the School of Forest
Resources and Conservation in the College of Agricultural and Life Sciences and to the
Graduate School and was accepted as partial fulfillment of the requirements for the
degree of Doctor of Philosophy. ^
Pramod P. Khargonekar
Dean, College of Engineering
December, 2005
Kenneth Gerhardt
Dean, Graduate School


143
mass losses were also correlated with volume loss represented by settlement of the
lysimeters. This relationship was found to be:
[settlement, %] = (AH, %) = 16.90 x log [mass loss, %] 6.24
The performance of the aerobic and anaerobic lysimeters was also evaluated by
analyzing lignocellulosic waste samples for biochemical methane potential (BMP). For
wood waste, which is typically categorized as a non-biodegradable material in landfills,
no significant influence of air injection on SYP block decomposition was found using
cellulose/lignin analysis. However, the methane yields of the SYP blocks excavated from
the aerobic lysimeter were significantly lower than those of the anaerobic lysimeter. BMP
analysis of the other lignocellulosic materials resulted in great differences in
biodegradation between the aerobic and anaerobic lysimeters.
6.2 The Implication of This Research
It may not prove surprising that waste can rapidly decompose in an aerobic
condition. It could be also anticipated that methane concentrations decreased as a result
of air addition. These results can be found from other research, and these may not be the
novel contributions of this research. However, in addition to these common consequences,
other results of this research, which could possibly occur in large scale landfill, may be
helpful for landfill operators and engineers when choosing between aerobic and
anaerobic systems.
As discussed in Chapter 4, all biodegradable wastes were not decomposed within
the test period of the aerobic lysimeter. However, Figure 2-5 in Chapter 2 showed that
BOD concentrations of the aerobic lysimeters lowered below 50 mg/L before day 300. In
comparison with the percentage of mass loss (20%) at low BOD concentrations of the


This document is dedicated to my parents and loving wife


140
Figure 5-8. Settlement prediction of the aerobic lysimeters


155
Table A-l. Actual mass loss and predicted values of the aerobic and anaerobic lysimeter
Actual mass loss (g)
Mass loss predicted (g)
Lys 2 (aerobic)
3,989 g
4,069 g
Lys 4 (anaerobic)
3,397 g
3,787 g
Table A-2. Mass and density of wastes excavated by de
pth
layers
depth
(from)
depth (to),
(inches)
wt(g)
wt (lb)
vol (cf)
Density
(pcf)
Density
(pcy)
2-1
21
24
1829.2
4.0
0.05
82.15
2218
24
28
2519.1
5.6
0.07
84.85
2291
28
33
2614.9
5.8
0.08
70.46
1903
2-2
33
45
5206.2
11.5
0.20
58.46
1578
2-3
45
57
4544.9
10.0
0.20
51.03
1378
2-4
57
66
4038.5
8.9
0.15
60.46
1632
4-1
22
24
1151.2
2.5
0.03
77.55
2094
24
26
1137.2
2.5
0.03
76.61
2069
26
28.5
1342.9
3.0
0.04
72.38
1954
4-2
28.5
32
1763.8
3.9
0.06
67.90
1833
32
34.5
1382.7
3.0
0.04
74.52
2012
34.5
41
3673.1
8.1
0.11
76.14
2056
4-3
41
53
5638.7
12.4
0.20
63.31
1709
4-4
53
66
6556.1
14.5
0.21
67.95
1835


175
100 200 300 400 500 600 700
Days
Figure C-7. The change in calcium (Ca) of the aerobic and anaerobic lysimeters over time


LIST OF REFERENCES
Agdag, O. N. and Sponza D. T., 2004, Effect of aeration on the performance of a
simulated landfilling reactor stabilizing municipal solid wastes, Journal of
Environmental Science and Health Part a-Toxic/Hazardous Substances &
Environmental Engineering 39(11-12), 2955-2972.
Akhtar, M., Blanchette, R. A. and Kirk, T. K., 1997, Fungal Delignification and
Biochemical pulping of wood. Advances in Biochemical Engineering &
Biotechnology 57, 159-195.
Alymore, M. G., 2001, Treatment of a refractory gold-copper sulfide concentrate by
copper ammoniacal thiosulfate leaching. Minerals Engineering 14(6), 615-637.
American Public Health Association (APHA), American Water Works Association
(AWWA), Water Environment Federation (WEF), 1995, Standard Methods for the
examination of water and wastewater, American Public Health Association,
Washington D.C., USA.
American Society for Testing and Materials (ASTM), 1992, Standard Test Method for
Determining the Anaerobic Biodegradation Potential of Organic Chemicals,
Annual Book of ASTM standard Vol. 11.01, 897-901.
Arzutug, M. E., Kocakerim, M. M. and Copur, M., 2004, Leaching of malachite ore in
NH3-saturated water. Industrial & Engineering Chemistry Research 43(15), 4118-
4123.
Audsley, E., Alber, S., Clift, R., Cowell, S., Crettaz, P., Gaillard, Hausheer, J., Jolliet, O.,
Kleijn, R., Mortensen, B., D., Roger, E., Teulon, H., Weidema, B., van Zeijts, H.,
1997, Harmonisation of environmental life cycle assessment for agriculture, final
report, Concerted Action AIR3-CT94-2028, European Commission DG VI,
Brussels, Belgium.
Babel, S. and Kumiawan, T. A., 2003, Low-cost adsorbents for heavy metals uptake from
contaminated water: a review. Journal of Hazardous Materials 97(1-3), 219-243.
Barlaz, M. A., Ham, R. K. and Schaefer, D. M., 1992, Microbial, chemical and methane
production characteristics of anaerobically decomposed refuse with and without
leachate recycling. Waste Management & Research 10(3), 257-267.
Barlaz, M.A., Ph.D. Dissertation, 1988, University of Wisconsin at Madison.
219


116
O 10 20 30 40 50
Days
Figure 4-4. Changes in cumulative methane volume of lignocellulosic materials over time


139
O 100 200 300 400
Time (days)
Figure 5-6.Correlation of logarithm of mass loss of the aerobic lysimeters over time
Figure 5-7. Different k values of anaerobic lysimeters at lag and log phases


Figure 3-16 (continued)
Metal concentrations (mg/L)
ro u
U <-/
8 8 8 8 8 8
r k>
*<
C/
3'
O
HT
55 <-
r
v;
c>
3
o
o
55
500
Metal concentrations (mg/L)
o
k> bo N>
N) -U 00 N>
so


98
Separation was performed using a vibratory shaker assembly. Materials greater
than 0.475 cm were sorted into office paper, cardboard, newspaper, and identifiable non-
biodegradable materials. Materials less than 0.475 cm in size were weighed and ground
to less than 0.25 mm using an Urschell mill (Fritsch, German). VS content was
determined by comparing weight loss after igniting the ground samples at 550C for 2
hours. Since the size of ground wood samples was found to be irregular, samples were
screened again using No 20 mesh (mesh size = 0.8mm).
4.2.3 Methane Yield Determination
The methane yield of both the new and excavated samples was measured. A
synthetic media containing buffers, nutrients and trace metals was prepared using ASTM
method El 196-92 (ASTM, 1992). Anaerobically digested sludge obtained from a
laboratory-scale anaerobic reactor was added as an inoculum to the prepared media while
flushing with nitrogen gas. Three serum bottles were used for each type of waste sample.
A 100-mL portion of inoculated media was then transferred into the prepared serum
bottles along with solid waste samples under anaerobic conditions. Pure cellulose was
utilized as primary positive controls. Sludge without waste sample was employed as a
negative control. A gas sample was collected from each serum bottle once per week over
a 50-day period. At each sampling, a 50-mL capacity syringe was used to measure the
biogas volumes. The collected biogas was analyzed for methane and carbon dioxide
using a gas chromatograph (Model 5890, Hewlett Packard, USA) equipped with a
thermal conductivity detector and GS-Carbon PLOT capillary column (30 m X 0.32 mm
ID, Agilent Technology, Palo Alto, CA, USA).
In this research, only biodegradable materials were included to estimate the average
BMP value. Biodegradable components of the new waste included office paper,


