Toxicity and hormonal activity in leachates from municipal solid waste (msw) landfills in Florida

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Toxicity and hormonal activity in leachates from municipal solid waste (msw) landfills in Florida
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Ward, Marnie Lynn
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Thesis (Ph.D.)--University of Florida, 2003.
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Includes bibliographical references.
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by Marnie Lynn Ward.
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Printout.
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Vita.

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TOXICITY AND HORMONAL ACTIVITY IN MUNICIPAL SOLID WASTE (MSW)
LEACHATES FROM FLORIDA LANDFILLS













By

MARNIE LYNN WARD


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

2003






























Copyright 2003

by

Marnie Ward






























This dissertation is dedicated to my family. To my husband, Bill, and my
daughter, Diana Mary, whose love and support fortified me on many long days
and nights. To my parents Nanny and Poppy, my sister Lisa, my brother
Jonathan and my nieces; Deonna, Megan, and Rebecca, for their presence and
guidance in my life.














ACKNOWLEDGMENTS

I wish to thank Dr. Gabriel Bitton for his guidance, patience, humor, and

unwavering steadfastness during the past eight years. He has inspired me to

investigate new ideas, question my theories, and to always ask why. I have been

awed by his wealth of knowledge and experience, while simultaneously humbled

by his understated persona. Dr. Bitton has been my mentor, my teacher, and a

model of all that I could hope to become. He continues to exemplify the strong

values and principles that are the foundation of graduate education.

Special thanks are extended to members of my doctoral committee for their

time and interest in my education and academic growth. Dr. Timothy Townsend

has been a continuous source of support and guidance. Dr. Matthew Booth

provided assistance with analytical questions and performed GC/MS analysis.

Dr. Angela Lindner maintained an "open-door" policy and was always interested

to hear updates on my research. Dr. Nancy Denslow made available

opportunities for further study and has been a reference for many research

questions.

I also extend my gratitude to my fellow students in the Department of

Environmental Engineering Sciences; both past and present. They were always

a source of support and inspiration to me, through many long nights in the lab

and even during times of personal crisis. I specifically wish to recognize the








following students; Kristin Stook, Roi Dagan, Thabet Toylamet, Diana Lee, Libby

Schmidt, Jenna Jambeck, Pradeep, Dubey, Jeff, Jenn, Tracy, and Marissa.

My heartfelt appreciation is extended to my dear friend and personal

mentor, Linda Tyson. Her house became my second home, which was a

welcome respite during my qualifying exams. I first met Linda in the early 1990s

when I was a student at Central Florida Community College and she was an

instructor. It was after listening to her talk about her graduate work that I became

inspired to attend the University of Florida. She has been an unwavering source

of inspiration and wisdom.

I owe a deep appreciation to Peter Meyers and Craig Watts at Hydrosphere

Research. They provided the starter cultures of Ceriodaphnia dubia, Daphnia

pulex, and Pseudokirchneriella subcapitata, which were subcultured to supply in-

house test organisms. Peter Meyers was a reliable and experienced source for

information concerning techniques and methods for culturing aquatic test

organisms and for conducting acute and chronic assays.

I also extend appreciation to the operators of the landfills at which the

leachates were collected. I specifically thank Jim Brunswick for his friendship

and guidance.














TABLE OF CONTENTS
Paae

ACKNOW LEDGM ENTS ................................................... ............................. iv

LIST O F TABLES................................. ..................................................... .... x

LIST O F FIG URES ................................... ....... ............................. xiv

A B ST R A C T ......................................................................... ...................... xviii

CHAPTER

1 INTRODUCTION ..................................... ......................... 1

2 LITERATURE REVIEW ............................ .................. ...................... 6

M SW Landfills ............................................................. ......... ............ 6
Modern MSW Landfills ......................................... ................................ 6
Characterization of the Chemical and Physical Composition of MSW
Landfill Leachates............................. ................. ...................... 9
Bioassays for the Evaluation of Toxicity in the Environment ...................... 14
Toxicity of MSW Landfill Leachates...................... ....................... 16
Hormonally Active Agents in the Environment......................................21
Phytoestrogens.................................................................. 23
Phthalates ..................... .................................... 24
Alkylphenols.......................... .......... ...................... 26
Natural and Synthetic Estrogens......................................................... 28
Effects of Hormonally Active Compounds on Humans.......................31
Bioassays to Identify Hormonal Activity ............................................... 33
In vivo assays for the determination of hormonal activity ................36
In vitro assays for the determination of hormonal activity ................37
A test battery to determine hormonal activity.................................43
Characterizing Hormonal Activity in MSW Landfill Leachates and Other
Environmental Samples................................................................44

3 TOXICITY OF LEACHATES FROM FLORIDA MUNICIPAL SOLID WASTE
(MSW) LANDFILLS USING A BATTERY OF TESTS APPROACH..............54

Introduction...................... ..... ..........................................................54
Materials and Methods ................... .... ..............................57
Leachate Collection ............................................ 57








Chemical and Physical Characterization of Leachates ........................58
Maintenance of Test Organisms ............................................58
Pseudokirchneriella subcapitata..........................................58
Ceriodaphnia dubia and Daphnia pulex................................. ..61
Preparation of aquatic invertebrate food........................................ 62
Maintenance of aquatic invertebrate cultures..................................63
Toxicity Assays ................................... .... ..................... 65
Pseudokirchneriella subcapitata..........................................65
Ceriodaphnia dubia and Daphnia pulex................................. ..67
M icrotoxT .............. ......................................................70
Data A analysis ..................................................... ........................ 72
Results and Discussion .............................................................................74
Chemical Analysis of MSW Leachates .......................................74
Toxicity of MSW Landfill Leachates .............................. ........... 81
Regression Analysis ........................................... .......................88
Ammonia Toxicity........................................................... ..................... 92
Influence of Site-Specific Factors on Leachate Toxicity ......................... 93

4 A SURVEY TO ASSESS THE ACUTE AND CHRONIC TOXICITY OF
LEACHATES FROM MSW LANDFILLS IN FLORIDA: ................................95

Introd auction ................................................................ ............................. 9 5
Materials and Methods ................... ................................98
Sampling Sites ................................... ...... ....................... 98
Collection of MSW Landfill Leachates....................................... 100
Chemical and Physical Characterization of MSW Leachates............... 101
Toxicity Assays .................................................. ...................... 101
Data Analysis............................................ 103
Results and Discussion .......................................................................... 104
Chemical and Physical Characteristics of the MSW Leachates........... 104
Toxicity of MSW Landfill Leachates .................................................. 117
Toxicity of MSW leachates to aquatic invertebrates...................... 117
Toxicity of MSW leachates with algae........................................... 121
Toxicity of MSW landfill leachates with Microtox ......................... 125
Heavy metal toxicity of MSW leachates using MetPLATE .............126
NOEC/LOEC vs. ECso or TU results........................................... 126
Monitoring MSW landfill leachate toxicity over time..................... 130
Comparative sensitivity of the bioassays..................................... 134
Relationship Between Chemical/Physical Leachate Characteristics and
Leachate Toxicity.................................. .............. 138

5 HEAVY METAL BINDING CAPACITY (HMBC) OF MUNICIPAL SOLID
W ASTE LANDFILL LEACHATES............................................................... 140

Introduction............................ ................................. ............. 140
Materials and Methods .......................................................... 143
Sam ple Sites..................................... ............... 143








Leachate Collection .......................... ............. ................... 143
Chemicals and Reagents.... ....... .......................................... 144
Chemical Analysis..........................................................................144
Determination of Heavy Metal Toxicity.............................................. 146
Determination of HMBC .................................................................... 146
Influence of Some Leachate Parameters on HMBC..........................149
Data Analysis.......................................................... 150
Results and Discussion .................................... ....................... 151
Heavy Metal Toxicity of Landfill Leachates........................................ 151
HMBC of MSW Landfill Leachates..................... .......... 157
Leachate toxicity as a function of time .............................................. 163
Influence of Selected Leachate Parameters on HMBC...................... 164

6 IDENTIFYING TOXICITY IN FLORIDA MSW LANDFILL LEACHATES WITH
A TOXICITY IDENTIFICATION AND EVALUATION (TIE) PROCEDURE .179

Introduction..................................... ..... 179
Material and Methods............................ .... ........................182
Sam ple C collection .......................................... ...... ..........................182
Chemical Analysis of MSW Landfill Leachates.................................... 183
TIE Procedure: Phase I..................................................................... 185
pH-adjustment of the MSW landfill leachates.............................. 185
Filtration of the MSW landfill leachates......................................... 186
Solid phase extraction of the MSW landfill leachates.................... 186
Aeration of the MSW landfill leachates .......................................... 187
Blank preparation ...................................................... 188
Zeolite test............................. ......................... 189
Toxicity Assays ................. ......................... ... ........................ 189
Initial toxicity assays...................... .......... ...................... 189
Baseline toxicity assays............................ ........................ 191
Post-manipulation toxicity assays............................................... 192
Data Analysis .... ..................................................... 192
Results and Discussion ............................................... 194
Chemical/Physical Characterization..................................................... 194
Determination of Heavy Metal Bioavailability ....................................... 195
Initial Toxicity ............................................................. 196
Blanks and Controls........................... .... ....................... 197
Baseline Toxicity .................................... ........................ 197
Effect of Zeolite Treatment................................ ............. 197
Post-manipulation Toxicity ..................................................................199
Site 7 ............................................................ 199
Site 8 ..................... .... ....... ...................... 202
Site 14 .................... ................................... 206






viii








7 HORMONAL ACTIVITY OF MUNICIPAL SOLID WASTE (MSW)
LEACHATES FROM FLORIDA LANDFILLS ..............................................209

Introduction ...................................... ...... 209
Materials and Methods .........................................................211
C hem icals ............................................ .......... ..................... 211
MSW Landfills and Leachate Collection.............................. ..........212
MSW Landfill Leachate Treatment Facility..........................................212
Solid Phase Extraction (SPE) of MSW Landfill Leachates................... 214
YES Assay for Determining Hormonal Activity.............................. 215
Toxicity of MSW Leachates to Yeast Cells...........................................217
GC/MS Analysis........ ...... ........................218
R results .............................................. ................ ........................ 2 19
Hormonal Activity of MSW Landfill Leachates....... ..................219
Effect of Biological Treatment on Hormonal Activity ..........................224
GC/MS Analysis of MSW Landfill Leachates ......................... ...........226
Influence of Concentration Factor on Hormonal Activity of Leachates.229
E2 Recovery in Spiked Methanol Extracts of Leachates....................232
Interpretation of GC/MS Results with the Hormonal Activity of MSW
Landfill Leachates...................................................234
Isolation of Hormonal Activity at LF 12................................. ... ..241
Issues Raised When Analyzing MSW Landfill Leachates for Hormonal
Activity .............................................................. ........ 242
Toxicity of MSW leachates to yeast cells.................................243
Coliform bacteria ................................... .............. .............. 246
CPRG activity issue............................. ... ..................248
Assessment of organic solvents ................... .......... ..........249

8 CONCLUSIONS ............................................. ....... 250

LIST OF REFERENCES.............................. ..... ..... .............253

BIOGRAPHICAL SKETCH....................... ..... .................. 287














LIST OF TABLES


Table page

1-1. Frequently used acronyms........................... .... .. .................... 2

2-1. Range of selected chemical and physical characteristics reported
in the literature for domestic wastewater and MSW landfill
leachates in Florida and internationally ........................................ ..... 10

2-2. Toxicity of individual constituents identified in MSW landfill leachates .......13

2-3. Reported toxicity in the literature for MSW landfill leachates.................... 17

2-4. Select phthalate compounds and their common usage ............................24

2-5. Concentrations of nonylphenols in food items in Germany.......................26

2-6. Rates for the urinary excretion of natural estrogens from men
and women.............................. ......... ... ......................29

2-7. Advantages and disadvantages associated with the use of in vivo and
in vitro assays for identifying hormonal activity ........................................ 33

2-8. In vivo assays for the determination of hormonal activity.........................34

2-9. Threshold dose for the induction of hormonal effects following
exposure of fish to natural and synthetic estrogens .................................35

2-10. In vitro assays for the determination of hormonal activity .......................37

2-11. Relative sensitivity of in vitro assays to 17 p-estradiol (E2) ....................... 38

2-12. The hormonal activity of selected metal species.......................................46

2-13. Reported phthalate concentrations in landfill leachates..........................47

2-14. Concentrations (ng/L) of natural and synthetic hormones in
wastewater treatment plants(WWTPs)...........................................49

2-15. Reported concentrations (ng/L) of natural and synthetic
estrogens in surface waters ........................ ...........................52








3-1. Amount of MSW generated and landfilled at six landfill
sites in Florida .................... .................................57

3-2. Components of the preliminary algal assay procedure (PAAP)
m edium .............. .................................... ........................ .......... 59

3-3. Physical and chemical characteristics of MSW landfill leachates
at six sites in Florida............................................. ................................ 73

3-4. Correlative analysis with the C. dubia, P. subcapitata, and MicrotoxTm
assay results versus leachate chemical characteristics ...........................87

4-1. Description of 14 MSW landfill sites where leachates were
collected .......................................................................... .......... ........ 99

4-2. Physical and chemical characteristics of MSW leachates collected
from 14 lined landfills in Florida............................................................. 107

4-3. Distribution of major ions in leachates from 14 lined MSW
landfills in Florida.......................... ...................... ..................... 110

4-4. Mean concentrations (mg/L) of total (NH4/NH3) and un-ionized (NH3)
ammonia in leachates from fourteen MSW landfills in Florida................113

4-5. Metal concentrations in leachates from fourteen MSW landfills
in Florida ............................................................... ................... 114

4-6. Toxicity of leachates collected from 14 lined MSW landfills with
C. dubia, D. pulex, and P. subcapitata.................................................. 120

4-7. Toxicity of the MSW landfill leachates from 14 sites in Florida
using the 15-minute Microtox acute assay ........................................... 124

4-8. Relationship between the toxic endpoints of ICso (%),NOEC (%)
and LOEC (%) with the results of the P. subcapitata assay
with leachate from site 1............................ ............................... 127

4-9. Coefficients of variation (CV)(%) for the P. subcapitata, C. dubia,
and D. pulex assays ....................... ...................................... 35

4-10. Classification system for ranking the toxicity of MSW landfill
leachates from 16 sites in Florida................................... ..................... 137

5-1. ECs5for Cu+2, Zn+2, and Hg*2 determined with the MetPLATE
assay ................................................................ ............................ 15 1

5-2. Toxicity of leachates from 16 lined MSW landfills using
MetPLATE............................................................................... ....... ... .....152








5-3. Physical and chemical characteristics of leachates collected
from 16 lined MSW landfills in Florida. .................................... ....... 154

5-4. Heavy metal binding capacity (HMBC) (unitless) of leachates
from 16 MSW landfills with copper, zinc, and mercury ....................... 158

5-5. MetPLATE and HMBC results with MSW landfill leachates
collected from sites 1, 4, 5, and 8........................... ...... ................ 164

5-6. HMBC of MSW landfill leachates with copper, zinc, and mercury
following fractionation.............................................................................. 165

5-7. Changes in physical and chemical characteristics during fractionation
of the MSW landfill leachates from site 1, 4, 5, and 8............................... 166