CHAPTER 5
LANDFILL SETTLEMENT BEHAVIOR WITH WASTE DECOMPOSITION
5.1 Introduction
Since Sowers (1973) proposed the equation to describe landfill settlement using
established soil consolidation models, various methods have been developed to interpret
and predict landfill settlement. Gibson and Lo (1961) applied a damping system to
account for landfill settlement and Ling et al. (1998) attempted to find best-matched
curves for landfill settlement behavior using various mathematical functions. In the past,
landfill settlement behavior has been interpreted similarly to soil consolidation models.
These types of interpretation were successfully applied to primary settlement, which
readily occurs as a result of overburden pressure. However, unlike soil consolidation,
landfill settlement is primarily determined by secondary settlement, which is controlled
for the most part by waste decomposition. Among numerous models developed so far,
only a few include waste decomposition to account for landfill settlement (Park and Lee,
1997). Secondary settlement is even more important for bioreactor landfills which are
designed to enhance waste decomposition by the addition of moisture and/or air. There is
a need to develop and to validate a model that can reliably predict secondary settlement.
It is proposed that methods for predicting the secondary settlement at MSW
landfills can be developed by relating waste decomposition over time with landfill
settlement over time. The most common approach for modeling, landfill gas production is
to model waste decomposition as a First-order function. A similar approach would be to
relate settlement (volume loss) to waste decomposition (mass loss) and to thus predict
119


145
in terms of Cr release to the environment since Cr should be reduced to trivalent form
when fixed in the wood. It was also reported that the oxidation of Cr would not happen in
landfill conditions due to the neutral pH and reducing conditions (Townsend et al., 2004).
The relationship between mass loss and volume loss would be another important
finding obtained through this research. As previously mentioned in Chapter 5, this
relationship could apply to the development of landfill settlement model assuming the
constant cross sectional area. Moreover, if it is possible to estimate the percentage of
biodegradable fraction of waste, ultimate settlement and long-term settlement could be
evaluated. More observations would be required to develop the relationship between
mass loss and volume loss since both aerobic and anaerobic lysimeters did not reach their
ultimate settlement. However, the exploration of the relationship between these two
parameters may be used to develop the landfill settlement model to describe waste
settlement more effectively.
The research results may provide insight into the various possibilities that may
occur in aerobic and anaerobic landfills. These results can be used for life-cycle
assessment to compare aerobic and anaerobic landfills. Aerobic landfills are expected to
reduce the cost for monitoring landfills and leachate treatment due to rapid waste
decomposition and low organic carbon in aerobic landfills. However, additional cost may
be required to treat the leachate with high metal concentrations and to install and operate
the facilities needed for air addition.
6.3 Conclusions
Conclusions obtained through this research can be summarized as follows:
The pH of aerobic lysimeters could increase up to 9.1


Figure 4-1. The dry weight differences between predicted and measured remaining mass
Lysimeter 2 Lysimeter 4
(Aerobic) (Anaerobic)
Dry weight (g)
K>
O
o
o
4^
o
o
o
Os
o
o
o
oo
o
o
o
o
o
o
o
K>
o
o
o
Initial
Calculated
Remaining mass
Predicted
lost from gas
Measured
Remaining mass
Initial
Calculated
Remaining mass
Predicted
lost from gas
Measured
Remaining mass
14000


102
methane yield of mechanical-sorted waste could be low because great amounts of
inorganic waste may be included in sorted organic fractions. The default methane
potentials of Clean Air Act (CAA) and AP-42 are 0.170 L/g and 0.10 L/g of waste,
respectively (USEPA, 1997), but these values included all organic and inorganic fractions
of solid wastes disposed of in landfills. These values also include the moisture of waste.
When applying the field capacity (58%), methane yield of the raw waste could be 0.086
g/L total waste, wet.
Based on the biodegradable volatile solids (BVS) of the organic fraction of the
fabricated waste, the percentage of ultimate decomposition of fabricated waste was
estimated. A total of 48% of the fabricated waste can be ultimately decomposed based on
the methane yield data. BVS of the organic was calculated using the methane yields of
those components; BVS values calculated are summarized in Table 4-3. The detailed
procedure used for the estimation of BVS is described in appendix A.
Forty eight percent of the fabricated waste into mass corresponds to approximately
6500 g in each lysimeter. The total mass of lignocellulosic materials (office paper,
newspaper, cardboard, wood and dog food) loaded in each lysimeter was 9600 g. Thus,
ultimately 67% of total lignocellulosic materials were biodegradable. When the mass of
the biodegradable fraction was compared to the mass loss measured from the excavated
waste, 42% and 37% of the total lignocellulosic waste of the aerobic and anaerobic
lysimeters were decomposed during a test period, respectively. This corresponded to 62%
and 54 % of the lignocellulosic wastes having decomposed.
4.3,2 Solid Waste Excavation
The characteristics of solid waste excavated are summarized in Table 4-4. The total
amounts of water contained in each lysimeter were 15,600 mL and 18,800 mL for


161
Figure B-3. A schematic of the temperature control system


CHAPTER 3
THE FATE OF HEAVY METALS IN SIMULATED LANDFILL BIOREACTORS
UNDER AEROBIC AND ANAEROBIC CONDITIONS
3.1 Introduction
An issue of current debate in the solid waste community is the fate of heavy metals
disposed in MSW landfills (SWANA, 2003). Heavy metals may be present as a result of
industrial residuals, but more importantly for MSW landfills, they result from
manufactured products. Examples include lead from electronic devices and copper,
chromate and arsenic from treated wood. This debate has taken on more immediate
concern as several US states have banned certain wastes containing heavy metals (e.g.,
leaded cathode ray tubes) from disposal in landfills (SWANA, 2003). These bans are in
part a result of fears regarding the fate of the disposal of metals in landfills.
For the most part, heavy metals have been thought to be relatively well contained in
typical anaerobic landfills. According to Kjeldsen et al. (2002), the amount of heavy
metals dissolved and contained in leachate is very low relative to those present in the
waste. Most metals are thought to be released during the initial stage of landfill
decomposition as a result of the lower pH. Once a landfill enters the methanogenic phase,
heavy metal concentrations in leachate dramatically decrease, and in many cases, their
levels decrease to lower than drinking water standards (Kjeldsen et al., 2002).
Bioreactor landfills are becoming a more common method of managing MSW, and the
impact of these systems on metal leachability should be evaluated. Since bioreactor
landfills involve exposing a much larger percentage of waste to moisture, the total mass
47


214
Table C-16. Metal concentrations of lysimeter 3 (mg/L)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
8/13/2003
5.04
2.14
0.42
0.08
20.59
6.45
0.02
77.06
8/15/2003
10.57
2.10
0.40
0.06
21.03
7.83
0.01
87.07
8/19/2003
20.35
1.89
0.42
0.05
31.37
7.86
0.06
87.99
8/22/2003
20.27
2.04
0.46
0.04
40.74
9.05
0.03
97.37
8/26/2003
9.26
2.09
0.38
0.05
40.01
9.18
0.02
103.86
9/9/2003
3.66
2.35
0.27
0.03
32.39
8.40
0.02
124.89
9/17/2003
2.61
2.72
0.24
0.03
31.84
8.80
0.03
155.53
9/24/2003
1.87
2.57
0.18
0.02
27.68
7.87
0.02
167.62
10/9/2003
1.35
2.62
0.15
0.03
26.36
7.80
0.04
181.27
10/15/2003
1.23
2.65
0.14
0.02
26.85
7.95
0.03
183.65
10/21/2003
1.23
2.72
0.14
0.03
27.91
8.03
0.03
182.82
10/28/2003
0.55
2.98
0.15
0.03
30.76
8.39
0.02
399.65
11/5/2003
0.53
2.91
0.13
0.06
30.86
8.23
0.03
398.20
11/12/2003
0.47
2.75
0.12
0.02
30.80
7.90
0.02
389.54
11/19/2003
0.44
2.74
0.11
0.02
33.28
7.83
0.03
386.99
11/26/2003
0.43
2.75
0.10
0.03
34.45
7.97
0.02
390.47
12/3/2003
0.44
2.80
0.10
0.02
38.23
8.13
0.03
389.78
12/10/2003
0.32
1.97
0.09
0.02
31.30
8.00
0.02
387.52
12/17/2003
0.38
2.64
0.09
0.02
41.92
7.78
0.04
378.75
12/23/2003
0.42
2.80
0.09
0.02
49.61
8.17
0.03
388.76
1/2/2004
0.40
2.67
0.08
0.01
53.79
7.99
0.03
384.01
1/9/2004
0.42
2.86
0.08
0.02
60.94
8.55
0.03
396.12
1/14/2004
0.41
2.66
0.08
0.01
62.34
8.02
0.03
380.51
1/22/2004
0.38
2.59
0.07
0.01
72.81
8.06
0.03
381.30
2/12/2004
0.36
2.22
0.07
0.01
85.17
7.76
0.01
403.33
3/16/2004
0.40
2.58
0.07
0.01
114.72
8.70
0.03
391.24
3/26/2004
0.41
2.74
0.07
0.01
103.28
8.91
0.04
397.52
6/1/2004
0.25
1.13
0.03
0.00
120.84
5.40
0.03
290.20
7/28/2004
0.31
0.91
0.03
0.00
192.79
5.37
0.03
289.10
8/9/2004
0.07
0.53
0.02
0.00
138.54
3.52
0.03
194.27
8/12/2004
0.20
0.86
0.03
0.01
253.34
5.70
0.05
269.86
8/17/2004
0.39
0.81
0.03
0.00
278.07
5.61
0.05
287.40
8/24/2004
0.19
0.78
0.04
0.01
301.47
5.66
0.09
262.24
9/3/2004
0.26
0.70
0.02
0.00
293.19
5.13
0.05
259.62
9/10/2004
0.11
0.60
0.02
0.01
269.31
4.50
0.04
217.89
9/15/2004
0.30
0.75
0.02
0.00
346.35
5.55
0.06
269.70
9/23/2004
0.60
1.26
0.04
0.02
534.21
8.08
0.09
263.15
10/7/2004
0.20
0.57
0.02
0.00
306.63
4.20
0.05
203.07
10/14/2004
0.00
0.37
0.01
0.00
217.10
2.91
0.03
140.47
10/19/2004
0.12
0.44
0.01
0.00
244.93
3.18
0.03
157.73


Metal concentrations (mg/L)
82
Figrue 3-11. Distribution of Cu over a C-pH diagram


Volatile fatty acids (mg/L) Volatile fatty acids (mg/L)
36
30000
25000
20000
15000 -
10000 -
5000 -
lys 1 (aerobic)

Acetic acids
o
Propionic acids
--T
Butyric acids
-fW
50 100 150 200 250 300 350 400
(B)
Figure 2-6. Changes in VFAs of aerobic and anaerobic lysimeters versus time (A) acetic
acid only and (B) acetic acid, propionic acid and butyric acid


48
of metals released might be expected to be high relative to dry landfills. On the other
hand, given that bioreactor landfills recirculate leachate to the landfill and that traditional
bioreactors promote anaerobic waste decomposition (and thus enhance metal removal by
the mechanisms described previously), the impact to the environment may be limited.
An alternative bioreactor landfill technique is to add air in addition to moisture.
Taking into consideration that the leaching behavior of heavy metals is mainly controlled
by redox, pH and the presence of ligands (Benjamin, 2002), it may be that the fate of
heavy metals in aerobic systems will differ from anaerobic systems. The long term fate of
heavy metals in aerated landfills is a question yet to be satisfactorily addressed.
In this research, the fate of heavy metals in simulated aerobic and anaerobic
bioreactor landfills was studied. Four stainless-steel lysimeters were used, two each for
aerobic and anaerobic conditions. Heavy metal-containing wastes were included in the
waste stream added to each lysimeter. Leachate collected from each lysimeter was
analyzed for heavy metals over time. In order to evaluate the heavy metals adsorbed in
solid wastes, waste samples were removed from two of the lysimeters at the end of the
experimental period and analyzed for heavy metal content.
3.2 Materials and Methods
A detailed description of the lysimeters was presented in chapter 2. The methods as
described here focus on the metal-containing components in the fabricated waste and on
metal analysis in the leachate and waste samples.
3.2.1 Heavy Metal Sources in Synthetic Waste
Several MSW components were chosen as sources of heavy metals. Each
component, its corresponding heavy metals and the percentage of each component are
presented in Table 3-1. Aluminum and galvanized steel sheets were purchased from a


153
of glucose would be 2.4 moles rather than 3 moles. Thus,
into CH4 can be written as follows:
[Mass converted into CH4] =
1PH ntSTPfl IT: lmolegaS :.162gC6H10Os
[ 4 generated 22.4 L, STp 2.4m0leCH4
the mass of waste converted
(12)
Therefore, BVS percentage can be determined by substituting the mass converted
into CH4 obtained from equation (12) and initial dry mass for parameters in equation (9).
A.3 Lysimeter Dismantlement
After lysimeter studies were completed, solid wastes were excavated from aerobic
and anaerobic lysimeters. Before excavation, temperature controllers and an air injection
pumps were shut down, and all tubes and cables connected to lysimeters were removed.
Excavation of solid wastes was conducted using 4-ft long narrow-tined fork used for
gardening. A lysimeter was laid down slightly and leaned on a support. In order to collect
all excavated solid waste, a large tub was prepared and placed under the mouth of the
lysimeter. Excavated samples were collected at various depths and measured for their wet
weight. All wood blocks were then collected from the excavated samples and divided into
two fractions for the purpose of compressibility testing and garbage separation. A
fraction of excavated samples for the compressibility test was put inside black garbage
bags and stored at 4C. The other fraction of garbage samples was measured for initial
wet weight and placed in a drying oven to measure moisture content. The mass of waste
and depth of each layer are summarized in Table A-2.
Figures B-8 through B-12 show the features of solid wastes excavated. Solid waste
excavated from the aerobic lysimeter was darker than the solid waste from the anaerobic


Sulfide (mg/L)
41
0 100 200 300 400
Days
Figure 2-11. Changes in sulfide and pH versus time
600
800