5-8. Coefficients of determination (R2) obtained between MSW landfill
leachate characteristics and the heavy metal binding capacity
(HMBC) for copper, mercury, and zinc.................................................... 175

6-1. Population served and amount of waste landfilled, as a percent
of total waste generated, at sites 7, 8, and 14.......................................... 182

6-2. Manipulations to identify suspected toxicants...........................................184

6-3. Chemical and physical characteristics of the MSW landfill leachates
from sites 7, 8, and 14........................................................ 193

6-4. The initial (day 1) and baseline (day 2) acute and chronic toxicity
of the whole MSW landfill leachates from sites 7, 8, and 14 prior
to fractionation.......................................... ........................................ 196

6-5. Ammonia concentrations in MSW landfill leachates before and after
treatment on a Zeolite cation exchange column....................................... 198

6-6. Acute toxicity of the whole and post-Zeolite MSW landfill
leachates to C. dubia neonates............................................................... 199

6-7. Summary of TIE results with MSW landfill leachates from
sites 7, 8, and 14 ......................................................... ........... ........ 208

7-1. Hormonal activity of raw MSW landfill leachates and their
methanol extracts............................................................. 220

7-2. Hormonal activity of raw MSW landfill leachates from LF 8
before (influent) and after (effluent) treatment in a powdered
activated carbon treatment (PACT) facility............................................ 224








7-3. Organic compounds tentatively identified in MSW landfill leachates
by GC/MS analysis in full scan mode................................ ..........226

7-4. Recovery (%) of 17 p-estradiol (E2) from the E2 spiked methanol
extracts of MSW landfill leachates....................... ............................232

7-5. Categories of hormonal activity in MSW landfill leachates......................234

7-6. Presence of hormonal activity in the raw leachates and methanol
extracts of MSW landfill leachates with identified hormonally active
compounds in parenthesis .................................................235

7-7. Effect of extraction procedures on the hormonal activity of leachates
from LF 12 (March 2002).......................... ............................240

7-8. Total and fecal coliform bacteria determined in MSW landfill
leachates with results expressed as the most probable
number (MPN) of bacteria/100 ml of leachate........................................ 247














LIST OF FIGURES


Figure page

1-1. Pathways for the characterization of the biological effects of MSW
landfill leachates............... ............... ...................... 3

2-1. Representation of the vertebrate endocrine system and the possible
influences of hormonally active compounds on various system
and organs .................... ..... ... .... ...................32

3-1. Locations of the MSW landfills for the collection of leachates
in Florida ................................................................. ....................... 56

3-2. Flowchart for the P. subcapitata assay .................... ......................64

3-3. Flowchart for the C. dubia assay ...................... .........................67

3-4. Flowchart for the MicrotoxT assay .............................................................69

3-5. Concentrations of total (NH4++NH3) (bars) and un-ionized ammonia
(NHa) (diamonds) in MSW landfill leachates from site 1 over a 6-month
sampling interval ................................................... 76

3-6. Concentrations of total (NH4+NH3) (bars) and un-ionized ammonia
(NH3) (diamonds) in MSW landfill leachates from site 2 over a 6-month
sam pling interval ................................................. .......................... ...... 77

3-7. Concentrations of total (NH4'+NH3) (bars) and un-ionized ammonia
(NHa) (diamonds) in MSW landfill leachates from site 3 over a 6-month
sampling interval ..................... ................................... 78

3-8. Concentrations of total (NH4'+NH3) (bars) and un-ionized ammonia
(NH3) (diamonds) in MSW landfill leachates from site 4 over a 6-month
sampling interval ..................... ............................. 79

3-9. Concentrations of total (NH4'+NH3) (bars) and un-ionized ammonia
(NH3) (diamonds) in MSW landfill leachates from site 5 over a 6-month
sam pling interval ........................ ... ................. ..................... 80








3-10. Concentrations of total (NH4+NH3) (bars) and un-ionized ammonia
(NH3) (diamonds) in MSW landfill leachates from site 6 over a 6-month
sam pling interval .............................................. ... .................................81

3-11. The mean toxicity of MSW leachates collected from six
landfill sites.................................................... .......................................82

3-12. Toxicity of MSW landfill leachates from site 1 over time......................... 82

3-13. Toxicity of MSW landfill leachates from site 2 over time........................... 83

3-14. Toxicity of MSW landfill leachates from site 3 over time......................... 83

3-15. Toxicity of MSW landfill leachates from site 4 over time .........................85

3-16. Toxicity of MSW landfill leachates from site 5 over time........................... 85

3-17. Toxicity of MSW landfill leachates from site 6 over time ...........................86

3-18. Relationship between the P. subcapitata (EC5o) and C. dubia (EC5a)
assay results with MSW landfill leachates.................................................89

3-19. Toxicity fluctuations in the leachates collected from the MSW
landfill at site 5 during February 2000 ...................... .......................90

4-1. Locations of the MSW landfills for the collection of leachates
in Florida .................................................................... .................. 98

4-2. Acute (48-hr C. dubia) toxicity of MSW landfill leachates collected
from 14 landfill sites in Florida.......................................... 116

4-3. Correlation between the 48-hour acute toxicity assays using
C. dubia and D. pulex assays with MSW leachates collected
from 14 landfill sites in Florida............................................................ 118

4-4. Chronic (96-hour P. subcapitata) toxicity of MSW landfill leachates
collected from 14 landfill sites in Florida.......................................121

4-5. Correlation between the results of the standard (125-ml) and
modified (25-ml) P. subcapitata chronic 96-hour assays.........................122

4-6. Influence of time on A.) conductivity, B.) chemical oxygen demand,
C.) total organic carbon of the MSW landfill leachates from site 1 ...........128

4-7. Influence of time on A.) conductivity, B.) chemical oxygen demand,
C.) total organic carbon of the MSW landfill leachates from site 5 ...........129








4-8. Acute (C. dubia and Microtox) and chronic (P. subcapitata) toxicity
of MSW landfill leachates collected from site 1 between February
2000 and May 2001 ........................................................... 132

4-9. Acute (C. dubia and Microtox) and chronic (P. subcapitata) toxicity
of MSW landfill leachates collected from site 5 between February
2000 and March 2001 ........................................................132

4-10. Relationship between the results of the chronic 96-hour P. subcapitata
(IC5o) and the acute 48-hour C. dubia (LCso) assays with MSW
landfill leachates from fourteen sites in Florida ...................................... 134

4-11. Ranking of sixteen MSW leachates with the results of the
Microtox (MT), P. subcapitata (P. sub), D. pulex (D. p.),
and Ceriodaphnia dubia (C.d.) assays.................................................. 138

5-1. The MetPLATE assay protocol for determining the heavy metal
toxicity of MSW landfill leachates.................................... ................... 145

5-2. The protocol for determining HMBC of MSW landfill leachates ..............147

5-3. The protocol used for fractionation of HMBC........................................ 149

5-4. MetPLATE results for MSW landfill leachates collected from site 1
and site 5 over time ................................ ... ... .................. 163

5-5. Effect of leachate treatment by filtration (Solids), DEAE resin
(Organics), and Dowex resin (Hardness) on the HMBC......................... 177

6-1. Results of the Phase 1 toxicity fractionation with the 24-hour
C. dubia assay for MSW landfill leachates collected from site 7 ............200

6-2. Results of the Phase 1 toxicity fractionation with the MicrotoxT
assay for MSW landfill leachates collected from site 7............................. 201

6-3. Results of the Phase 1 toxicity fractionation with the 24-hour
C. dubia assay for MSW landfill leachates collected from site 8 ............202

6-4. Results of the Phase 1 toxicity fractionation with the MicrotoxT
assay for MSW landfill leachates collected from site 8...........................204

6-5. Results of the Phase 1 toxicity fractionation with the 24-hour
C. dubia assay for MSW landfill leachates collected from site 14 ............205

6-6. Results of the Phase 1 toxicity fractionation with the MicrotoxT
assay for MSW landfill leachates collected from site 14........................... 207








7-1. Procedure for preparing MSW leachates and methanol extracts
of leachates for analysis of hormonal activity ......................................... 213

7-2. Response of the YES assay to 17-p estradiol (E2) ...................................219

7-3. Total ion chromatogram of MSW leachates from LF 8 (Feb. '02#1)
in the full scan mode ..................................... ... ..................... 228

7-4. Dose-response of the LF 1 (Nov. '01) methanol extracts versus
concentration factor................................................. ........ .................. 230

7-5. Dose-response of the LF 8 (Dec. '01) methanol extracts versus
concentration factor................................ ..................... 231

7-6. Dose-response of the LF 12 (Nov. '01) methanol extracts versus
concentration factor...................... .... .. .......................231

7-7. The solid phase extraction protocol used with MSW landfill
leachates........................ ......... ... ....................... 241

7-8. The toxicity of MSW landfill leachates to yeast cells according
to the INT procedure .............................................................................244














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

AN INVESTIGATION OF TOXICITY AND HORMONAL ACTIVITY IN
LEACHATES FROM MUNICIPAL SOLID WASTE (MSW) LANDFILLS IN
FLORIDA

By

Marnie Lynn Ward

December, 2003

Chair: Dr. Gabriel Bitton
Major Department: Environmental Engineering Sciences

The purpose of this research was to characterize the chemical composition

and biological effects of leachates from MSW landfills in Florida. Samples were

collected from 16 engineered landfills to encompass a cross-section of leachate

quality and characteristics. The MSW landfill leachates were tested using a suite

of bioassays, which included the chronic Pseudokirchneriella subcapitata and the

acute Ceriodaphnia dubia, D. pulex, and MicrotoxTM. Leachates were tested with

MetPLATE, a heavy metal specific assay. Additionally, using a yeast reporter

assay, the leachates were tested for hormonal activity.

Landfill leachates are complex mixtures of organic and inorganic

contaminants with compositions heavily influenced by site-specific parameters

(e.g. waste composition and age). The chemical composition of the Florida

landfill leachates varied widely. In some leachates, high levels of un-ionized








ammonia, inorganic components, CBOD, and COD were recorded. The

corresponding toxicity at these sites was high. Significant relationships were

shown between the ammonia content of the leachates and toxicity as determined

by the C. dubia (R2 = 0.62) and P. subcapitata (R2 = 0.69). The assays were

ranked for their sensitivities to the MSW landfill leachates as follows: C. dubia

D. pulex >P. subcapitata (125-ml) P. subcapitata (25-ml) > MicrotoxT.

The heavy metal toxicity/bioavailability and heavy metal binding capacity

(HMBC) of landfill leachates was determined with MetPLATE. The heavy metal

toxicity was low, which was attributed to the presence of complex-forming

ligands. The magnitude of the HMBC was investigated with the metals copper,

zinc, and mercury. The results showed that the HMBC ranged from 3 to 115, 5

to 93 and 4 to 101 for HMBC-Cu+2, HMBC-Zn*2, and HMBC-Hg*2, respectively.

The leachates were chemically/physically treated to reduce or fractionate their

complexity. Fractionation of selected leachates revealed that HMBC was

influenced by the solid, organic, and hardness content of the tested leachates.

Additionally, other unidentified components influenced the HMBC of the landfill

leachates.

The leachates were evaluated for their hormonal activity using a yeast

reporter assay. The hormonal activity of the raw MSW landfill leachates was

highly variable, with an E2 equivalent range from 2.6 to 45.7 ng E2/L. A similar

range from 6.2 to 59.7 ng E2/L was reported for the methanol extracts of the

leachates. The presence of unidentified substances in the leachates reduced the








recovery of hormonal activity. Treatment processes utilizing powdered activated

carbon (PAC) removed hormonal activity.













CHAPTER 1
INTRODUCTION


Historically, domestic wastes were disposed in open pits or surface piles,

and these sites often posed a significant risk to the surrounding environments

(Reinhart and Townsend, 1998). Concerns with the environmental fate of

discarded materials led to the construction of engineered systems for long-term

storage and management of waste materials and their degradation products.

These engineered systems included lined cells for the disposal of waste and

collection systems for leachate recovery. Government regulations prohibit the

disposal of waste materials except at regulated municipal solid waste (MSW)

disposal facilities. Although MSW landfills may contain small quantities of

hazardous waste materials, they are primarily designed to receive domestic

wastes. These items may include packaging materials, food scraps, furniture,

clothing, and grass clippings (USEPA, 2002).

The acute and chronic toxicity of MSW landfill leachates has been

extensively studied using various bioassays (Rutherford et al., 2000; Clement et

al., 1997; Kaur et al., 1996; Plotkin and Ram, 1984); however, little information is

available concerning the biological effects of MSW leachates collected from

landfills in Florida. Reinhart and Grosh (1998) have reported a lower chemical

strength in Florida MSW leachates, resulting from the predominant environmental

conditions, e.g., abundant rainfall and warm temperatures. Therefore, the basis








Table 1-1. Frequently used acronyms
Acronym Definition Acronym Definition
Carbonaceous Hormonally active
CBOD biochemical oxygen HAA agent
demandagent
COD Chemical oxygen HAC Hormonally active
demand compound
CPRG Chlorophenol red HMBC Heavy metal binding
galactopyranoside capacity
Sample concentration
DOC Dissolved organic IC responsible for 50 %
carbon inhibition in test
organism
Sample concentration
DOM Dissolved organic LC5 lethal to 50 % of the
matertest organisms
E1 Estrone MSW Municipal solid waste
E2 17 p-estradiol MHW Moderately hard water
E3 Estriol ONPG Ortho-nitrophenyl
galactopyranoside
EE2 17 a-ethinyl estradiol TDS Total dissolved solids
Sample concentrationcity identification
ECso responsible for a 50 % TIE To identification
effect in test organism
ED Endocrine disrupter TRE Toxicity reduction
evaluation
ERa Estrogen receptor alpha TS Total solids
United States
ERp Estrogen receptor beta USEPA Environmental
Protection Agency
Estrogen receptor -
ER-CALUX chemically activated YES Yeast estrogen
luciferase reporter gene sc


of this investigation was to determine the biological affects associated with


exposure to MSW leachates from Florida landfills.








ACUTE CHRONIC
TOXICITY TOXICITY




LEACHATE




I ME-AL
HORMONE BINDING
ACTIVITY ME
METAL
BIOAVAILABILITY




Figure 1-1. Pathways for the characterization of the biological effects of MSW
landfill leachates

In recent years, the number of operating landfills in the U.S. has

decreased; however, the overall size of the remaining landfills has increased

(USEPA, 2002). Landfills continue to be the most economically feasible and

least environmentally intrusive method for the disposal of discarded waste

materials. In Florida, more than 14 million tons (total 24.8 million tons) of MSW

were landfilled in 1998, the most recent year data were available, with a per

capital generation rate of 9.1 pounds/person/day (FDEP, 2000).

While the chemical and physical composition of MSW landfill leachates in

Florida have been summarized (Reinhart and Grosh, 1998), only one study has

evaluated the toxicity of Florida leachates (Ward et al., 2000) and then only on a

very limited scale. The scientific community, environmental regulators, landfill

operators, and the operators of facilities treating landfill leachates require








comprehensive databases to provide information relative to biological effects and

chemical constituents of Florida MSW landfill leachates.

Numerous acronyms are frequently used in the scientific research

community, and many are commonly recognized; however, some are relatively

new and discipline-specific. Therefore, a table of the acronyms used in the

following chapters has been included to aid the reader as a quick reference

(Table 1-1).