16
al. (1999) explained that various enzymatic reactions in microorganisms dictate a greater
decreasing rate of butyric acids than that of other VFAs. However, more biosynthetic
processes are involved in butyric acid production than acetic acid due to longer carbon
chains. In aerobic lysimeters, all three major VFAs were depleted together like other bulk
organic carbon.
The ratio of BOD5 to COD is often used to assess the biodegradability of the
organic matter in leachate, and thus to assess the degree of landfill stabilization. In old
stabilized landfills, the BOD5/COD ratio is below 0.10 (Kjeldsen et al, 2002). A low
BOD5/COD suggests that a leachate is low in biodegradable organic carbon and relatively
high in hard-to-biodegrade organic compounds such as humic compounds. In this
research, low BOD5/COD ratios were observed with the aerobic lysimeters after day 200
(0.04 on average) (Figure 2-7). Relatively high BOD5/COD ratios were exhibited from
the anaerobic lysimeters (0.36 on average). These values fall into the range of average
BOD5/COD ratios proposed by Kjeldsen et al. (2002) for the acid phase (0.58) and the
methanogenic phase (0.06).
2.3.3 Nitrogen
Figure 2-8 shows the changes in ammonia-nitrogen in the lysimeter during the
course of experiment. Ammonia concentrations from the aerobic lysimeters remained
relatively constant, showing a general increase during the course of the experiment. In a
different fashion, ammonia concentrations in the anaerobic leachate increased
dramatically at a point corresponding to an increase in system pH. Ammonia
concentrations in the anaerobic lysimeters increased to concentration in the range of
1000-1600 mg/L. These values then dropped to 800-1000 mg/L and stabilized. Small
increases of ammonia concentrations in leachate of aerobic column were observed after


BOD/COD
37
Figure 2-7. Changes in the ratio BOD/COD of the aerobic and anaerobic lysimeters over
time


Mass loss (g) Mass loss (g)
157
Days
(A)
Days
(B)
o
100
400


61
Figure 3-14 depicts metal concentrations adsorbed on selected lignocellulosic
materials (newspaper and cardboard) and plastic waste. Greater concentrations of most
metals except for Pb and Zn adsorbed on lignocellulosic materials were observed in the
aerobic lysimeter. Bradle (2005) explained that many metals tend to adsorb the organic
matter as the pH increases under oxidizing condition. Zhang and Itoh (2003) reported that
carbonized mixture of polyethylene terephthalate (PTE) and waste ash could be used as a
metal sorbent. However in this research, metal concentrations adsorbed on plastic waste
were substantially low in comparison with metal concentrations adsorbed on the
lignocellulosic materials.
3.4 Discussion
Among the various factors affecting heavy metal leaching under landfill conditions,
the redox and pH may play the most critical role. Under the given redox condition and
pH, the metal oxidation state, ligands, adsorption behavior can be determined. In many
cases, metal precipitation can be controlled by Fe (II) and sulfide concentrations in both
aerobic and anaerobic condition. Cr, Cu, Pb, Zn and As are reported to adsorb on hydrous
ferric oxide at pH > 6, and the precipitation of Cu, Fe, Pb, Mn and Zn is controlled by
sulfide in anaerobic condition. In addition to those ligands, hydroxide ion (OH ) also can
play an important role to precipitate A1 and Cr (Drever, 1988). The various chemical
interactions are depicted in Figure 3-15.
3.4.1 Overall Comparison of Metal Behavior
Figure 3-16 describes the overall trends of metal leaching in aerobic and anaerobic
lysimeters. Among 8 metals under consideration, (Al, As, Cr, Cu, Pb, Mn, Fe and Zn)
greater concentrations of Al, Cr, Cu and Pb were observed in the leachate of the aerobic
lysimeters, and As, Mn, Fe and Zn were observed in the leachate of the anaerobic


Metal concentrations (mg/L) Metal concentrations (mg/L)
90
3.5
3.0
2.5
2.0
1.5
1.0
0.5
0.0
As
0.5
0.4
0.3
0.2
0.1
0.0
1
4
Lysimeters
Lysimeters
Figure 3-16. Comparison of concentrations of metal leached between aerobic and
anaerobic lysimeters.


18
lysimeters after the pH was stabilized at alkaline condition. For the anaerobic lysimeters,
relatively high alkalinity was maintained over the entire test period. Generally, alkalinity
could be generated by CO2 accumulation and ammonification (Fannin, 1987). It is also
consumed by nitrification (Gujer and Jenkins, 1974).
2.3.5 Oxidation Reduction Conditions
Figure 2-11 depicts the change of sulfide concentrations in the aerobic and
anaerobic lysimeters over a period of time. Little changes of sulfide were observed during
the acid phase of both aerobic and anaerobic lysimeters. However, rapid increases in
sulfides along with an increase of pH were exhibited from the aerobic lysimeters and
lysimeter 4. The sulfide level of lysimeter 1 was lowered on day 210 again, but increased
up to 2,600 pg/L during the alkaline phase. The trends of change in sulfide concentration
of lysimeter 2 exhibited are similar to that of lysimeter 1. The highest sulfide level found
in lysimeter 2 was 1,200 pg/L. In lysimeter 4, sulfide concentrations increased as the pH
increased.
It is notable that great concentrations of sulfide were found in the system where air
injection had been taking place. It is hard to understand how sulfide could be presented in
an aerobic environment. In comparison with sulfate concentrations, sulfide was formed
by sulfate reduction (Figure 2-12). However, Figure 2-10 shows that high dissolved
oxygen concentration was also found in the same condition. Snoeyink and Jenkins (1980)
reported that sulfide could be detected under aerobic conditions. They explained that this
phenomenon was caused by a non-equilibrium situation for the reaction between oxygen
and sulfide. Therefore, it can be concluded that sulfide can be found before it oxidizes for
a second time by dissolved oxygen. This result indicates that anaerobic zones were


CHAPTER 2
COMPARATIVE STUDIES OF LEACHATE AND GAS QUALITY OF AEROBIC
AND ANAEROBIC SIMULATED LANDFILL BIOREACTORS
2.1 Introduction
The operation of municipal solid waste (MSW) landfills as bioreactors for the
purpose of rapid landfill stabilization has historically been proposed as an anaerobic
process. Conditions within the landfill are controlled to accelerate the activity of the
anaerobic microorganisms responsible for waste decomposition. The addition of air has
also been proposed as a method to enhance landfill stabilization (Stessel and Murphy,
1992), and recently this technique has gained more attention (Read et al., 2001; Reinhart
et al., 2002). In addition to an enhancement of waste decomposition that is more rapid
than anaerobic operation, a major benefit often cited for air addition is the reduction in
methane emissions relative to anaerobic landfills (Borglin et al., 2004). These studies also
find that the overall strength of leachate (with respect to readily degradable carbon
compounds and oxygen demand) is lower in aerobic systems, offering a potential
advantage with respect to leachate treatment.
Research examining the relative differences in leachate quality between aerobic and
anaerobic systems is very limited. Though the performance of aerobic bioreactor landfills
has been simulated in several studies (Agdag and Sponza, 2004; Warith and Takata,
2004), these studies are often limited with respect to their ability to control several key
parameters, and their lack of a complementary anaerobic system for comparison purposes.
This chapter reports the results of research performed to examine the characteristics of
6


151
these organic carbons, with the rest of the organic carbons remaining in leachate, which
can be measured using bulk organic parameters, such as total organic carbon (TOC).
These relationships can be expressed as shown in equation 8.
[Total mass loss of wastes (g)] = [Mass loss (g) by gas generation]
+ [Changes in TOC (g/L)*leachate remained in lysimeters (L)] (8)
Figure A-2 depicts changes in TOC, mass loss by gas generation and total mass
loss in both aerobic and anaerobic lysimeters. Since aerobic bacteria generate gas and
degrade organics rapidly, the impacts of TOC on total mass loss in aerobic lysimeters
may not be critical. Although small gaps between total mass loss and mass loss by gas
generation were observed at high TOC levels, it can be concluded that gas generation is
the main path of total mass loss. In the anaerobic lysimeters, however, total mass loss was
highly affected by both TOC and gas generation, especially during the first acid-forming
phase. The effects of TOC were then reduced as TOC concentrations decreased.
Ultimately, total mass loss of both aerobic and anaerobic lysimeters may be expressed as
mass loss by gas generation and dissolved organic carbons.
After lysimeter studies were completed, total dry mass of the actual waste was
compared to mass loss predicted using equation (8). Actual mass loss of excavated
garbage and mass loss predicted are summarized in Table A-l. For the aerobic lysimeters,
both values were within 2%. However, some discrepancies between actual mass loss and
the predicted values in the anaerobic lysimeter were observed. The most likely
explanation of this difference would be because of 1) the high methane yield of dog food
and 2) some fraction of waste was used for energy generation rather than CH4 conversion.


14
The pH of the aerobic lysimeters measured in the latter half of the experiment (9.0)
was more alkaline than the pH measured from the anaerobic lysimeters (7.2). According
to other lysimeter studies, higher pH was observed from the aerobic lysimeters in
comparison with that of the anaerobic lysimeter. The range of pH of aerobic lysimeters
has been reported as 7 9 (Stessel and Murphy, 1992; OKeefe and Chynoweth, 2000;
Agdag and Sponza, 2004). Summerfelt et al. (2003) also observed an increase of pH
when air was injected into their aquaculture system. They reported that this increase was
because of CO2 stripping by air; a decrease in CO2 leads to a decrease of carbonic acid
(H2CO3) and bicarbonate concentrations (HCO3) consuming H+ ions. These relationships
can be described by carbonate systems as follows:
C02gas<-> H2CO3 (1)
H2CO3 HC03' + H+ (2)
HC03 C03 + H+ (3)
They additionally concluded that, because the dehydration of carbon acid is rate-limiting,
pH may not increase instantaneously.
2.3.2 Organic Carbon Concentration
Figure 2-4 depicts the change of COD concentrations for the lysimeters versus
time. The initial average COD concentrations in the leachate of the aerobic and anaerobic
lysimeters were 36,000 mg/L and 66,000 mg/L, respectively. The COD values for the
aerobic lysimeters increased up to greater than 84,000 mg/L and decreased rapidly after
pH was stabilized. Although one of the aerobic lysimeters showed high COD (70,000
mg/L) at day 50, the overall COD concentrations of the aerobic lysimeters were lower
than values in the anaerobic lysimeters. Similar trends occurred for COD as were
observed for BOD5 (Figure 2-5). The BOD values of the aerobic lysimeters decreased


30000
25000
20000
15000
10000
5000
0
30000
25000
20000
15000
10000
5000
0
35
200
400
Days
(A)
600
800


26
Table 2-2. Parameters and methods for analysis.
Parameters
Method
Alkalinity
Standard Method 2320B
Ammonia
Standard Method 4500-D
BOD
Standard Method 521 OB
COD
HACH 2720
Conductivity
Standard Method 2510
PH
Standard Method 4500-Fl+
TOC
EPA SW846, Method 9060
Sulfide
HACH 8131
Sulfate, Floride and Chloride
EPA SW846, Method 9056
Sodium
EPA SW846, Method 9060A
VFA
VFA analysis method using GC
(Innocente et al 2000)


55
Copper solubility is controlled by several Cu-containing minerals forming
complexation with Fe and sulfide. In addition, Cu sulfides may coexist with the sulfides
of other metals such as Zn, Pb and As (Faure, 1991). Representative mineral deposits
formed with OH', Fe and/or sulfide include chalcocite (CU2S), chalcopyrite (CuFeS2),
cuprite (Cu20) and malachite (Cu2(0H)2C03). These minerals are widely distributed over
a pe-pH diagram. Ionic Cu is present only at a pH less than neutral and under highly
oxidizing conditions (pe > 2.5). For these reasons, high concentrations of Cu may not be
found under landfill conditions where low ORP and neutral pH are predominant. These
concepts can be applied to the Cu distribution patterns displayed at a low pH in the pH-
Cu concentration chart shown in Figure 3-11.
However, there is a disparity in the Cu concentrations observed and those
thermodynamically predicted for the alkali conditions of the aerobic lysimeters; most of
the Cu precipitated by complexation with various Cu-containing mineral deposits at
alkali and oxidizing conditions. Edwards et al (2000) reported high Cu concentrations in
drinking water at alkali conditions, calling it the blue color phenomenon since water
color changed to blue with high concentrations of Cu. Critchley et al (2004) explained
this blue water was caused by microorganism-intermediated-Cu leaching from a part of
the water delivery system. In this research, blue water was also observed from
condensate passing through a copper tube which connected to a gas collection system of
an aerobic lysimeter. Further development of Cu corrosion caused a small hole on the
same copper tube and called for a replacement of the copper tube with plastic materials
(Figure B-7).


80
pH
Figure 3-9. Distribution of As over a C-pH diagram


Cumulative gases,
135
Figure 5-2 The changes in settlement, cumulative gas (CO2 and CH4) and pH over time


Total mass of metals adsorbed on lignocellulosic materials (mg)
85
120
100
80
60
40
20
0
300
250
200
150
100
50
0
Figure 3-13. The comparison of aerobic and anaerobic lysimeters in respect of total mass
of metals adsorbed on lignocellulosic materials


Zn, mg/L Zn, mg/L
79
300
250
200 -
150
100
50 r


AEROBIC
Lys 1
O Lys 2
O
9 o<£
o
CP
r>Q0ooft 8cgoo 8
BDL
50
100
150 200
Days
250
300
350
700
Figure 3-8. Changes of Zn concentrations over time


Cumulative metals (Al) released out of the lysimeters
(mg)
92
Mass loss, %
Figure 3-17. Changes in cumulative mass of meta released over a mass loss, %


205
Table C-9. Volatile fa
tty acids (VFA) of lysimeter 1
Acetic
Acid
Propionic
Acid
Isobutyric
Acid
Butyric
Acid
8/18/2004
25045.1
695.9
2160.3
5020.9
9/22/2004
13103.4
1421.7
4984.8
9269.7
9/30/2004
10662.5
1166.9
4304.6
8284.1
10/21/2004
14481.1
1457.8
5599.2
9500.1
11/3/2004
13667.2
1310.8
5194.9
9088.5
12/12/2004
6856.0
773.9
2681.6
5399.9
12/20/2004
3581.9
419.9
1532.8
2810.4
1/22/2005
3303.1
353.9
2515.4
5576.7
1/30/2005
117.4
21.1
166.1
404.0
2/8/2005
14.8
0.0
10.3
30.2
2/16/2005
7.9
0.0
3.7
4.1
3/3/2005
43.3
41.3
19.5
29.8
3/12/2005
58.5
0.0
21.9
36.5
3/25/2005
31.0
42.9
18.7
27.5
4/3/2005
55.2
45.8
19.4
29.4
4/17/2005
35.0
0.0
1.5
3.8
4/24/2005
24.1
0.0
0.0
0.0
4/30/2005
12.6
2.0
0.0
3.7
5/7/2005
32.1
1.7
1.6
3.7
5/16/2005
9.9
0.0
0.0
3.8
5/23/2005
3.9
3.0
1.9
3.7
5/31/2005
6.3
2.8
1.5
3.7
6/6/2005
5.2
2.1
1.5
4.3
6/14/2005
9.9
2.5
0.0
4.6
6/24/2005
6.4
2.5
1.6
4.0
7/5/2005
0.4
0.0
2.2
4.9
7/11/2005
0.9
0.0
1.8
4.5
7/19/2005
2.0
0.0
2.4
6.1


103
lysimeters 2 and 4, respectively. The water in the column was approximately 3,500 mL
less than that initially added. This difference is thought to be a result of water lost by air
stripping for the aerobic lysimeter. The water loss (mL) was calculated assuming gas
emitted from the aerobic lysimeter was fully saturated. The calculated amounts of water
removed by gas-striping were 3,900 mL and 200 mL for lysimeters 2 and 4.
The dry weight of the solid waste excavated was compared with dry weight
calculated by volume of CH4 and CO2 and TOC concentration in leachate (see appendix
A). Dry weights measured and calculated were 8,715 g and 8,741 g for lysimeter 2 and
8,997 g and 9,258 g for lysimeter 4. The measured weight loss and the weight loss
calculated from gas production and leachate solute were comparable. They were less than
2% different for the aerobic column and less than 7% different for the anaerobic column
(Figure 4-1).
4.3.2 Mass Loss for Individual Components
The mass loss for individual components was calculated on the basis of the results
of waste separation procedure. Among four biodegradable wastes (office paper,
cardboard, newsprint and wood blocks), the largest difference of mass loss between the
raw and decomposed waste was observed from the office paper fraction for both
lysimeters (Figure 4-2). Theses differences were reduced in the following order: office
paper (OP) > cardboard (CB) > newspaper (NP) > wood (WD). The percentages of each
component of the excavated waste sample also changed in comparison to the raw waste
(Figure 4-3). The decomposition trends of these wastes were related to the
biodegradability of each component, as described next.