The purpose of this research was to evaluate MSW landfill leachates in

Florida, for their toxicity, hormonal activity, and chemical characteristics. To

date, there have been no reported investigations of MSW landfills of this scope or

magnitude. Most investigations have focused on leachate from one or several

landfills (Kaur et al., 1996; Wong, 1989), but few evaluated multiple leachates

(Clement et al., 1996, 1997), and none have tracked patterns over time. The

toxicity of leachates from Florida landfills has received little attention. Ward et al.

(2000) evaluated the toxicity of landfill leachates from three Florida landfills;

however, leachates were evaluated with only one acute toxicity assay.

The basis of this research project was to provide a statewide database

for landfill operators, regulators, and other research investigators, that

established a range of biological effects from exposure to MSW landfill leachates.

The overall objectives of this research project were as follows (Figure 1-1):

1. To characterize the toxicity in the MSW landfill leachates using a battery of

toxicity assays including the chronic 96-hr Selenastrum capricomutum, and





5

the acute 48-hour Ceriodaphnia dubia, 48-hour Daphnia pulex, and

MicrotoxTM assay (Chapters 3 and 4).


2. To characterize the chemical and physical composition of MSW landfill

leachates (Chapters 3, 4, 5, 6, and 7).


3. To determine the toxicity of heavy metals in MSW landfill leachates with the

heavy metal specific MetPLATE assay and to quantify the ability of MSW

leachates to reduce the bioavailability of heavy metals with a heavy metal

binding capacity (HMBC) assay (Chapter 5).


4. To conduct a toxicity identification evaluation (TIE) with selected MSW

landfill leachates (Chapter 6).


5. To determine the hormonal activity of MSW landfill leachates with a yeast

estrogen screen (YES) and identify by gas chromatography/mass

spectrometry (GC/MS) organic compounds responsible for hormonal

activity (Chapter 7).


6. To determine the effect of powdered activated carbon treatment on

hormonal activity in landfill leachates (Chapter 7).


7. The results obtained during this research investigation are summarized

(Chapter 8).













CHAPTER 2
LITERATURE REVIEW

Forty years after Rachel Carson first revealed the risks to humans and the

environment posed by a diverse array of man-made substances, the threat from

these anthropogenic chemicals persists (Carson, 1962). Over 100,000 synthetic

chemicals including pesticides, solvents, domestic cleaners, plasticizers, and

flame-retardants are produced yearly for domestic and industrial usage; however,

little is known about the biological effects from long-term exposure to these

compounds (Darnerud et al., 2001; Hale et al., 2001; Lyytikainen et al., 2001). In

the US, the Resource Conservation and Recovery Act (RCRA), Toxic

Substances Control Act (ToSCA), and the Clean Water Act (CWA) and their

amendments have been effective in the control of point source pollution.

However, non-point sources of pollution continue to pose significant threats to

the environment.

MSW Landfills

Modem MSW Landfills

MSW landfills are engineered systems that are designed through the use

of heavy-duty plastic liners and/or low permeability clay barriers to retard the

escape of pollutants to surrounding environments and for the recovery of

leachates. Leachates are formed following rainwater infiltration. Percolation of

this water through the waste materials mobilizes soluble substances (Ross,

1990). These leachates represent the mobile fraction of landfill toxicants, and








they contain high concentrations of inorganic and organic compounds (Bozkurt et

al., 2000). The majority of MSW landfill leachates are closely regulated;

therefore, their release of leachates from modern landfills is unlikely (Barlaz et

al., 2002). Accidental leachate releases may occur, e.g. during periods of high

rainfall when leachates are contaminated by stormwater or from the improper or

faulty installation of leachate collection systems. Additional sources for the

unintentional release of MSW landfill leachates exist. Prior to 1990, federal

regulations did not require landfill liners, therefore, the escape of leachates from

these sites represents a potential adverse environmental impact (Assmuth,

1996). Florida was more proactive and required landfill liners for MSW landfill in

the Solid Waste Act of 1988.

Landfilling remains the predominant management method for municipal

solid waste (MSW), accounting for more than 50% (128.3 million tons) of the

MSW generated in the United States (USEPA, 2002). Under authority granted by

RCRA, subtitle D, the U.S. Environmental Protection Agency (USEPA) regulates

the construction, operation, and post-closure of municipal solid waste (MSW)

landfills (CFR 258, 1996). Strict regulations require landfill operators to minimize,

recover, and treat the leachates generated in MSW landfills. Landfill liners are

used to restrict the flow of leachates to ground and surface waters, and they may

be composed of clay, high or low-density polyethylene, or concrete, depending

on the local conditions. In some landfills, leachates are recycled through the

waste to encourage waste stabilization and improve leachate quality (Reinhart

and Townsend, 1998; Nopharatana et al., 1998).








The chemical characteristics of MSW leachates are dependent on the type

and amount of waste landfilled, the landfill age and various environmental

conditions, e.g. temperature and rainfall. Environmental regulations in the U.S.

greatly limit the disposal of hazardous wastes in municipal landfills, but waste

materials containing toxic chemicals may enter many landfills. Some sources of

these substances include generators of small amounts of hazardous waste, non-

hazardous industrial wastes, and household hazardous wastes (HHW), such as

electrical devices, fluorescent light bulbs, thermometers, batteries, pesticides,

and other chemical products (Boyle and Baetz, 1993). Using a risk-based

assessment, comparable health risks were associated with MSW landfill

leachates and industrial waste leachates (Brown and Donnelly, 1988). Their

assessment was based on the individual leachate constituents, while not

accounting for potential synergistic and/or antagonistic responses (Brown and

Donnelly, 1988).

Landfills are classified based on the composition of the waste materials

landfilled. Class I and II landfills are designated for the disposal of non-

hazardous household wastes and some commercial, industrial and agricultural

wastes, while Class III landfills are designed for yard waste, construction and

demolition debris, carpet, furniture, and similar non-putrescible waste materials.

The basis for the distinction between Class I and Class II landfills is the volume

of MSW landfilled daily. Class I landfills receive 20 or more tons of MSW per

day, while Class II landfills less than 20 tons of MSW per day.








Characterization of the Chemical and Physical Composition of MSW
Landfill Leachates

Waste stabilization is based on microbial degradation processes, with the

conversion of organic matter to methane gas occurring predominately under

anaerobic conditions (Barlaz, 1997). Initially, aerobic conditions dominate in

landfills, but oxygen is rapidly depleted with the continuous addition of waste.

The phases of waste degradation are loosely defined by the predominant

chemical characteristics. Early phases are defined by the transition from aerobic

to anoxic and finally strictly anaerobic conditions within the waste.

Concomitantly, microbial degradation converts the large organic molecules to

smaller organic acids, which are further degraded to hydrogen and acetate.

Methanogenesis, the conversion of these small molecules to methane gas by

methanogenic bacteria, is a strictly anaerobic process. Researchers theorize

that landfills may revert to aerobic conditions as oxidized micro-environments

begin to form at the boundaries of the waste. However, no landfill currently

under study has reached this stage of decomposition (Bozkurt et al., 2000;

Kjeldsen et al., 2002). As landfilling is a continuous and on-going process,

various stages of waste decomposition occur simultaneously, and this may be

reflected in the chemical characteristics of the leachates.

MSW landfill leachates are complex mixtures, with a wide ranging

chemical strength (Table 2-1). In a recent review, Kjeldsen et al. (2002)

summarized the chemical and physical characteristics of MSW landfill leachates.








Table 2-1. Range of selected chemical and physical characteristics reported in
the literature for domestic wastewater and MSW landfill leachates in
Florida and internationally
Florida MSW International MSW
Parameter Raw Wastewater Landfill Leachate Landfill Leachate
BODs 110-400c 0.3-4800e 42-10,900a
(mg/L)
COD 250-1,000c 7-50,000e 40-90,000a
(mg/L)
Alkalinity 50-200c 350-8775' 1,350-3,510b
(mg/L)
pH -6.2-9.7e 3-7.9a
NH-N 12-50c 0-4110e <0.3b-13,000d
(mg/L)
Chloride 20-500 1.9-2720e 125-2,400b
(mg/L)
Sulfide <0.01-3.8e <0.02-30b
(mg(L)
TDS 250-850c 1,800-31,700' 2,000-60,000g
(mgfL)
Na* 40-70c 25.6-1963' 128-840b
(mg/L)
Ca+2 45-4,400' 10-7,200g
(mg/L)
Mg+2 25-122' 30-15,0009
(mg/L)
Total phosphorus 4-15c 0.1-39.6' <0.01-2.7.
(mg/L as PO42)
Conductivity 3.4-39.6' 1,200-16,000a
(mS/cm)
References: aKadlec and Knight, 1996; 'Cameron and Koch, 1980; "Metcalf and Eddy;1991; 0Lo,
1996; eReinhart and Grosh, 1998; 'this research; gKjeldsen et al., 2002.

They stressed four primary categories of contaminants to consider in discussions

of leachate quality, and these were dissolved organic matter, organic xenobiotics,

inorganic components, and heavy metals (Christensen et al., 1994).

In MSW landfill leachates, dissolved organic matter (DOM) includes the

dissolved and colloidal particles (Gounaris et al., 1993). The molecular structure

and elemental composition of dissolved organic matter in MSW landfill leachates

is strongly influenced by microbial degradative processes. Calace et al. (2001)








reported a narrow distribution of organic molecular weight groups in young

landfills (< 5 years old), with primarily low molecular weight constituents (<500

Dalton). This contrasted with their findings in older landfills (>10 years old)

with an increased distribution of molecular weight fractions and high molecular

weight constituents (> 10,000 Dalton) (Calace et al., 2001). These high

molecular weight fractions contain structurally complex humic materials (Croue et

al., 2003). Kang et al. (2002) reported an increased presence of humic

substances with increasing landfill age and a decrease in the easily degraded

lower molecular weight organic materials. Attributed to the high ammonia

concentrations in the leachate, the humic materials contained a large distribution

of nitrogen functional groups (Kang et al., 2002). This is significant when

considering the strong metal complexes that are formed with organic ligands

containing nitrogen functional groups (Croue et al., 2003; Stumm and Morgan,

1995).

Xenobiotics are frequently detected at low levels in MSW landfill leachates

(Schwarzbauer et al., 2003; Kawagoshi et al., 2002; Yasuhara et al., 1999).

Some xenobiotics of significant environmental concern, relative to reproductive

effects, have been identified in MSW landfill leachates. Wintgens et al. (2003)

reported nonylphenol, a surfactant, and Bisphenol A, a plasticizer, at 60 and 37.5

ig/L, respectively, in MSW landfill leachates. In a Japanese landfill, xenobiotics

identified included Bisphenol A, nonylphenol, octylphenol, and some dioxin-like

substances (Behnisch et al., 2001). The organic contaminants in MSW landfill

leachates have been summarized (Kjeldsen et al., 2002).








Inorganic contaminants in MSW landfill leachates include anionic and

cationic species, and some of the most widespread are NH4*, Ca+2, Mg'2, Na1,

ClI', HCO3-1, S04-2. Typically, the ammonia in MSW landfill leachates occurs as

the ionized ammonium (NH4*) species. Ammonia speciation is pH dependent,

with a pKa of 9.3. Therefore, the dominant species is ammonium, rather than the

highly toxic ammonia (NH3) form (McBean et al., 1995). Generally, total

ammonia concentrations are high, with reports of up to 740 mg/L (Kjeldsen et al.,

2002). These high ammonia levels are often attributed to the absence of

degradative pathways for the removal of ammonia from landfills (Burton and

Watson-Craik, 1998). Also found at high concentrations are the hardness

cations. Hardness is a measure of multivalent metallic cations; e.g. Ca *2, Mg2,

Sr'2, Fe.2, Mn+2, but mainly Ca2 and Mg+2(Stumm and Morgan, 1995).

Hardness has a strong influence on heavy metal bioavailability in landfill

leachates (Heijerick et al., 2003).

Heavy metals are generally reported in the low mg/L range (Reinhart and

Grosh, 1998). However, this represents only a fraction of the total metal

associated with the waste. In some cases, metals that are strongly associated

with waste materials are not easily leached under landfill conditions. Flyhammer

(1995), in a mass balance on cadmium in Swedish landfills, concluded that the

total concentrations associated with the landfilled wastes were up to four orders

of magnitude greater than leached concentrations (Flyhammer, 1995).

Kjeldsen et al. (2002) discussed leachate characteristics in MSW landfills

with a primarily organic composition and the influences of time on these leachate








Table 2-2. Toxicity of individual constituents identified in MSW landfill leachates
S C dubia S. capricomutum MicrotoxT D. magna MetPLATET
Compound (L) (mg) (mL) (mgL) (mgL)
Copper 0.01b 0.04 7.4 0.02 0.22
5-min ECo 48-hr LC5o

Cadmium 0.05c 0.34- 0.03

Zinc -0.18h 12' 5.1 0.11J
15-min EC5o 48-hr LC5o
Total 3607'
ammonia 5-min ECo
Unionized 1.1a 1.7' 3.3e
ammonia 15-min ECso 24-hr ECso
14.5d
Manganese (MHW)
Alkalinity 7819 9219
(HCO3)
Chloride -

Sodium
renamed Pseudokerchneriella subcapitata, References: 'Andersen and Buckley, 1998; Jung,
1995; CRhodes,1992; dLassier et al., 2000; "Clement and Merlin, 1995; 'Qureshi et al., 1982;
"Hoke et al., 1992; hChen et al., 1997;'Doherty et al., 1999; 'Bitton et ., 1994.


characteristics. In earlier work, Bozkurt et al. (2000) discussed leachates

generated in MSW landfills with either a primarily organic or inorganic waste

composition. The latter case represented ash monofills or co-disposal facilities

for MSW and ash. The focus of their investigation was to predict the long-term

fate of heavy metals using a conceptual model, which included influences from

various organic and inorganic ligands (Bozkurt et al., 2000).

The toxicity of MSW landfill leachates is heavily influenced by chemical

and physical interactions (Table 2-2). In this regard, complexation reactions

have a mitigating influence on toxicity and are frequently underestimated and

poorly understood, due to the sheer number of possible ligands in solution

(Martensson et al., 1999; Stumm and Morgan, 1995). Some of the typical








inorganic ligands in leachates include carbonate (Sletten et al., 1995), chloride

(Bolton and Evans, 1991), and sulfide ions (Bozkurt et al., 2000). These

inorganic ligands can form insoluble precipitates with heavy metals (Majone et

al., 1996). Other ligands present are dissolved organic matter (Kaschl et al.,

2002) and colloidal solids (Gounaris et al., 1993). These complexes exert a

strong influence on heavy metal toxicity (Heijerick et al., 2003), with up to 98 % of

metals in some landfills present as organo-metallic complexes (Kang et al., 2002;

Weng et al., 2002). However, there have been some questions concerning the

toxicity of these organo-metallic complexes (Palmer et al., 1998). Fraser et al.

(2000) suggested low-level toxicity associated with complexes between copper

and dissolved organic materials (DOM).