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Adsorption capacity (mg/g)
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105
This was because the pH of the layer 4-4 remained in the acid phase for more than 400
days. As previously described in chapter 2, a buffer (sodium bicarbonate) was added
from the top of the lysimeter. After the buffer addition, the pH of the waste layers was
measured on occasion using the front ports (see Figure 2-1 in chapter 2). While the pH of
the other layers changed to neutral, the pH of the layer 4-4 remained acidic. The pH of
layer 4-4 then increased to 7.3 at day 550.
Based on the methane yields of each layer, the mass losses of lysimeter 2 and 4
were calculated. The total potential gas volume generated from each layer was estimated
on the basis of methane yield of the raw waste. Calculated dry mass loss for lysimeter 2
using the methane yield of the all layers was 4,285g (wd, measured = 4,044g). The loss for
lysimeter 4 was 3088g (wdimeasured= 3,524g). These values were equivalent to 65.8 % and
46.9% of the biodegradable fraction of lignocellulosic materials (Figure 4-6). Though a
small difference between estimated and measured values was observed, the BMP assay
results of the excavated waste were shown to be a good parameter for the evaluation of
waste decomposition in landfills.
4.3.4 Biodegradability of Wood Waste
The excavated SYP blocks were analyzed for cellulose and lignin to access the
performance of the aerobic and anaerobic simulated landfills with respect to the
decomposition of waste with a high lignin content. Table 4-6 summarized the cellulose
and lignin analysis results of raw and decomposed ground SYP blocks. According to the
statistical analysis results, overall lignin and cellulose concentrations of SYP excavated
from the aerobic and anaerobic lysimeters were not significantly different from those of
raw SYP blocks (p > 0.05). However, a relatively large difference in cellulose content of
the SYP blocks from aerobic lysimeter samples 2-1 (42.3 %) and 2-2 (46.1 %) were


97
component was determined so that any differences in the lignin content between aerobic
and anaerobic landfill conditions could be observed.
4.2 Materials and Methods
4.2.1 Composition of Fabricated Waste
The waste stream fabricated for this research was based on the typical MSW
composition previously reported for the U. S. and Florida (see Figures D-4 and 5). For
simplification purposes, several minor components, such as textiles and tires were
excluded from the fabricated waste stream. A greater portion of commingled paper was
allotted as a substitute for those excluded materials. The ratio of office paper, cardboard
and newsprint in commingled paper (4.6:2.6:1) was estimated from previous published
data (FDEP, 2003 and USEPA, 2005). Southern yellow pine (SYP) was selected for the
wood waste used in this study and comprised 5% of the total waste stream. A detailed
classification of wood species that occur in MSW may not be available, but SYP is
known as one of the most widely used species for manufactured wood products (USDA,
1999).
4.2.2 Excavation and Processing of Decomposed Solid Waste
After 1 and 2 years of operation, one aerobic lysimeter and one anaerobic lysimeter
were dismantled and the decomposed waste was removed. When the carriage system was
removed, some amount of leachate remained pounded above the top waste layer in each
lysimeter. This leachate was removed using a vacuum pump; the volume was recorded.
The waste in each lysimeter was divided into four fractions during the removal process.
The wet weight of each fraction of solid waste was recorded. The wood blocks were
separated from the excavated samples which were then dried. The dried samples were
separated by each lignocellulosic component.


81
pH
Figure 3-10. Potential- pH diagram of Cr (Richard and Bourg, 1991)


Metal concentrations (mg/L)
182
Figure C-14. Mn concentration versus pH in leachate from the lysimeters


21
was calculated using the water carrying capacity of the exit gas assuming that the gas was
100% saturated with water vapor. The overall performance of the aerobic lysimeters with
respect to waste decomposition and leachate quality was substantially greater than those
of the anaerobic lysimeters. However, the concentrations of sodium in the aerobic
lysimeters were still too high to meet drinking water standards. The final concentration of
ammonia in the aerobic lysimeters was also substantially higher than the criteria value of
ambient water (0.897 at 30C and pH 8.0) (USEPA, 1999). Although large quantities of
waste were decomposed, leachate of the anaerobic lysimeters still contained high
concentrations of organics, ammonia and anions (Table 2-4). Leachate generated would
be used for recirculation, but excessive volume of leachate must be treated at an on-site
or off-site wastewater treatment plant.
2.4.2 The Comparison of Leachate Parameters with Other Studies
Tables 2-5 and 2-6 summarize the comparison of leachate constituent
concentrations from this study with those from other studies. For aerobic landfill studies,
the maximum concentrations of COD, BOD and ammonia in this study appeared to be
greater than those of other studies, but they were in a similar range overall. The pH of the
aerobic lysimeters of this study was, however, greater than other studies. As previously
discussed, this would be because of the relocation of carbonic acids, bicarbonate, and
carbonate due to CO2 removal by air stripping (Summerfelt, 2003). The high pH of the
aerobic lysimeters implies that great concentrations of carbonate ions were dissolved due
to high partial pressure of CO2, and these carbonate ions might consume more H+ ions
when CO2 was removed. If alkalinity data of other studies were available, it would be


128
moisture addition and air addition may increase the compression index. In contrast,
highly compacted waste may have low compression index due to relatively lower void
ratio. In this study, however, initial conditions of all lysimeters were the same and the
flow rate of air addition (70mL/min) would not be high enough to change the void ratio.
Thus, the changes in compression indices could be contributed by the influence of waste
decomposition.
The compression index could be important and useful tool to compare the landfill
performance in terms of settlement, but Holtz and Kovacs (1981) pointed out that the
compression index for the secondary long-term settlements were oversimplified to
describe more complicated real behavior. Moreover, the compression index may not be
useful to compare the lab-scale landfill settlement occurred within a short period of time.
For example, the compression index of lysimeter 4 appeared to be 2 times greater than
that of lysimeter 2. However, the differences of time and the change in strain of lysimeter
2 were more advanced than those of lysimeter 4 (Table 5-1). Since the date starting with
the second phase was too early, log (t2/fi), a denominator of compression index equation,
was not comparable with that of the anaerobic lysimeters.
5.4.2 Correlation of Mass Loss and Volume Loss
As Figures 5-3 and 5-4 shown, the percentage of mass loss may not correspond
with the same percentage of volume loss; small volume loss may require relatively larger
mass loss. This phenomenon is difficult to assess because waste is heterogeneous and less
understood. Wall and Zeiss (1995) reported that total settlement of a landfill would be
range from 25% to 50%. If it takes consideration that more than 40% of biodegradable
lignocellulosic materials still remained, the relationship between mass loss and volume
loss (equation (4)) may be changed, ultimately.