Bioassays for the Evaluation of Toxicity in the Environment

Luoma (1995) described bioassays as tools for investigating the complex

continuum of biochemical, physiological, and reproductive responses that occur

in organisms following exposures to suspect toxicants. Traditionally, bioassays

with various vertebrate and invertebrate organisms were used to monitor and

track environmental perturbations (USEPA, 1993a, 1994a). These bioassays

measured chronic effects in low-level, long-term exposures and acute toxicity in

high-level, short-term exposures. Critically important to all environmental

researchers was how best to reconcile chronic low-level environmental exposure

with laboratory investigations utilizing high dose acute substances (Mowat and

Bundy, 2001; Degen and Bolt, 2000). Some of the most common bioassays

have used algae (Eullaffroy and Vemet, 2003; van der Heever and Grobbelaar,

1998), aquatic plants (Mohan and Hosetti, 1999; Klaine and Lewis, 1995),








invertebrates (Heijerick et al., 2003; Kim et al., 2003; Preston and Snell, 2001;

Pereira et al., 1999), or microorganisms (LeBlond et al., 2001; Doherty et al.,

1999; Jung et al., 1997).

While most bioassays have been extensively validated, each has intrinsic

advantages and disadvantages that are specific to the method. For example, in

some algal assays, cell exudates extracellularr organic material) mitigate metal

toxicity by acting as ligands that form complexes with free metal ions. In

microplate assays, these and similar problems are controlled, which may explain

the recent increase in the use of microbiotests. Additionally, these microbiotests

offer an increased affordability, portability, and the availability of results in a short

interval of time (Chial and Persoone, 2002; Bitton et al., 1994). Gabrielson et al.

(2003) recently developed a microplate assay, referred to by the acronym MARA

(microplate assay risk assessment), that utilized 11 lyophilized microbial strains

for determining the toxic fingerprint of a chemical. This assay allows for the

testing of multiple species simultaneously; however, it is not sensitive to any

specific class of toxicants. There are microplate assays (e.g., MetPLATE and

MetPAD) that are designed specifically for the detection of heavy metal toxicity

(Bitton et al., 1992b, 1994).

Developments in the field of environmental chemistry have produced

analytical methodologies and techniques that are highly successful at identifying

and quantifying contaminants, even in highly complex matrices (Richardson,

2001). They use a suite of analytical tools that have in common an electrode,

which senses changes based on electronic signals. Some typical electrodes








measure dissolved oxygen (DO), conductivity, pH, select ions (e.g. ionized

metals), and oxidative potential. Parallel techniques for application in the field of

environmental toxicology would allow for the rapid identification of toxicity, while

simultaneously reducing the time and cost involved in continuous monitoring

programs (Arikawa et al., 1998). Biosensors are a rapid and convenient

monitoring tool, which incorporate biological tissues in a system highly sensitive

to a broad spectrum of toxic substances (Botre et al., 2000; Buffle and Horvai,

1998; Argese et al., 1996). Although their use is currently limited, biosensors

have been successfully applied to the evaluation of wastewater toxicity (Farre

and Barcello, 2003).

Toxicity of MSW Landfill Leachates

Research investigations that simultaneously combine bioassays with

methods for chemical characterization are highly valued, but the associated

expenses and labor demands constrain their extensive utilization (Ferrari et al.,

1999; Atwater et al., 1983). To date, the most extensive study of waste

leachates was conducted in France. Clement et al. (1996) investigated the

toxicity of ten domestic landfill leachates and various other hazardous and non-

hazardous waste leachates. Bioassay results with protozoa, bacteria, algae, and

invertebrates demonstrated that the toxicity of the domestic waste leachate was

higher than the industrial or hazardous waste leachates (Clement et al., 1996).

Furthermore, the chemical characterization of the domestic leachates revealed

that ammonia, alkalinity, conductivity, and COD were highly associated with

increased leachate toxicity (Clement et al., 1997). In earlier work, a significant

contribution of ammonia to the acute toxicity of landfill leachate to duckweed








Table 2-3. Reported toxicity in the literature for MSW landfill leachates
Leachate Origin Species Endpoint Reference
Solid waste landfill Tilapia
Soidnwat aditio) (Sarotherodon 96-hr LC = 1.4 to 12% Wong,
(unknown composition) mossambicus) 1989
mossambicus)
Solid waste landfill
(40 % household and Fathead minow Plotkin and
60 % (Pimephales 96-hr LC50 = 100 % Ram, 1984
industrial/commercial) promelas)
Solid waste landfill
(40 % household and Plotkin and
60 % Daphnia magna 48-hr LCo = 62 to 66 % Ram, 1984
industrial/commercial)
Solid waste landfill
(40 % household and Selenastrum Plotkin and
60% capricomutum 13 day EC5= to 10 % Ram, 1984
industrial/commercial)
Solid waste landfill
(40 % household and Plotkin and
(nnw60 % Microtox 5 min. ECso= 17 % Ram, 1984
60 % Ram, 1984
industrial/commercial)
Devare
Solid waste landfill and
(unknown composition) Aquaticplants ECo = 10% Bahadir,
1994
Devare
Solid waste landfill and
(unknown composition) Microtox ECo= 18-35% Bahadir,
1994


(Lemna sp.) was reported (Clement and Merlin, 1995). Clement et al. (1997)

performed correlative analyses between their bioassay results and various

chemical characteristics and revealed a strong correlation (R2 = 0.92) between

Daphnia magna and combined ammonia and alkalinity concentrations. Similar

relationships were shown between the bioassay results with aquatic plants,

algae, and other crustaceans and ammonia and alkalinity (Clement et al., 1997).

Using the MicrotoxTm assay, the relationship between COD and toxicity was








significant (p< 0.01) but weaker (R2 = 0.58). This was the only assay sensitive to

toxicity associated with increasing organic content (Clement et al., 1997).

The toxicity of MSW landfill leachates have been well characterized

around the world (Ernst et al., 1994; Lambolez et al., 1994; Devare and Bahadir,

1994; Cheung et al., 1993; Wong, 1989; Radi et al., 1987; Plotkin and Ram,

1984; Atwater et al., 1983; Millemann and Parkhurst, 1980; Cameron and Koch,

1980), but the toxicity of Florida landfill leachates remains poorly characterized

(Table 2-3). Ward et al. (2002) studied the leachates from six MSW landfill

leachates in Florida and concluded that the leachates were highly toxic. The

toxicity of the MSW landfill leachates varied widely, due to site-specific chemical

characteristics of the leachates. Furthermore, on a monthly basis, fluctuations in

leachate toxicity indicated the heterogenous composition of the waste materials

and local conditions (Ward et al., 2002).

Recently, investigations of waste leachates in countries that do not require

landfill liners, minimization of leachate generation, or leachate collection and

treatment have been reported. These studies offer insight to researchers

concerned with the potential for leachate release to the environment. Magdaleno

and De Rosa (2000) characterized leachates from a waste dump in Argentina

with an algal assay using Selenastrum capricomutum (renamed

Pseudokirchneriella subcapitata), while Sisinno et al. (2000) evaluated waste

leachates in Brazil with the Zebrafish (Brachydanio rerio). The chemical strength

of these leachates were comparable to reports of others (Kjeldsen et al., 2002;

see Table 2-1).








The chemical strength of the Argentinian leachates was demonstrated by

COD concentrations from 502 to 4640 mg/L, ammonia from 26.5 to 35 mg/L, and

pH from 7 to 7.3 (Magdaleno and De Rosa, 2000). Similar chemical

characteristics were reported with the leachates from Brazil, with conductivity

values from 3.1 to 6.2 mS/cm, alkalinity from 212 to 372 mg/L as CaCO3, and

COD from 5,200 to 11,500 mg/L (Sisinno et al., 2000). Overall, the toxicity of the

Argentinian leachates was low with TU (toxicity unit) values that ranged from 1 to

2.1 (Magdaleno and De Rosa, 2000). Higher toxicity was reported in the

Brazilian leachates and ranged from 17.5 to 45.5 TU (Sisinno et al., 2000). The

leachates from Brazil demonstrated a toxicity similar to that reported with Florida

leachates (Ward et al., 2002). However, the leachates from Argentina displayed

reduced toxicity. The predominance of plastics and other disposable materials in

US MSW landfills may be a factor contributing to their higher toxicity.

The contamination of groundwater by waste leachates is a primary

concern, relative to the escape of leachates into the environment. Baun et al.

(1999) investigated the toxicity of groundwater contaminated by MSW leachates

in Denmark with an algae assay, a crustacean assay, and a bacterial

genotoxicity assay. Using the algae assay, the leachate-contaminated

groundwater sample displayed an EC20 of 17 %; however, the toxicity decreased

by 75 % at twice the distance from the landfill. Similar toxicity was demonstrated

by the crustacean, Daphnia magna, with an EC20 of 18 % at the landfill, but,

further downstream, no toxicity was reported. Further analysis of this

contaminated groundwater revealed that organic contaminants were responsible








for the high toxicity, and with increasing distance from the landfill the organic

toxicity decreased, suggesting metabolic degradation or dilution effects (Baun et

al., 2000). Additional bioassays with the organic fraction revealed a low

sensitivity of the D. magna assay to organic toxicants, which contrasted with the

algal and MicrotoxT assay results (Baun et al., 2000). Other researchers have

reported the higher sensitivity of MicrotoxTM to organic contaminants (Bitton et al.,

1994).

In 1988, reports of comparable carcinogenic risk associated with exposure

to MSW landfill leachates or hazardous waste leachates raised concerns in the

regulatory community (Brown and Donnelly, 1988). Subsequent investigations,

to determine the genotoxic potential of MSW landfill leachates, have revealed

conflicting results. Beg and AI-Muzaini (1998) investigated the genotoxicity of

MSW landfill leachates in Kuwait using a dark mutant strain (nonluminescent) of

Vibrio fisher, a bioluminescent bacterium. In the presence of a mutagen, the

dark strain reverts to the luminescent state and this response is quantified by

increased light intensity. These results suggested, in some of the Kuwaiti

leachates there was a high degree of genotoxicity, and this was dependent on

the type of waste landfilled and seasonal conditions (Beg and AI-Muzaini, 1998).

Helma et al. (1996) used four bacterial assays to characterize genotoxicity in

landfill leachates, wastewater effluents, pulp and papermill effluents, and

contaminated groundwater. Overall, the highest genotoxicity was displayed by

the MSW landfill leachates with more than 35,000 revertants/L of leachate. This

was comparable to the genotoxicity of the leachates produced by mixed industrial








and domestic wastes, which were reported as approximately 40,000 revertants/L

(Helma et al., 1996). Baun et al. (1999) using the umuC strain of Salmonella

typhimurium showed that leachate-contaminated groundwater was not genotoxic

at concentrations up to 25% by volume. However, bacterial toxicity at higher

concentrations prevented the evaluation of genotoxic effects. After isolation of

the organic fraction of these contaminated groundwaters, a similar mutagenicity

was identified. These results suggested that the organic fraction contained the

mutagens (Baun et al., 2000).

Hormonally Active Agents in the Environment

Some natural and anthropogenic substances may interact or interfere with

the nuclear receptors and chemical messengers of the endocrine system and

have been identified as a threat to the environment and wildlife (McLachlan,

2001). Alterations in both the developmental and reproductive functions of cells

and whole organisms are increasingly documented and attributed to exposure to

these exogenous substances (NRC, 1999). These substances are ubiquitous

environmental contaminants, which are commonly referred to as hormonally

active agents (HAAs) or compounds (HACs), xenoestrogens, estrogen-like

compounds, endocrine disruptors, estrogen-mimics, or estrogen

agonists/antagonists. Substances are labeled based on their interaction with

and/or displacement of an endogenous hormone from its conservative function

(McLachlan, 2001). Recent congressional mandates in the Safe Drinking Water

Act (1996) (Bill No.S.1316) and the Food Quality Protection Act (1996)(Bill

No.P.L.104-170) have required the USEPA to evaluate the hormonal activity of

all chemicals produced in the U.S.








In a recent survey of 139 contaminated U.S. surface waters, alkylphenols,

phthalate compounds, and natural and synthetic estrogens were shown to

comprise roughly 75% of the organic contaminant load (Kolpin et al., 2002). Of

critical concern is the exposure to these three classes of compounds, because of

reports of hormone-like effects. Their origin may be due in part to agricultural

non-point source runoffs (Casey et al., 2003), but the majority are discharged

from domestic and industrial wastewater treatment plants (WWTPs) (Sheahan et

al., 2002b; Snyder et al., 2001; Baronti et al., 2000; Rudel et al., 1998).

Environmentally relevant concentrations of these compounds have been linked to

altered sexual characteristics (Jobling et al., 1995) and elevated tissue levels of

hormonally active compounds (Sheahan et al., 2002b) in fish. However, links are

not easily established in some situations (Jacobsen and Guildal, 2000; Fawell et

al., 2001; Sepulveda et al., 2002).

According to the European Union Scientific Committee, hormonally active

agents (HAAs) (referred to as endocrine disruptors) are "exogenous substances

or mixtures that alter functions) of the endocrine system and consequently

cause adverse health effects in an intact organism, or its progeny, or sub-

populations" (Baker, 2001). Sweeping in its brevity, this definition fails to address

some concerns (McLachlan, 2001; Ashby, 2000), specifically, altered cellular

functions in relation to the overall health of the organism. Although, the National

Research Council (1999) was even less direct when they defined hormonally

active compounds as any "substance that possesses hormone-like activity,








regardless of the mechanism" of action. In light of new research (Wu et al.,

2003), changes to the definition may read, "any substance or influencing factor."

Hormonally active compounds are arranged in three groups; the natural

and synthetic estrogens, anthropogenic chemicals, and phytoestrogens (naturally

produced substances in plants). The effects of these substances may be

agonistic or antagonistic. Agonistic hormonally active compounds act in a

manner similar to an endogenous hormone, while antagonists block the activity

of endogenous substances. Beginning with reports of the estrogen-like effects

following exposure to the insecticide DDT (Burlington and Lindeman, 1950),

researchers continue to study the interaction of non-steroidal compounds with

the estrogen receptor (Miksicek, 1994).

Phytoestrogens

Phytoestrogens are naturally occurring compounds in plants and plant-

derived products (Nilsson, 2000) and include genistein, equol, formononetin,

biochanin A (Latonelle et al., 2000). The hormonal activity of phytoestrogens has

been reviewed, with special emphasis on environmentally relevant dosages

(Nilsson, 2000). One source of phytoestrogens is the urine of vegetarians (Fotsis

and Adlercreutz, 1987). Phytoestrogens impact the reproduction and sexual

health of wildlife (Hughes, 1988), but there is no evidence in humans (Strauss et

al., 1998). In fact, limited evidence suggests phytoestrogens are beneficial in

treating some types of human cancers (DiPaola et al., 1998). Ju et al. (2000)








Table 2-4. Select phthalate compounds and their common usage
COMPOUND USES
Rain gear, footwear, upholstery materials, I.V.
Di-ethylhexyl phthalate fluid bags, waterproof gloves
(DEHP) Heat seal coating on metal foils used on portioned
food items.
Dispersant in insect repellants and perfumes
Butyl benyzl phthalate Component of cellulose plastics
(BBP) Floor tiles
Di-butyl p e Coatings on cellophane, insect repellants,
Di-butyl phthalate Hair spray
(DBP) Carpet backing
Cellulose acetate plastic films- used as
Di-ethyl phthalate carton windows to display foods
(DEP) Molded plastics, i.e. toothbrushes, car
components and children's toys
Di-isononyl phthalate Vinyl wall coverings, toys, and medical devices
(DINP)


showed low concentrations of some plant substances reduce estrogenic effects,

while at high doses estrogenic effects may be increased. The contamination of

foodstuff by zearalenone, a fungal phytoestrogen, is common and human

consumption is estimated at 3 pg/person/day in North America (McLachlan,

2001).