125
5.3 Results
5.3.1 Settlement Behavior over Time
Figures 5-1 and 5-2 show the settlement of the aerobic and anaerobic lysimeters
with respect to gas generated and pH. Though the aerobic lysimeters remained in an acid
phase for 200 days, CO2 gas was constantly generated over time. This trend corresponded
with the settlement behavior of the aerobic lysimeters. As previously discussed in chapter
2, the accumulation of organic acids might be localized at the bottom of the aerobic
lysimeters due to improper air addition. Thus, it is concluded that the pH of leachate in
the aerobic lysimeters during the acidic phase might not be representative of the pH of
the entire lysimeter and that waste was decomposed aerobically regardless of the pH of
leachate. For the anaerobic lysimeters, however, little change in settlement was observed
while the anaerobic lysimeters remained in the acid phase. For the first 100 days,
approximately 5 7% of the settlement was observed along with CO2 gas increases, but
the settlement was retarded until pH of the anaerobic lysimeter started to change at day
400. Most of the settlement observed from the aerobic and anaerobic lysimeters took
place in a similar trend as the increase in gas. These results indicate that long-term
landfill settlement is mainly affected by waste decomposition. This could be confirmed
by comparing different phase of landfill settlement behavior.
Figure 5-3 depicts the changes in settlement over time on a logarithm scale. The
settlement trends were separated into two phases, (Ca)mjn and (Ca)max by the changes in
slope. For the lysimeters 1 and 2, time differences (At) for (Ca)min were only 50 and 30
days, while At for (Ca)mm were 210 and 410 days for the lysimeters 3 and 4 (Table 5-3).
These results indicate that the settlement of the aerobic lysimeter mainly occurred by


230
Zhang, F. S. and Itoh, H., 2003, Adsorbents made from waste ashes and post-consumer
PET and their potential utilization in wastewater treatment, Journal of Hazardous
Materials 101(3), 323-337.
Zysset, M., Blaser, P., Luster, J. and Gehring, A. U., 1999, Aluminum solubility control
in different horizons of a Podzol, Soil Science Society of America Journal, 63(5),
1106-1115.


62
lysimeters. For the anaerobic lysimeters, the metal leaching occurred in the acid phase,
while occurrence of the metal leaching was relatively well distributed over the pH (5 <
pH < 9) for the aerobic lysimeters except for Pb.
Cumulative masses of metals were calculated by the multiplication of the amount
of leachate used for analysis by the concentration of the metals in the leachate sample
(Figure 3-17). The total amounts of leachate produced and used for the analysis are
summarized in Table 3-3.
Among the 8 metals under consideration, As, Fe, Pb, Mn and Zn increased
substantially for the first 10 to 15% of mass loss and reached a plateau. These leaching
patterns indicate that metal leaching mainly occurred at the initial phase, an acidic
environment, in both aerobic and anaerobic conditions. In contrast to these metals, only
minor change in cumulative masses of A1 and Cu was observed from the anaerobic
lysimeters whereas a consistent increase in these metals was exhibited from the aerobic
lysimeters. Relatively high concentrations of A1 were exhibited during the initial stage of
the anaerobic lysimeters, but no further changes were observed. It is notable that
cumulative concentrations of A1 and Cr increased more rapidly after 25% mass loss
occurred. An increase in the rate of accumulation of these metals corresponds to an
increase in pH to 9.
Overall metal leaching behavior is strongly associated with pH and redox
conditions. Since the anaerobic lysimeters remained in the acidic condition (pH < 6) for
more than 400 days, great amounts of metals such as As, Mn, Fe, Cr and Zn were
released through the leachate. In contrast to the anaerobic lysimeters, greater cumulative
concentrations of Al, Cu and Pb were observed from the aerobic lysimeters. Leaching of


197
Table C-5. Total organic contents (TOC) o
date
lys 1
lys 2
date
lys 3
8/9/2004
6288
0
8/8/2003
26266
21771
8/12/2004
6227
7268
8/13/2003
24492
22581
8/17/2004
8053
5219
8/15/2003
21679
25649
8/9/2004
6418
8/19/2003
28491
28005
8/24/2004
7238
5847
8/26/2003
16688
18640
9/15/2004
18738
4678
9/2/2003
18480
19881
10/5/2004
18881
3363
9/12/2003
18148
18088
10/9/2004
1447
9/19/2003
19209
24694
10/26/2004
17985
1614
10/3/2003
19431
23817
11/19/2004
5205
1496
10/10/2003
18166
21096
12/12/2004
7178
6634
10/17/2003
17519
22649
1/6/2005
12977
13166
10/26/2003
18357
22802
1/13/2005
5617
10720
10/31/2003
15581
21267
1/21/2005
6864
11/12/2003
14397
1/30/2005
630
1997
11/19/2003
14112
16858
2/5/2005
1671
845
11/26/2003
13627
16335
2/16/2005
1454
1392
12/3/2003
14428
16779
2/24/2005
1359
2102
12/10/2003
13523
18360
3/12/2005
1525
1833
12/17/2003
13362
16588
3/20/2005
1565
1614
12/23/2003
13337
17636
3/26/2005
1394
1771
1/2/2004
13158
16128
4/3/2005
1673
2337
1/14/2004
18156
20742
4/17/2005
2225
1562
1/22/2004
14387
7892
4/24/2005
1935
1203
1/29/2004
17721
12777
4/30/2005
2451
1160
2/12/2004
13306
17529
5/7/2005
2382
3548
2/26/2004
12293
11076
5/16/2005
2428
1671
3/12/2004
14975
17391
5/31/2005
3718
2993
3/16/2004
14464
18522
6/6/2005
4233
3186
6/1/2004
13041
16700
6/28/2005
2469
1419
8/9/2004
13961
17174
7/5/2005
3016
8/12/2004
16567
15254
7/11/2005
3432
2102
8/17/2004
12648
16307
7/19/2005
3894
2905
8/9/2004
12983
16356
7/27/2005
2296
8/24/2004
14460
16431
8/10/2005
2639
2228
9/15/2004
14762
16696
10/5/2004
14238
16406
10/26/2004
15047
17252
11/19/2004
15995
19933
12/12/2004
13760
13121
1/6/2005
14561
13351
1/13/2005
16461
12668
1/21/2005
12567
12499
1/30/2005
16806
11683
2/5/2005
16595
12062
2/16/2005
17044
11606
2/24/2005
11526
aerobic and anaero
3C lysimeter (unit: mg/L)


20000
15000
10000
5000
0
20000
15000
10000
5000
0
e 2-10.
40
100
200
300
400
200
400
Days
600
800
hanges in alkalinity of the aerobic and anaerobic lysimeters versus time


65
Cu concentrations between the anaerobic and aerobic columns proved too high to be
comparable. Within the same anaerobic systems, As and Cr concentrations between
Jambeck and this study were not much different despite different initial As and Cr
masses. Similar leaching trends could be observed in the same anaerobic or aerobic
system, but overall concentrations of As and Cr per initial mass may depend more on the
chemistry of the system.
3.4.3 Implication for Disposal of Heavy Metals
In this research, CCA-treated wood and CRT monitor glass were used as metal
sources to represent treated wood and electronic waste. Though the land-disposal of these
wastes was banned in several states in the U.S. (SWANA, 2003), they can still be land-
disposed as a form of home appliances and ash. Since the greatest amount of As may be
leached during the first acid phase of anaerobic landfills, landfill owners need to monitor
the leachate quality during landfill construction. Though thermodynamically As may
precipitate with sulfide, a maximum of 70% of As dissolved in solution may combine
with sulfide at pH 8 (Carbonell-Barrachina, 1999). This adsorption ratio decreased as the
pH decreased. In addition to the sulfide, Fe, organic matters, and carbonate are also
known as As adsorbent. However, the adsorption efficiency of those ligands were low
relative to sulfide (Carbonell-Barrachina 1999). Thus, in the presence of an As source,
As can be found in landfill leachate for all operation periods. For aerobic landfills, As
concentrations may slightly increase during the alkaline phase. Overall As concentrations
found in aerobic landfills may be lower than those of anaerobic landfills.
Thermodynamically, Cr concentration in landfill condition during all phases may
be low. During the first acid phase, Cr may be combined with high concentrations of Fe
(II), and Cr may precipitate with sulfide during the methane phase. The lower