Phthalates

Phthalates are plasticizers that are commonly used as softeners in the

production of paints, inks, adhesives, and various plastic goods (Table 2-4). In

an extensive study of phthalate compounds, the National Institute for Health

(NIH) concluded that benzyl butyl phthalate (BBP) was both a developmental and

reproductive toxicant (NIH, 2003). Additionally, extensive phthalate

contamination has been reported for over-the-counter beauty products

(Environmental Working Group, 2002). With concem, researchers have shown








that body burdens of phthalate compounds in women of child-bearing age (20-40

years) are higher than males and any other age group (Blount et al., 2000), and

the long-term consequences of this are unknown.

Phthalate compounds are ubiquitous contaminants of both terrestrial and

aquatic environments. Freshwater levels of di-ethyl hexyl phthalate (DEHP) and

di-butyl benzyl phthalate (DBP) ranged from 4.6 to 90.5 1pg/L and 0.1 to 75.6

pg/L, respectively, while marine concentrations of DEHP and DBP ranged 0.1 to

2306.8 ig/L and 1.0 to 1028.1 ig/L, respectively (Fatoki and Noma, 2002).

Additionally, contamination of raw drinking water by di-ethyl phthalate (DEP) has

been reported (USEPA, 2001). Some phthalate contamination in the

environment may be traced back to WWTP discharges. Fromme et al. (2002)

surveyed 39 German wastewater treatment plants and showed the

concentrations of DEHP and DBP were highly variable in the effluents and

ranged from 1.7 to 182 [ig/L and 0.2 to 10.4 pg/L, respectively. Furthermore, the

concentrations of phthalate esters in the WWTP sludge ranged from 27.9 to 154

mg/kg dry weight for DEHP and 0.2 to 1.7 mg/kg dry weight for DBP (Fromme et

al., 2002).

Phthalate esters, including the commonly used di-ethyl phthalate (DEP),

DBP, BBP, and di-isobutyl phthalate (DIBP), are capable of inducing an

estrogenic response in reporter assays, but their potency was one millionth that

of 17 p-estradiol (Jobling et al., 1998; Harris et al., 1997). Legler et al. (2002)

identified the hormonal activity of BBP using an estrogen receptor-chemically

activated luciferase reporter gene (ER-CALUX) construct; however, the








Table 2-5. Concentrations of nonylphenols in food items in Germany
Concentrationa
Food Item (ig/kg)
Peanut butter 5.2
Marmalade 7.3
Butter 14.4
Tomatoes 18.5
Apples 19.4
Breast milk 0.3
Infant formula 1.6 2.1
a from Guenther et al. (2002).

responses of other phthalates, e.g., DEP and DBP, were weaker. Researchers

continue to investigate the hormonal activity of phthalate compounds; in fact,

there is still a debate surrounding the hormonal activity of the most commonly

used phthalate, DEHP (Metcalfe et al., 2001).

In humans, the main pathway for the conjugation of phthalates prior to

excretion is via glucuronidation (Albro et al., 1982). Evidence for the reduced

hormonal activity of phthalate conjugates comes from rodent assays (Foster et

al., 2000). Although the conjugated phthalates are excreted at concentrations in

the microgram per liter range, they may be rapidly deconjugated in the presence

of the glucuronidase enzymes (Blount et al., 2000). These glucuronidase

enzymes are present in high concentrations in domestic wastewaters.

Alkylphenols

Alkylphenol polyethoxylates (APEs) are one class of non-ionic surfactants,

with numerous industrial and domestic uses (Talmage, 1994). These hydrophilic

APEs are rapidly degraded during biological treatment, e.g. in wastewater

treatment plant (WWTP), to hydrophobic and recalcitrant alkylphenols (AP). Due

to the high degree of ethoxylation, APEs are not estrogenic; however, activity has








been reported in the degradation products nonylphenol (NP) and octylphenol

(OP) (Routledge and Sumpter, 1996a). Nonylphenols are ubiquitous

contaminants of commercially available food items (Guenther et al., 2002) (Table

2-5).

The affinity of the hydrophobic APs to sediment increases with organic

content (Lye et al., 1999). As a result, the reported half-life of sediment-

associated alkylphenols is roughly 60 years (Shang et al., 1999). In the outfall of

WWTPs, reported concentrations of APs in the sediments range from 2 to 9,050

ng/g dry weight in freshwater environments (Lye et al., 1999) and from 1370 to

1630 ng/g in marine environments (Shang et al., 1999).

Humans excrete AP compounds as glucuronide conjugates (Muller et al.,

1998), and these conjugates are then subjected to biological degradation

processes. Current wastewater treatment technologies are not effective for the

complete removal of APs (Sheahan et al., 2002b); hence, pg/L levels are

discharged to receiving waters and induce hormonal responses in fish (Sheahan

et al., 2002a). In aquatic environments, one of the biomarkers for exposure to

hormonally active compounds is the presence of vitellogenin (Vtg), a fish egg

yolk protein. Jobling and Sumpter (1993) reported a 20- to 90-fold increase in

Vtg production in Rainbow trout (Oncorhynchus mykiss) exposed to various

concentrations of alkylphenols (1 to 100 jpM) in a laboratory study. Alkylphenols

bioaccumulate (Sheahan et al., 2002b); in fact, 10-30 ng NP/g of liver (wet

weight) was reported in male flounder (Platichthys flesus) (Lye et al., 1999).

Although WWTP effluent concentrations of the hydrophobic NP and OP have








been reported in the mid ng/L range (Snyder et al., 1999), higher concentrations

may be found in the wasted sludge due to partitioning (Ejlertsson et al., 1999; La

Guardia et al., 2001). While octylphenol and nonylphenol are weakly estrogenic

(Legler et al., 2002), their conjugated forms are not capable of inducing estrogen-

like responses (Moffat et al., 2001).

Natural and Synthetic Estrogens

Research indicates that estrogens, both natural and synthetic, represent

the predominant fraction of organic wastewater contaminants and concurrently

induce the highest hormone activity (Metcalfe et al., 2001; Snyder et al., 2001;

Rodgers-Gray et al., 2000; Desbrow et al., 1998). The vertebrate endocrine

system produces chemical messages, called hormones, which regulate body

functions, e.g., reproduction, growth, and homeostasis (Figure 2-2). Estrogens

are the hormones produced by the ovaries and they are responsible for the

development and regulation of female secondary sexual characteristics. The

endogenous estrogens, 17 p-estradiol (E2) and estrone (E1), together with their

degradation product estriol (E3) are rapidly conjugated and excreted from the

body. This made their use in hormone therapies ineffective and led to the

development of synthetic hormones (Bolt, 1979). Although the synthetic

hormones are rapidly absorbed in the bloodstream, they are slowly metabolized

and are; therefore, better suited for drug therapies (Guengerich, 1990). The

most commonly prescribed synthetic hormones are 17a-ethynylestradiol (EE2)

and mestranol, which are both utilized in the production of birth control pills

(BCP) and as inhibitors of ovulation (Ranney, 1977). Although E1 and E3 are the








Table 2-6. Rates for the urinary excretion of natural estrogens from men and
women
Women

Estrogen Pre- Pre- Post- Menc
Menopausala menopausalb menopausala (n=2)
(n=114) (n=25) (n=146)
17p-Estradiol (E2) 3.5 1.1-2.8 0.7 1.5
( g/day)
Estrone (Ei) 7.0 2.6-7.8 1.4 3.9
(gg/day)
Estriol (E3) 8.7 4.7-5.6 1.6 1.5
(lg/day)
Key et al., 1996, reported as geometric mean, "Adlercreutz et al., 1994, reported as range, and
CFotsis and Adlercreutz, 1987 reported as mean. Standard deviations were not reported.

main metabolites of E2, there is also a group of minor metabolites with

inconsequential hormonal activity.

The metabolic pathways for natural and synthetic estrogens have been

extensively reviewed (Bolt, 1979; Guengerich, 1990). Generally, natural and

synthetic hormones in the human body are metabolized to inactive glucuronide or

sulfonide conjugates before excretion (Bolt, 1979). The age distribution and;

hence, the reproductive conditions of women in a population determines the total

concentration of excreted estrogens. Table 2-6 summarizes reported excretion

rates of natural estrogens from both men and women. Keys et al. (1996) showed

that pre-menopausal women excreted 3.5 pg/day of E2, 7.0 ig/day El, and 8.7

pg/day E3. In a separate study with pre-menopausal women, similar estrogen

concentrations in urine were reported by Adlercreutz et al. (1994). The slight

variation in estrogen excretion reported by Key et al. (1996) and Adlercreutz et

al. (1994) was probably due to the menstrual phase during urine collection.








Adlercreutz et al. (1994) collected urine samples during the mid-follicular phase

(3-11 days after the onset of the last menstruation), while Key et al. (1996)

analyzed urine collected throughout the entire menstrual cycle.

Overall, pregnant women excrete the highest concentrations of natural

hormones at 600 jig/day, 259 pg/day, and 6000 pg/day for Ei, E2, and E3,

respectively (Fotsis et al., 1980). The estrogen concentrations reported for post-

menopausal women were 0.7 jig/day for E2, 1.4 pg/day for El, and 1.6 pg/day for

E3 (Key et al., 1996) and were comparable to male estrogen excretion rates.

Male (n=2) excretion was reported for El, E2, and Esat 3.9 pg/day, 1.5 ig/day,

and 1.5 ig/day, respectively (Fotsis and Adlercreutz, 1987). Predicting the

excretion rates of synthetic hormones is more difficult and depends on the

number of pre-menopausal females in a population, cultural mores and the brand

of birth control pill used (Johnson et al., 2000). A search of pharmaceutical

information on the Internet showed a range of 30 40 j.g EE2I tablet, with a

typical dosing regime of 21-28 days, followed by 7 days of inactive tablets.

Larsson et al. (1999) estimated EE2 excretion rates at 4 jg/day per female

consuming oral contraceptive pills in Sweden. The excretion of endogenous

hormones is predominantly via the urine, while fecal elimination generally

exhibits a minor secondary role; however, for the excretion of synthetic hormones

the fecal route is primary (Ranney, 1977).

Fecal excretion rates of endogenous estrogens from pre-menopausal

women (n=25) were reported as 0.5 ng/day for E1, 0.4 ng/day for E2, and 0.8

ng/day for E3 (Adlercreutz et al., 1994). Daily excretory rates for feces and urine








have been reported at 100-400 grams and 1-1.3 kg wet volume/person/day,

respectively. These ranges generally apply to men, with excretion rates for

women generally at the lower limit of this range (Polprasert, 1989 as cited in

Bitton, 1994).

Effects of Hormonally Active Compounds on Humans

Over the past 30 years, what began as anecdotal observations of altered

reproductive and sexual development in humans have coalesced into concem for

long-term species survival (McLachlan, 2001). The early onset of middle-age

vaginal carcinomas and deformed uteri in young women have been linked to the

potent synthetic estrogen diethylstilbesterol, widely prescribed to pregnant

women throughout the 1950's and 60's (Colbum et al., 1996). During the first

trimester of pregnancy, human fetuses are highly sensitive to exposures from

hormonally active compounds.

Industrialized nations, including the U.S., Scandinavia, and Japan, have

reported an increased incidence of hypospadias (displacement of the urethral

opening toward the scrotum) and cryptorchidism (failure of the testicles to

descend into the scrotum) in males (Paulozzi, 1999). Some researchers have

questioned this conclusion and instead cite increased reporting and stricter

definitions as factors artificially inflating the data. Widespread trends are difficult

to establish, but adverse sexual effects from exogenous substances have been

confirmed. The feminization of males has been attributed to work place

exposure to formaldehyde (Finkelstein et al., 1988) and therapeutic treatments

with herbal supplements (DiPaola et al., 1998). Gray (1998a) showed that

sexual differentiation in male rats was altered after exposure to hormonally active








HORMONALLY ACTIVE COMPOUNDS


CENTRAL NERVOUS SYSTEM


HYPOTHALAMUS

PITUITARY GLAND


N PANCREAS OVARY
Jl 1t
ggn i" ""


I

]


GLTESTESND
J TESTES PNA GLAND


MUSCLE. NERVOUS LIVER, I CIRCADIAN
[IVERSYSTEMRLE SYSTEM RYTHYMS
LYMPH TISSUE MKDNEYS MUSCLES REPRODUCTIVE ORGANS R
KIDNEYS U C

Figure 2-1. Representation of the vertebrate endocrine system and the possible influences of hormonally active
compounds on various systems and organs. (adapted from Mathews and van Holde, 1996)


L

I


THYMUS


I


-U
THYROID








Table 2-7. Advantages and disadvantages associated with the use of in vivo and
in vitro assays for identifying hormonal activity
Advantages Disadvantages
In Vivo
Metabolic capability High cost

Multiple nuclear receptors Assay duration (weeks to months)
Non-standardized protocols (dosing
Established assays regime, food, endpoint)

Sensitivity to non-hormonal effects
In Vitro
Low cost No metabolic capability
Rapid (hours to days) Predominately measure ER mediated
Rapid (hours to days) effects
Simplified culture techniques Lack of pathways to clear hormones

Minimize endocrine system complexity


compounds and pesticides. The human reproductive system is regulated by a

plethora of chemical messengers in a complex relay of signals that control

gametogenesis, ovulation, fertilization and sexual differentiation (Thomas, 1997).

The vertebrate endocrine system produces chemical messages, called

hormones, which regulate body functions, e.g. reproduction, growth, and

homeostasis (Figure 2-1). Numerous reviews have been published that discuss

the effects of endocrine-disrupting compounds on humans (Sultan et al., 2001;

Degen and Bolt, 2000; Paulozzi, 1999; Neubert, 1997).

Bioassays to Identify Hormonal Activity

Pursuant to congressional mandates, the US environmental protection

agency (USEPA, 1998) developed a framework for a tiered screening program








Table 2-8. In vivo assays for the determination of hormonal activity
ASSAY ENDPOINT
Measure uterine weight of ovariectomised
Rodent rodents
Measure vaginal cornification" of
Rodent ovariectomisedb rodents
Measures androgen sensitive tissue weight of
Hershberger Castrated Rat castrated male rodents

Fish Gonadosomatic index, Vtgc induction

Turtles Vtg induction
vaginal lesions, ovaries removed surgically, CVtg, vitellogenin (a female egg yolk protein)

that integrated in vivo and in vitro bioassays for the identification and

quantification of hormonally active agents in the environment (Gray, 1998b).

Researchers continue to investigate novel approaches for identifying hormonal

activity, and, although these new methods increase the knowledge base of

hormonal effects, more work is needed in establish the foundation of adverse

hormonal effects. Primarily, increased validation of the more widely utilized

assays, e.g., the yeast estrogen screen (Routledge and Sumpter, 1996b) and

rodent assays (Ashby, 2000) and inter-laboratory comparison of these

established assays (Ashby, 2003) are needed. In vitro and in vivo assays each

have their own advantages and disadvantages; therefore, a battery of assays

utilizing both types of assays has the greatest value (Table 2-7).