114
Figure 4-2. Comparison of dry weights between raw and decomposed lignocellulosic
wastes


156
Figure A-l. Schematic of mass loss by waste decomposition


Table 4-
. The physical characteristics of excavated waste
layers
depth (inches)
Wet wt
(lb)
vol (cf)
Density
(pcf)
dry density
(pcf)
Moisture
content
2-1
21-33
15.4
0.20
78.4
42
45.70%
2-2
33-45
11.5
0.2
58.46
21.7
62.9%
2-3
45-57
10
0.2
51.03
16.2
68.2%
2-4
57-66
8.9
0.15
60.46
23.7
60.7%
Lys 2
21-66
45.8
0.74
62.2
26.2
57.9%
4-1
22 28.5
8
0.11
75.2
37
50.20%
4-2
28.5-41
15
0.20
73.3
31
57.40%
4-3
41-53
12.4
0.2
63.31
25.3
59.9%
4-4
53-66
14.5
0.21
67.95
24.5
64.0%
Lys 4
28.5 66
41.2
0.61
67.1
33.5
58.8%


2.4.4Limitations 23
2.5 Conclusions 24
3. THE FATE OF HEAVY METALS IN SIMULATED LANDFILL
BIOREACTORS UNDER AEROBIC AND ANAEROBIC CONDITIONS 47
3.1 Introduction 47
3.2 Materials and Methods 48
3.2.1 Heavy Metal Sources in Synthetic Waste 48
3.2.2 Sampling Methods 49
3.2.3 Analytical Methods 49
3.3 Results and Discussions 50
3.3.1 Changes in Metal Concentrations versus Time and the Percentage of
Mass Loss 50
3.3.1.1 Aluminum 50
3.3.1.2 Arsenic 51
3.3.1.3 Chromium 53
3.3.1.4 Copper 54
3.3.1.5 Lead 56
3.3.1.6 Iron 57
3.3.1.7 Manganese and Zinc 58
3.3.2 Organic Wastes as Absorbents of Heavy Metals 59
3.4 Discussion 61
3.4.1 Overall Comparison of Metal Behavior 61
3.4.2 Comparison to Other Studies 63
3.4.3 Implication for Disposal of Heavy Metals 65
3.4.4 The Impact of Air on Metal Mobility 66
3.5 Conclusions 67
4. THE EVALUATION OF LIGNOCELLULOSIC WASTE DECOMPOSITION OF
AEROBIC AND ANAEROBIC SIMULATED LANDFILLS 95
4.1 Introduction 95
4.2 Materials and Methods 97
4.2.1 Composition of Fabricated Waste 97
4.2.2 Excavation and Processing of Decomposed Solid Waste 97
4.2.3 Methane Yield Determination 98
4.2.4 Cellulose and Lignin Determination 100
4.2.5 Data Analysis 101
4.3 Results 101
4.3.1 Methane Yield of Raw Waste 101
4.3.2 Solid Waste Excavation 102
4.3.2 Mass Loss for Individual Components 103
4.3.3 Biodegradability of Excavated Wastes 104
4.3.4 Biodegradability of Wood Waste 105
4.4 Discussion 106
4.5 Conclusions 108
vi


71
Table 3-6. Comparison of average metal concentrations of the aerobic and anaerobic
lysimeters with MSW leachate and regulatory levels (SAIC, 2000; USEPA,
1996 and 2003; Kjeldsen et al., 2002) (unit: mg/L)
MSW
leachate
Aerobic lysimeters
Anaerobic lysimeters
TCLP TC
limits
Drinking
water
standards
Acid phase
Alkali
phase
Acid phase
Methane
phase
Al
15.05
4.56
9.07
0.31
0.17
-
0.2*
As
0.44
0.19
0.55
1.5
0.43
5
0.01
Cr
0.24
0.09
0.26
0.1
0.14
5
0.1
Cu
0.14
4.01
1.75
0.02
0.07
-
1.3
Fe
3.00
61.66
3.89
188.26
10.32
-
0.3*
Pb
0.13
0.37
0.03
0.03
0.01
5
0.015
Mn
6.08
3.35
0.09
5.63
0.07
-
-
Zn
5.1
96.32
8.26
250.52
4.58
-
5*
* secondary drinking water standards
Table 3-7. Comparison of characteristics of CCA-treated wood used for Jambeck (2004)
and this study
Jambeck (2004) | This study
The % of CCA-treated included in waste stream
1%
1%
1%
As
1390 20.0
1960 27.
1330 10
Cu
814 52.4
1340 54.0
2350 50
Cr
1450 68.3
2550 48.0
2890 56


217
Table C-17 (continued)
Sample
A1
As
Cr
Cu
Fe
Mn
Pb
Zn
12/8/2004
0.10
0.78
0.05
0.01
432.43
4.82
0.04
207.50
1/6/2005
0.06
0.71
0.03
0.00
216.03
1.89
0.02
84.38
1/13/2005
0.03
1.11
0.03
0.01
247.93
2.14
0.04
90.68
1/21/2005
0.03
1.10
0.03
0.01
151.48
1.29
0.01
56.25
2/5/2005
0.00
1.31
0.03
0.02
123.73
1.07
0.01
43.28
2/16/2005
0.00
1.26
0.04
0.02
80.33
0.67
0.01
23.90
2/24/2005
0.00
1.25
0.05
0.03
66.61
0.55
0.02
18.83
3/1/2005
0.12
1.47
0.07
0.07
57.44
0.47
0.01
14.51
3/12/2005
0.00
0.72
0.06
0.04
22.78
0.18
0.00
6.42
3/20/2005
0.00
0.65
0.09
0.05
17.27
0.13
0.00
5.33
3/26/2005
0.03
0.58
0.11
0.06
15.20
0.12
0.01
4.81
4/3/2005
0.03
0.55
0.11
0.06
13.75
0.09
0.00
4.11
4/17/2005
0.02
0.47
0.10
0.05
11.66
0.08
0.05
3.78
4/24/2005
0.24
0.72
0.14
0.09
21.29
0.10
0.02
5.12
4/30/2005
0.09
0.52
0.13
0.07
12.10
0.08
0.01
4.61
5/7/2005
0.01
0.31
0.09
0.04
7.63
0.05
0.00
2.89
5/16/2005
0.13
0.51
0.15
0.08
11.25
0.07
0.01
4.43
5/31/2005
0.16
0.38
0.14
0.09
9.31
0.05
0.00
4.07
6/6/2005
0.24
0.41
0.19
0.08
8.27
0.05
0.00
4.95
6/14/2005
0.31
0.44
0.17
0.09
8.46
0.05
0.00
5.59
6/28/2005
0.21
0.33
0.15
0.07
5.83
0.03
0.00
4.37
7/5/2005
0.25
0.32
0.17
0.08
6.14
0.03
0.00
4.64
7/11/2005
0.27
0.32
0.18
0.09
6.53
0.08
0.00
4.90
7/27/2005
0.45
0.31
0.20
0.11
6.80
0.03
0.01
5.74
8/10/2005
0.39
0.27
0.19
0.11
5.62
0.03
0.01
5.74