Endogenous estrogen ligands bind with estrogen receptors at the cellular

level in a well-defined cascade of cellular events (Okamura and Nakahara,

1999). The endogenous ligand enters the cell via active transport mechanisms

or diffusion. Once inside the cell, the ligand enters the nucleus and binds with

the estrogen receptor displacing the heat shock proteins (e.g., Hsp90) associated








Table 2-9. Threshold dose for the induction of hormonal effects following
exposure of fish to natural and synthetic estrogens
Conc.
Species Hormone on. Response Source
(ng/L)

Oncorhynchus
mykiss Ei 25-50 Vtg induction Routledge et al.,
(Rainbow E2 1-10 Vtg induction 1998
trout)
Oncorhynchus
mykiss EE2 1.5 Vtg induction Larsson et al., 1999
(Rainbow
trout)
Rutilus
Rutilus 1-10 Routledge et al.,
rutilus E2 Vtg induction 1998
(Roach)

Oryzias EE2 0.0b Alteration in
latipes E2 8b reproductive Metcalfe et al., 2001
(Medaka) E3 750b characteristics

Danio Vtg induction; a n B l.
rerio EE2 5-10 Erratic
(Zebrafish) spawning
Ictalurus
punctatus E2 2.7 Vtg induction Monteverdi et al.,
(Channel E2 2.7 Vtg induction 1999
(Channel 1999
catfish)
Platichthys
flesus EE2 10 Vtg induction Allen etal., 1999
(Flounder)
aAbbreviations: El, estrone; E2, 17p-estradiol; E3, estriol; EE2, ethynyl estradiol; Vtg, vitellogenin
lowest observed effect concentration


with the receptor. These proteins maintain the structural conformation of the

estrogen receptor (Fliss et al., 2000). The receptor-ligand complex then binds to

a specific ligand-binding domain on the nuclear DNA (Massaad et al., 1998),

which codes for the transcription of messenger RNA (mRNA). In the cellular








machinery, the genomic message on the mRNA is translated into protein. This

suite of events is initiated in response to the estrogenic ligand. Estrogen

receptors are part of a "superfamily" of nuclear receptors and include a large

number of orphan receptors, with no recognized ligands (McLachlan, 2001).

There are two forms of the estrogen receptor, estrogen receptor a (ERa) and

estrogen receptor p (ERp). Although, the tissue distribution of ERa and ERp

differ based on sex (male or female) and organ type, they display similar

sensitivities to the endogenous estrogen 17 p-estradiol (Couse et al., 1997).

While the majority of hormonally active substances exert their influence via

ligand-dependent activation of the estrogen receptor, some substances do not

act via receptor interactions (EI-Tanani and Green, 1997).

In vivo assays for the determination of hormonal activity

Traditionally, the potential for hormonal activity was assessed with in vivo

assays (Table 2-8). Typical endpoints measured are increased uterine weight,

altered sex ratios, skewed gonado-somatic index, and induction of vitellogenin

(Vtg) (an egg yolk protein in vertebrates). In vivo assays using various

crustacean species (Andersen et al., 1999; Fingerman et al., 1998), including

Daphnia magna (Shurin and Dodson, 1997; Baldwin et al., 1997; Baldwin et al.,

1995) have been used to evaluate hormonal activity as decreased steroid

metabolism and developmental abnormalities.

Common in vivo assays use rodents (Prinsen and Gouko, 2001), fish

(Sepulveda et al., 2002; Bowman et al., 2000) and some invertebrates (Gagne et

al., 2001; Blaise et al., 1999), but these assays are expensive, labor-intensive








Table 2-10. In vitro assays for the determination of hormonal activity
ASSAY MODE OF ACTION
measures the ability of a substance to
stimulate proliferation in estrogen sensitive
Cell Proliferation Assays cells
Ex. E-Screen (Soto et al., 1992)
measures the affinity between a substance
Receptor Binding Assays and the estrogen receptor
Ex. hER a or p (Gutendorf and
Westendorf, 2001)
measures the ability of a substance to induce
Reporter Gene Assays the reporter gene
Ex. YES (Routledge and Sumpter, 1996b)
measures the induction of a specific proteins
Cell Line Assays or enzymes by a substance
Ex. Liver cells (Monteverdi et al., 1999)
measures the ability of substance to stimulate
Cell Prolifer Ge cell proliferation and induce reporter gene
Cell Proliferation/Reporter Gene transcription
transcription
Ex. ER-CALUX (Legler et al., 1999)


and, in some cases, raise ethical concerns. One of the advantages of in vivo

assays is the cellular machinery for the metabolism and/or conjugation

ofhormonally active compounds (HAC). The degradation of the parent HAC may

produce a metabolite with no hormonal activity (Harris et al., 1997). Table 2-9

summarizes literature reports for the threshold dose of natural or synthetic

estrogens required for the induction of a hormonal response in various fish

species.

In vitro assays for the determination of hormonal activity

Generally, the premise on which in vitro assays are based is the defined

mechanism of action between hormonally active ligands and nuclear receptors,

usually the estrogen receptor. Due to the lack of metabolic pathways in the in

vitro assays, the hormonal activity may be over-predicted. In most cases, in vitro









Table 2-11. Relative sensitivity of in vitro assays to 17 p-estradiol (E2)
MDLa EC5o Source
Assay (ng/L) (ng/L)ource
RCBAb .03 Coldham et al., 1997

RCBA .02 Klein et al., 1994

YES 3 Routledge and Sumpter, 1996b

YES 2.7 27 Murk et al., 2002

YES 19 Beresford et al., 2000

YES 27 Tanaka et al., 2001

YES 13 Elsby et al., 2001

YES 5.4 Vinggaard et al., 2000

YES 22.8 Layton et al., 2002

YES 2.7 27 Legler et al., 2002


E-SCREEN

E-SCREEN

E-SCREEN
Cell line/reporter
(MVLN)
Cell line/reporter
(HGELN)
Cell line/reporter
(ER-CALUX)
Competitive
binding (ER a)
Competitive
binding (ER p)
Competitive
binding (ER (?))


1.4

2.7

1.7

1.4

10.9

0.1 1.6

900

17700


Gutendorf and Westendorf, 2001

Soto et al., 1992

Behnisch et al., 2001

Gutendorf and Westendorf, 2001

Gutendorf and Westendorf, 2001

Legler et al., 2002

Gutendorf and Westendorf, 2001

Gutendorf and Westendorf, 2001


272 1360 Murk et al., 2002


Abbreviations: MDL, minimum detection limit; "RCBA, recombinant cell bioassay; 'estrogen
receptor form not indicated








assays are constructed with recombinant molecules or cells from either

mammalian or fish tissues (Zacharewski, 1997; Diel et al., 1999). Most

frequently, yeast cells (Saccharomyces cerevisiae) are used as carriers for the

hormone receptors. Early in vitro assays incorporated an estrogen receptor (ER)

and reported only ER-mediated hormonal activity. Recently, in vitro assays have

been designed with other nuclear receptors, including androgen receptors (AR)

and progesterone receptors (PR) (Nishikawa et al., 1999). Table 2-10

summarizes the general types of in vitro assays currently available.

The hallmark of an in vitro assay is the sensitivity of the assay to the

endogenous estrogen E2. Table 2-11 summarizes the sensitivities of various in

vitro assays. Advantages afforded by in vitro assays include the rapid

identification of hormonal effects, relatively lower cost, and the ability to screen

numerous samples simultaneously. Typical endpoints for in vitro assays include

cell proliferation, enzyme expression (e.g. p-galactosidase), and protein

synthesis. Some of the limitations inherent in in vitro assays are the absence of

metabolic pathways, the over-estimation of binding in a single receptor systems

(Jobling et al., 2002), and reliance on estrogen mediated effects, while ignoring

other receptors (Diel et al., 1999). Despite these concerns in vitro

assayscontinue to be widely used for the identification of hormonal activity and

for the quantification of the contributions from individual chemicals to overall

activity (Degen and Bolt, 2000; Gutendorf and Westendorf, 2001).

The most widely used reporter gene assay incorporates the human

estrogen receptor (hER) into the genome of the yeast Saccharomyces cerevisiae








(Rehmann et al., 1999; Coldham et al., 1997; Gaido et al., 1997; Routledge and

Sumpter, 1996b). Yeast cells are stably transfected with the human ER (hER)

and expression plasmids for a reporter gene, usually lac-Z (codes for the enzyme

p- galactosidase). When estrogenic ligands enter the cell, they bind with the ER

to form a ligand-ER complex. This complex then interacts with the estrogen

responsive element (ERE) on the plasmid and initiates transcription of the

reporter gene. The reporter gene product is quantified by the addition of a

suitable substrate. Due to their easy quantification by spectrophotometers,

chromogenic substrates are typically used, e.g., chlorophenol red

galactopyranoside (CPRG) (Routledge and Sumpter, 1996b) or ortho-nitrophenol

galactopyranoside (ONPG) (Lascombe et al., 2000; Klein et al., 1994). These

estrogen receptor/reporter assays are rapid, reproducible, and have

demonstrated a high degree of sensitivity to hormonally active compounds.

The YES (Routledge and Sumpter, 1996b) has been widely used to

identify hormonal activity in pure compounds (Beresford et al., 2000), wastewater

treatment plant influents and effluents (Holbrook et al., 2002), flue gases

(Muthumbi et al., 2002) and recycled materials (Vinggaard et al., 2000). This

assay has also been adapted to include the androgen receptor and thus quantify

androgenic effects (Beresford et al., 2000).

Other yeast-based assays have been developed, and some are gaining

wide acceptance due to their use of multiple nuclear receptors. A novel ligand-

receptor binding assay was constructed in a yeast (Y190) two-hybrid assay with

expression plasmids (pGBT9 and pGAD424) to determine the interaction








between selected hormone receptors (ER, AR, PR, MR, TR) and co-activators

(TIF2, SRC1, TIF1, RIP140) in the presence of HACs (Nishikawa et al., 1999).

This assay is highly sensitive to phytoestrogens and nonylphenol (Nakano et al.,

2002) and environmentally relevant concentrations of HACs (Kawagoshi et al.,

2003). The role of the co-activator is poorly understood, but, after binding of the

ligand to the nuclear receptor, it appears to influence processes that initiate gene

transcription (Nishikawa et al., 1999). The TIF 2 co-activator had the greatest

influence on initiation (Kawagoshi et al., 2002).

Exploitation of the estrogen sensitivity of breast cancer cells led to the

development of assays that quantify cell proliferation in the presence of

estrogens or estrogen-like substances. The E-Screen (MCF-7 breast cancer

cells) assay (Soto et al., 1992) provides highly reproducible results when assay

protocols are strictly adhered to and the cell line source is consistent. Payne et

al. (2000) evaluated three MCF-7 cell lines (BUS, UCL, and SOP) for their

sensitivity to E2 and showed EC5o values of 3.6, 3.1, and 2.6 ng/L, respectively.

However, the proliferative effect (growth in excess of control cells) of E2 varied

markedly among the cell lines at 8.9 with BUS, 0.98 with UCL, and 1.45 with

SOP (Payne et al., 2000). The functionality of some estrogen receptors in breast

cancer cell lines may be low, with up to 50 % non-functional (Balmelli-Gallacchi

et al., 1999). In cell proliferation assays, direct counts occur with

hemocytometers or automated Coulter counters.

The addition of reporter genes to cell lines has increased their sensitivity

and allowed for the elucidation of multiple mechanisms of action in a single test








system. Luciferase reporter gene constructs were designed in human breast

cancer cell lines. Addition of the lux reporter gene to MCF-7 and HeLa cells

produced the MVLN and HGELN systems (Balaguer et al., 1999; Gutendorf and

Westendorf, 2001). This adaptation allowed for the identification of substances

whose mode of action was via cell proliferation or ER activation. Katori et al.

(2002) demonstrated that di-butyl phthalate induced cell proliferation, but not

gene transcription.

Additionally, the ER-Chemical Activated Luciferase gene eXpression (ER-

CALUX) assay was developed in T47D human breast cancer cells by

transfection with reporter genes (pEREtata-Luc) (Legler et al., 1999). The steep

E2 dose-response curve with the ER-CALUX assay was nearly 20 times greater

than the reporter gene response in the YES, indicating the higher sensitivity of

the ER-CALUX assay (Legler et al., 2002) (Table 2-11). Other advantages of the

ER-CALUX assay are the small sample volume requirements, which were

roughly 1/10t and 1/100t* the volume required by the YES and ER binding

assays, respectively (Murk et al., 2002).

Cell line assays are not limited to mammalian cells. Monteverdi and

Giulio (1999) combined primary liver hepatocytes from Channel catfish (Ictalurus

punctatus) with an enzyme-linked immunosorbant assay (ELISA) to measure the

induction of Vtg. Petit et al. (1999) developed a test system in yeast that

expressed the Rainbow trout estrogen receptor (rtER).

Competitive binding assays measure the displacement of E2 from the

estrogen receptor. The ERa has a higher sensitivity than the ERI for E2








(Gutendorf and Westendorf, 2001). Generally, these competitive binding assays

have a lower sensitivity than other in vitro assays and are not suited as screening

assays for hormonal activity (Table 2-11). In competitive binding assays the

response to agonistic and antagonistic xenobiotics are measured simultaneously,

which accounts for the lower sensitivity (Murk et al., 2002).

Some novel approaches for detecting hormonally active compounds have

been developed. A biosensor has been constructed, which incorporates the

human estrogen receptor a (ERa) into a lipid bilayer with direct contact to a gold

electrode for quantification (Granek and Rishpon, 2002).

A test battery to determine hormonal activity

An optimum test battery to evaluate aquatic toxicity includes algal,

invertebrate, and bacterial components (Rojickova-Padrtova et al., 1998).

Similarly, a successful investigation of hormonal activity should use a suite of

assays to include a cell proliferation assay, a yeast reporter assay, and a

competitive binding assay (Fang et al., 2000; Coldham et al., 1997). When in

vitro assays are combined in an array, then multiple mechanisms may be studied

simultaneously including the effects of metabolism and/or transport (Baker, 2001;

Vinggaard et al., 1999). The sensitivity of in vitro assays to E2 stands as the

hallmark by which assays for hormonal activity are judged. Gutendorf and

Westendorf (2001) demonstrated an E2 sensitivity that increased in the order of

estrogen receptor binding assays (ERa and ERp) < reporter gene assays < cell

proliferation assays (Table 2-11). The ER-CALUX assay has demonstrated the








overall highest sensitivity to E2, with an EC50 of 1.6 ng/L compared to 27 ng/L

with the YES and 1,360 ng/L with ER binding assays (Murk et al., 2002).

Characterizing Hormonal Activity in MSW Landfill Leachates and Other
Environmental Samples

Until recently, little was known about the hormonal activity of MSW landfill

leachates, despite the threat they posed to the environment (Ejlertsson et al.,

1999). Modern MSW landfills are engineered with barriers to restrict the mobility

of the liquid fraction (leachate) of waste and collection systems; however, this

has not always been the case. In the past, the disposal of waste was largely

unregulated allowing for direct release of toxic substances to ground and surface

waters. Some of these older unregulated landfills continue to release leachates

of unknown strength and chemical composition.

In MSW landfills, biological and chemical processes produce leachates

with high concentrations of organic contaminants (Yasuhara et al., 1999).

Shiraishi et al. (1999) demonstrated the presence of compounds with known

hormonal activity in the leachates of Japanese landfills. Behnisch et al. (2001),

using the E-Screen assay (MCF-7 cells), confirmed the hormonal activity of

waste leachates in a Japanese landfill and tracked reductions after treatment

processes. They reported an estradiol equivalency (EE) of the untreated

leachate at 4.8 ng EE/L and after biological and activated carbon treatment at 2.8

ng EE/L. This was equivalent to a 58 % reduction in hormonal activity (Behnisch

et al., 2001). The composition of waste materials in Japanese domestic landfills

differs significantly from U.S. MSW landfills. Due to space constraints and other

considerations, domestic wastes in Japan are first incinerated to reduce volume








and reactivity, then the ash material is landfilled. The composition of the waste in

the landfill studied by Behnish et al. (2001) was primarily inorganic, with about 70

% incinerator ash. Characterizing some of the organic compounds in the

leachate revealed the presence of known hormonally active compounds,

specifically bisphenol A, nonylphenol, and estradiol at 0.13, 2.8, and 0.005 pg/L,

respectively.

Kawagoshi et al. (2002) used a yeast two-hybrid reporter assay

(Nishikawa et al., 1999) to demonstrate the hormonal activity of waste leachate-

contaminated groundwater at an E2 activity equivalent to 27.2 ng/L. Leachates

that were collected from sites for the disposal of solid municipal wastes and

dredged soils were not hormonally active. In a continuing investigation, the

efficiency of the extraction procedures for the recovery of hormonal activity were

evaluated (Kawagoshi et al., 2003). Their extraction procedures used C-18 SPE

columns and were highly efficient for the recovery of hormonal activity. Together

with the high recovery of activity following elution with polar solvents (acetone),

these results implicated non-polar hydrophobic compounds as causative agents

for the activity (Kawagoshi et al., 2003). Based on the comparison of bioassay

results with one raw leachate and acetone extracts of the same leachate,

Kawagoshi et al. (2003) suggested that anti-estrogenic compounds in the

leachate interfered with the recovery of hormonal activity.

Considering the potential for release of MSW leachate from landfills, the

fate of hormonally active compounds in soils is a concem. The mobility of

hormonally active compounds (HACs) (E2, EE2, nonylphenol, octylphenol, and








Table 2-12. The hormonal activity of selected metal species
MCF-7 luciferasee E-Screen
Sub e reporter assay Assay
Substance (ECo) (EC0)
(EC5o) (EC5o)
(nM) (nM)
E2 0.03 0.14
Bis(tri-n- 1.84 0.55
butyltin)
Antimony 16.4 14.8
chloride
Chromium 34.5 33.3
chloride
Lithium 47.1 49.7
chloride
Cadmium 108 176
chloride
Barium
Barium 743 458
chloride
aChoe et al. (2003).


bisphenol A) in soil (93 to 94 % sand) was investigated in lysimeter experiments,

with bioassays (Dizer et al., 2002). Although the leachates (water extracts)

produced by the lysimeters displayed a low hormonal activity, no attempt was

made to quantify the concentrations of hormonally active compounds in the

leachates. Hence, reduced hormonal activity may have resulted from the

adsorption of the HACs to soil particles (Dizer et al., 2002).

While organic substances are the most widely recognized hormonally active

compounds, some heavy metals are also hormonally active (Stoica et al., 2000).

Although the concentrations of heavy metals in MSW landfill leachates are low,

there is a potential for increased leaching of heavy metals as landfills age

(Bozkurt et al., 2000; Flyhammer, 1997). Stoica et al. (2000) reported cadmium

(as CdCl2) activated the estrogen receptor at low concentrations, but at high








Table 2-13. Reported phthalate concentrations in landfill leachates


Compound Concentration Reference
(ng/L)


Bis(ethylhexyl) 1350a Yasuhara et al.,
phthalate 1999
S61400 Yasuhara et al.,
Bisphenol A 61400 1999
Yasuhara et al.,
Dimethyl phthalate 300a araet1999 al.,
Yasuhara et al.,
Dibutyl phthalate 1800 1999haraet
Yasuhara et al.,
Diethyl phthalate 1600 1999raal
Welander and
Diethyl phthalate 479 Henrysson, 1998
Bis(ethylhexyl) 10,850 Welander and
phthalate Henrysson, 1998
Butyl-benzene 2300 Welander and
sulfonamide Henrysson, 1998
"median concentration

concentrations it blocked the binding of estradiol to the receptor. Choe et al.

(2003) used the E-Screen assay and a cell proliferation/ luciferase reporter gene

assay in MCF-7 cells to determine hormonal activity in twenty species of eight

metals (Table 2-12). The metal species were ranked for their hormonal activity in

the cell proliferation/reporter assay as Bis(tri-n-butyltin) > cadmium chloride >

antimony chloride > barium chloride = chromium chloride. Although the

sensitivity of the E-Screen assay was lower, the ranking was similar with Bis (tri-

n-butyltin) > cadmium chloride > antimony chloride > lithium chloride > barium

chloride (Choe et al., 2003).

As previously discussed, phthalates have demonstrated hormonal activity

in a variety of test systems. Their presence in MSW landfill leachates can be

attributed to the composition of the waste material and the increasing use of


~ --`--








excess packaging in consumer goods. Table 2-13 summarizes the

concentrations of various phthalates identified in MSW landfill leachates.

Hormonal activity has been identified in a variety of other environmental samples,

which include tree debarking mill effluents (Mellanen et al., 1996), wastewater

treatment plant effluents (Shang et al., 1999), industrialized rivers (Lye et al.,

1999) and surface waters (Witters et al., 2001). The sources of hormonal activity

with the widest distribution are the WWTPs. This is attributed to the

concentrations of natural and synthetic hormones in domestic wastewater (Table

2-14). In the absence of WWTPs, septic systems represent a source of

hormonally active compounds (Rudel et al., 1998). The largest sources of

hormonal activity in wastewater treatment plants (WWTPs) are the natural and

synthetic estrogens. They are excreted as inactive glucuronide and sulfonide

conjugates; however, rapid deconjugation occurs in WWTPs (Ternes et al.,

1999a). Deconjugation occurs in the presence of the enzyme glucuronidase,

which is abundantly produced by Escherichia coli (Ternes et al., 1999b). This

enzyme is responsible for the degradation of both the glucuronide and sulfonide

estrogen conjugates (Belfroid et al., 1999), but sulfonide to a lesser degree

(Huang and Sedlak, 2001). Roughly 25 % of excreted estrogens occur as

sulfonide conjugates, and their degradation is closely associated with

arylsulfatase enzymes. Low concentrations of these enzymes in treatment plants

is responsible for the greater persistence of the sulfonide conjugates (D'Ascenzo

et al., 2003). Regardless of the reason for incomplete deconjugation of excreted

estrogens, the underestimation of estrogen loads on treatment facilities may








Table 2-14. Concentrations (ng/L) of natural and synthetic hormones in
wastewater treatment plants(WWTPs)
17p3- Estrone Estriol 177a-ethynyl
Estradiol (EI) (E3) estradio Source
(E2) (ng/L) (ngL) (EE2)
(ng/L) (ngnL)L
WWTP ND ND ND 263 ND
Inf Sole et al.,
TP ND ND ND ND 2000
Eff
WWTP 11.6 51.8b 80.4b 3.0bti .,
inr" Baronti et al.,
WWTP 1.4b 18.4b 3.0b 0.4b 2000
Eff~
WTP <0.5-20 <0.5-75 2-120 <0.5-6 Johnson etal.,
WWTP 2000
Eff <0.5-7 <0.5-52 <0.5-28 <0.5-2.2 2000
T 50.7 NM NM NM Matsuial.,
InO Matsui et al.,
WWTP 7.1 NM NM NM 2000
Eft
WWTP 11b 44b 73b NM
n NM D'Ascenzo et
WWTP 1 23b N al., 2003
Eff' 1.6 17b 2.3b NM
WWTP Rodgers-Gray
Eff' 7-88 15-220 NM NM eta
(winter) et al., 2000
(winter)
VWVTP Rodgers-Gray
Eft 4-8.8 27-56 NM NM etal.2
(summer)
WWTP ND 9b M b Ternes et al.,
EffaN 1999a
WWTP 6b 3b NM 9b Temes et al.,
EfF 1999a
WWTP Belfroid et al.,
Eff. 0.9 4.5 NM <0.3 1999
WWTP Desbrow et
Efr' 2.7-48 1.4-76 NM 0.2-7 al.,1998
WWTP Huang and
E0.2-4.1 NM NM 0.2-2.4 Sedlak,2001
WWTP Snyder et al.,
ElP 1.9-14.6 NM NM <0.05-3.0 ydetal.
Abbreviations: Wastewater treatment plant, WWTP; Inf, influent; Eff, effluent; ND, not detected;
NM, not measured. indicates median value. Samples collected from: d Germany, CCanada,
dUnited Kingdom, USA, 'Netherlands, 'Italy, hSpain, Japan








result (Johnson et al., 2000).

To preface any discussion of reported concentrations of hormonally active

compounds in the environment, it is important to consider the detection limits of

the analytical methods. Often the analytical methods for the identification and

quantification of low-level organic contaminants are ineffective and grossly under

predict environmental burdens (Castillo and Barcelo, 1999). For example, while

some laboratories have reported detection limits for E2, El, E3, and EE2 as low as

0.1-0.6 ng/L for surface waters and 0.1-2.4 ng/L for wastewaters (Belfroid et al.,

1999), others have reported detection limits up to three orders of magnitude

higher for E2(250 ng/L), E1(100 ng/L), E3(50 ng/L) and EE2 (500 ng/L) in WWTP

effluents (Sole et al., 2000).

Alterations in the sexual and developmental characteristics of aquatic

species have been reported worldwide and attributed to the release of

hormonally active micro-organic contaminants (Sheahan et al., 2002b; Desbrow

et al., 1998; Jobling et al., 1993). Although researchers have reported a range of

estrogens in WWTP effluents, concentrations are generally at the low ng/L level.

Chiefly, E2 and EE2 as the hormones with the greatest reported activity have

been identified in effluents at <0.2 to 88 and <0.2 to 9 ng/L, respectively (Table 2-

14). The metabolite El has been identified in effluents at concentrations of 1.4 to

220 ng/L (Desbrow et al., 1998). While E2 is widely recognized as the strongest

estrogen, its metabolites are also potent, with E1 about 1.5 times less active

(Jurgens et al., 2002) and E3, the weakest metabolite, also inducing hormonal

affects albeit orders of magnitude less (Metcalfe et al., 2001). The synthetic








estrogen, EE2, while found at lower concentrations in wastewater (Johnson,

2000) has an activity comparable to that of E2 (Larsson et al., 1999). The lower

activity of E1 is offset by its extensive presence, and gives rise to concerns about

the equivalent E2 activity.

Rodgers-Gray et al. (2000) investigated the influence of seasonal changes

on activated sludge biology and its subsequent effect on the removal of E2 and

E1 from an activated sludge WWTP in England. During the cooler winter months,

concentrations of E2 and E1 ranged from 7 to 88 ng/L and from 15 to 220 ng/L,

respectively. In contrast, there were lower concentrations reported in the

summer months with E2 concentrations ranging from 4 to 8.8 ng/L and E1

concentrations ranging from 27 to 56 nglL. Overall, the concentrations of E1

exceeded those of E2, indicating a greater recalcitrance of E1 to biological

treatment (D' Ascenzo et al., 2003). Few studies have looked at E3, and its

effects may be under estimated.

A limited number of studies have evaluated the concentrations of

endogenous and synthetic estrogens in influents and corresponding effluents of

WWTPs. Despite reported removal rates of up to 90 %, concentrations of

estrogens remain at threshold levels for inducing hormonal effects. Six activated

sludge treatment facilities in Italy (Cobis, Fregene, Ostia, Roma Sud, Roma Est,

and Roma Nord) have been extensively studied by three research teams over a

two-year period. Baronti et al. (2000) reported mean influent concentrations for

the six facilities at 51.8, 11.6, 80.4, and 3.0 ng/L for E1, E2, E3, and EE2,

respectively. Following biological treatment these concentrations were reduced








Table 2-15. Reported concentrations (ng/L) of natural and synthetic estrogens in
surface waters
17p- Estrone Estriol 17a-ethynyl
estradiol (E) (E) estradiol Source
(E2) (ng/L) (ng/L) (EE2)
(ng/L) (nglL)
Surface water <0.5 ND NM <0.5 Tees et al.
1999a
Belfroid et al.,
Surface water <0.3-5.5 NM NM <0.3-4.3 Be et al
1999
Surface water 3.6-5.2 NM NM <0.05-1.4 Snyder et al.,
1999
Surface water <0.05-0.8 NM NM <0.05-0.07 Hang and Sedlak,
2001


to 18.4, 1.4, 3.0, and 0.4 ng/L for E1, E2, E3, and EE2, respectively. These results

are consistent to those reported by Johnson et al. (2000) and D' Ascenzo et al.

(2003), although the later did not measure EE2. Notably, E1 displayed the

greatest range of concentrations (Baronti et al., 2000; Johnson et al., 2000; D'

Ascenzo et al., 2003) and remained the most prevalent estrogen following

treatment. The only other study to evaluate multiple hormone concentrations in

influents and corresponding effluents was conducted with four WWTPs in Spain

(Sole et al., 2000). While analyses were conducted for E1, E2, E3, and EE2, the

only hormone detected was E3 and then only in the influents of 2 WWTPs at

approximately 262 ng/L (Sole et al., 2000). These levels were nearly double the

maximum concentrations detected in the Italian WWTPs (Baronti et al., 2000).

Comparatively, low ambient levels of natural and synthetic hormones have

been detected in surface water samples, however, the potential for extensive

contamination of surface waters exists from animal manure's (Casey et al., 2003)

and WWTP discharges (Kolpin et al., 2002). In surface waters, E2 concentrations





53


ranged from <0.05 to 5.5 ng/L, while those for EE2 were <0.05 to 4.3 ng/L (Table

2-15). As a point of reference, altered sexual characteristics are induced in male

rainbow trout (Oncorhynchus mykiss) at -1.5 ng/L EE2 (Larsson et al., 1999).

Additionally, threshold doses of 1 to 10 ng/L and 25 to 50 ng/L for E2and Ei,

respectively, have been reported for the induction of hormonal effects in O.

mykiss (Metcalfe et al., 2001).













CHAPTER 3
TOXICITY OF LEACHATES FROM FLORIDA MUNICIPAL SOLID WASTE
(MSW) LANDFILLS USING A BATTERY OF TESTS APPROACH

Introduction

The State of Florida currently generates more than 25 million tons of

municipal solid waste (MSW) a year. Fifty-six percent of this waste is disposed in

engineered Class I landfills (FDEP, 2000). The state has sixty-one Class I

landfills (permitted to accept only MSW) that are lined and contain systems for

the collection and transport of waste leachates, that are then subsequently

subjected to biological treatment. Researchers have extensively characterized

the chemical and physical characteristics (Townsend et al., 1996; Booth et al.,

1996; Gettinby et al., 1996) and biological toxicity (Plotkin and Ram, 1984;

Ferrari et al., 1999; Ernst et al., 1994) of waste leachates world-wide. These

waste leachates are a complex mixture of both inorganic (e.g., heavy metals,

ammonia) and organic substances (e.g. pesticides and chlorinated

hydrocarbons). It has been suggested that exposure to MSW landfill leachates

may pose as great a cancer risk as does the exposure to industrial waste

leachates (Brown and Donnelly, 1988) due to their mutagenic properties (Beg

and AI-Muzaini, 1998). The genotoxic potential of MSW landfill leachates was

shown to be higher than that for industrial wastewater, groundwater or drinking

water samples (Helma et al., 1996).








When evaluating the toxicity of complex effluents, the use of a battery-of-

tests approach allows for multiple mechanisms of action to be evaluated

simultaneously (Deanovic et al., 1999; Rutherford et al., 2000). A battery-of-

tests approach with algal, crustacean, and bacterial assays was used to

successfully characterize landfill leachate toxicity (Rojickova-Padrtova et al.,

1998; Clement et al., 1996). No characterization of MSW landfill leachate toxicity

is complete until toxicological assays are combined with analytical procedures for

chemical characterization (Lambolez et al., 1994).

The unique sub-tropical climate in Florida with generally abundant rainfall

and warm temperatures reduces the chemical strength of MSW landfill leachates

(Reinhart and Grosh, 1998). Although researchers have characterized the

composition and site-specific parameters for MSW landfills throughout Florida,

the biological effects of these leachates have not been assessed (Reinhart and

Grosh, 1998). The toxicity of MSW leachates from separate landfills, while

related, may differ due to specific characteristics of the wastes, e.g. pH,

temperature, ammonia levels, presence of recalcitrant organic substances, and

microbiological activity. Little information is currently available conceding the

toxicity of MSW landfill leachates in Florida and such a database of information

could be useful in evaluating leachate treatment options and reuse possibilities

(Ward et al., 2000). The research community now recognizes the importance of

using a tandem approach, with both biological and chemical analyses, when

analyzing environmental samples to achieve a better understanding of the

possible causes of toxic effects.












Site 5

Site 1 Site4

Site 2 and 3
S\ Site 6










Figure 3-1. Locations of the MSW landfills for the collection of leachates in
Florida

The objectives of this research were to 1.). characterize the toxicity of

Florida MSW landfill leachates, 2.). evaluate the use of a battery of algal,

invertebrate and bacterial toxicity assays with Florida leachates, 3.). characterize

the chemical composition of the MSW leachates, and 4.). determine relationships

between selected chemical components and leachate toxicity. Six landfill sites

in north and north-central Florida were sampled monthly over a six-month

sampling period. The sites sampled represented a variety of factors, including

rural and urban areas, some industrial activity, leachate recycle, enhanced

biological treatment in-situ, and those sites currently accepting waste and capped

sites.








Table 3-1. Amount of MSW generated and landfilled at six landfill sites in Florida
Landfill MSW Collected Amoun Waste Landfill
Site (Tons/year)a Land d Type

1 235,662 65 Leachate recycle

2 226,477 69 Rural

3 NAb NA Rural/capped

4 317,694 68 Semi-urban

5 10,197 83 Regional
Urban/enhanced
6 1,890,112 73 biological treatment
'The data presented represents information collected in 1998. "Data were not available (NA) for
site 3 a capped landfill site, no longer permitted to accept MSW.

Materials and Methods

Leachate Collection

Municipal solid waste (MSW) leachates were collected from six sites

located in five landfills in central Florida, USA. The sites were designated as 1

through 6. Site 2, an operating landfill unit, and site 3, a capped landfill unit, are

located at the same landfill (Figure 3-1). Capped landfill units no longer accept

waste materials and are surrounded by a high-density polyethylene (HDPE) liner

to prevent the infiltration of water and the potential for subsequent escape of

leachates. Table 3-1 summarizes the rates of municipal waste disposal at the

sites under study (FDEP, 2000). MSW landfill leachates were collected from

leachate collection wells using a Teflon baler. One sample was collected at each

site, and then apportioned to separate containers for chemical analysis and

toxicity assays. Leachates for chemical analysis were collected in polyethylene

or glass containers and preserved according to U.S. Environmental Protection








Agency (USEPA, 1993b). Samples for toxicity analysis were collected in plastic

cubitainers, transported to the lab on ice and immediately stored at 4C until

sample analysis, within 1 to 2 days.

Chemical and Physical Characterization of Leachates

The chemical/physical characterization of the MSW landfill leachates

began with field measurements that included pH and temperature (Orion, Model

290A), conductivity (HANNA Instruments, Model H19033), dissolved oxygen

(DO) (YSI Inc., Model 55/12 FT), and oxidation/reduction potential (ORP)

(Accumet Co., Model 20). In the laboratory, the MSW landfill leachates were

analyzed for a number of standard chemical and physical parameters, which

included alkalinity, biochemical oxygen demand (BOD), chemical oxygen

demand (COD), ammonia and sulfides, according to methods described by

USEPA (1993b) and APHA (1999). Leachates for metal analysis were digested

and analyzed by Inductively Coupled Plasma (ICP) (Thermo Jarrell Ash, Model

Enviro 36). For ion analysis, a Dionex ion chromatograph (Dionex, Model DX-

500) was used. Total ammonia (NH4* + NHa) and un-ionized ammonia (NHs)

were analyzed by a selective ion probe (Accumet, Model 15).

Maintenance of Test Organisms

Pseudokirchneriella subcapitata

Pseudokirchneriella subcapitata (previously known as Selenastrum

capricomutum) is a freshwater unicellular green algae routinely utilized in both








Table 3-2. Components of the preliminary algal assay procedure (PAAP)
medium

Growth Assay
medium medium
MACRO SALTS
Magnesium sulfate X X
(MgSO4*7H20)
Magnesium chloride X X
(MgCI2*6H20)
Calcium chloride X X
(CaCI2 2H20)
Sodium bicarbonate X X
(NaHCOs)
Sodium nitrate x
(NaNO3)
Potassium phosphate X X
(KH2PO4)
Disodium(Ethylene-
dinitrilo)tetraacetate X
(EDTA)
TRACE METAL SOLUTION
Zinc chloride X X
(ZnCI2)
Cobalt chloride X X
(CoCI2* 6H20)
Sodium molybdate X X
(Na2MoO4 2H20)
Cupric chloride X X
(CuCI2* 2H20)
Boric acid
(H3BO3)
Manganese chloride X X
(MnCI2)
Ferric chloride
(FeCI3*6H20)


FDEP (FDEP, 1997) and USEPA protocols (USEPA, 1978; USEPA, 1994a). The

algae cultures were started from an original algae seed graciously provided by

Hydrosphere Research. Subsequent cultures of algae were maintained in the

laboratories at the University of Florida. The algae growth medium was prepared








according to FDEP (1997) and is referred to as the preliminary algal assay

procedure (PAAP) medium. The components of the PAAP are listed in Table 3-

2. The PAAP was prepared by combining 1 ml of each of the macrosalts with 1

ml of the trace metal solution in a 1-L volumetric flask. The flask was filled with

DDI water and thoroughly mixed. The pH of the PAAP medium was adjusted to

7.5 0.1 with either 0.1 NaOH or 0.1 N HCI. The PAAP medium was then filter-

sterilized (0.45 pm membrane filter) and stored under refrigeration.

The algal cells were grown under a specially designed light unit

(constructed by Martin Dolley). The light unit consisted of a wooden platform (3

feet by 4 feet) supported with legs (4 feet), but with open sides. Three

fluorescent light fixtures (3 1/2 feet long) were suspended from the wooden

platform on adjustable chains. The light intensity inside the unit was regulated by

adjustment of the chain length. Black plastic sheeting surrounded the light unit,

and the interior of the sheeting was lined with aluminum foil to maximize the light

reflection and minimize temperature fluctuations.

When propagating the algal cells, cultures were grown in sterile 1- or 2-L

glass erlenmeyer flasks. The flasks were filled to the three-quarter mark with

sterile PAAP medium and then approximately 100 ml of an algae culture (3 to 5

days old) was added. A 10-ml glass pipette was placed in the flask, which was

then wrapped with parafilm to seal the top of the flask. The flask was swirled

vigorously to mix, placed under the light unit, and then the culture was aerated by

attaching a small hose in series with filters (0.2 im, Acrodisc) to the glass

pipette. A small aquaculture pump supplied air to the algal culture and








continuous gentle mixing of the algae medium was maintained. The algae cells

were cultured for up to one week at 25oC under constant illumination (400 40 ft-

c). After 3-5 days in the light unit, the algae cells were harvested for toxicity

assays, and after 1 week the cells were transferred to start new algae cultures or

were recovered for use as aquatic invertebrate food. The 1-week old algae cells

were transferred to wide-mouth glass containers, covered loosely and placed in

the refrigerator. After settling for approximately one week, the algae cells were

recovered by siphoning off the overlying spent medium. The recovered algae

cells were then resuspended in a minimum volume of DDI. Past experience has

shown that the washing of the algal cells was not required. Algae cultures were

periodically checked for uniformity by transferring small volumes of cells to glass

slides and visually inspecting cell morphology under a phase contrast

microscope. All settled cultures were combined and a final algal cell density was

determined with a hemacytometer. The algae cell density was maintained at 3.5

X 107 cells/ml with DDI before use as aquatic invertebrate food.

Ceriodaphnia dubia and Daphnia pulex

The aquatic invertebrates, Ceriodaphnia dubia and Daphnia pulex are

both members of the family Daphnidae (commonly referred to as daphnids) and

have similar species distributions and life cycles (USEPA, 1993a). Traditionally,

D. pulex (or D. magna) were the invertebrates of choice for determining aquatic

toxicity. Over the past 20 years, assays with C. dubia have increased in

popularity. This is directly related to the greater sensitivity of the C. dubia to

aquatic toxicants (Versteeg et al., 1997). Both invertebrate species are able to








reproduce by parthenogenesis, which ensures a continuous supply of identical

offspring.

Preparation of aquatic invertebrate food

The aquatic invertebrate cultures (C. dubia and D. pulex) were fed a

yeast, cereal leaves, and trout chow (YCT) based food. The YCT was prepared

over a one-week period beginning with the digestion of the trout chow pellets.

The trout chow digestion was performed in a bottomless 3-L inverted plastic

container by combining a 0.5-g portion of trout chow with 1-L of distilled water

(DDI). To ensure adequate mixing, an aquaculture pump was attached to a

glass pipette secured in the inverted cap of the plastic container. After 7 days,

the digestion was complete and the container was covered and placed in the

refrigerator to settle for at least 1 hour. No water was added during the digestion,

despite evaporative losses; however, water was added if needed in the final step

of food preparation to reach the desired solids content. The supernatant from the

trout chow digestion was filtered through a fine mesh, e.g. nylon hose, and then

reserved. Simultaneously, on day 6 of the trout chow digestion a 5-gram portion

of cereal leaves (Sigma) was combined in a blender with 1-L of DDI and mixed

on high speed for 5 minutes. The cereal leaf mixture was covered and reserved

in the refrigerator overnight to settle. Finally, on day 7 of the digestion, a 5-gram

portion of dry yeast (Fleischmann" or equivalent) was combined with 1-L of DDI

water and mixed well on a magnetic stir plate. Equal volumes of the yeast

solution, cereal leaf supematant, and trout chow filtered supernatant were

combined and mixed thoroughly. The YCT food was apportioned into 50-ml








plastic bottles, labeled, and stored in the freezer (-400C) until needed. YCT food

was stored in the refrigerator and unused portions were discarded after 1 week.

The total solids (TS) content of the YCT was maintained between 1.7 and 1.9 g

solids/L by the addition of DDI as needed. The C. dubia and D. pulex cultures

were also fed P. subcapitata algae cells (3.5 X 107 cells/mi), as previously

described.

Maintenance of aquatic invertebrate cultures

Starter cultures of C. dubia and D. pulex were graciously donated by

Hydrosphere Research (Gainesville, FL). The aquatic invertebrates were

cultured in dedicated glassware, which was maintained separately and

thoroughly washed and rinsed between each usage. Daphnids were cultured in

reconstituted moderately hard water (MHW), which was composed of NaHCO3,,

96 mg; CaSO4* 2H20, 60 mg; MgSO4, 60 mg; and KCI, 4 mg per liter of DDI

water. The MHW had the following specifications: pH, 7.4 -7.8; hardness, 80 -

100 mg/L as CaC03; and alkalinity; 60 -70 mg/L (USEPA, 1993a).

Aquatic invertebrate cultures were maintained by adding neonates (< 24-

hour old) of C. dubia or D. pulex to 1-L glass beakers containing MHW. The

daphnid beakers were kept in an environmental chamber (Percival, model E-

30BX) at 20 + 20C and with a light regime of 16 hours of light and 8 hours of

dark. The daphnid cultures were fed 7 ml of YCT and 7 ml of algae cells per liter

of invertebrate culture. Invertebrate cultures were culled daily to remove

neonates and ensure a population with a uniform age distribution. Following the

removal of neonates for toxicity assays, any remaining neonates were used to









Prepare leachate dilutions with
PAAP minus EDTA






Transfer 50 ml of leachate, or its dilution, to
triplicate 125-mi erienmeyer flasks






Spike each flask with 1 ml
P. subcapitata (500,000 cells/ml)







Place erlenmeyers under growth light for
96 hours, shaking daily






Count number algae cells with
hemacytometer and phase/contrast
microscope


Figure 3-2. Flowchart for the P. subcapitata assay








start new cultures or were discarded. Adult females were retained for neonate

production for a period no longer than two weeks.

Toxicity Assays

Pseudokirchneriella subcapitata

The chronic toxicity of the MSW landfill leachates was evaluated

according to the protocols of the 96-hour P. subcapitata assay (USEPA, 1994a).

The leachates were filtered with glass fiber (Whatman, GFIB) and membrane

filters (0.45 pm). The glass fiber pre-filters were used to minimize clogging of the

membrane filter. The dilution media for the algal assays was a PAAP solution

prepared without the disodium (Ethylenedinitrilo) tetra-acetate (EDTA). EDTA

has been shown to form complexes with heavy metals, which confounds assay

results when metal toxicity is suspected. The PAAP growth medium requires the

addition of EDTA, because its presence is crucial for the uptake of many

micronutrients (USEPA, 1994a).

For each algal assay, five dilutions were prepared in a laminar flow hood

with PAAP minus EDTA at a dilution factor of 0.5 (Figure 3-2). A 50-ml aliquot of

the leachate, or its dilution, was added to triplicate 125-ml sterile erlenmeyer

flasks with styrofoam stoppers. The flasks were then inoculated with a 1-ml

aliquot of algae cells (500,000 cells/ml). The inoculum was prepared by

centrifuging a 4 to 5-ml portion of algae stock (3-5 days old) at 4000 rpm for

fifteen minutes. The supernatant was discarded, and the algae cells were

resuspended in PAAP minus EDTA and mixed by vortexing gently. Using a

hemacytometer and a phase contrast microscope, the cell density was

determined. The volume of resuspended algae cells required to prepare an








algae seed with a density of 500,000 cells/ml was determined by the following

equation:

Number test flasks x Vol. test Solution per flask x 10,000 cells per ml
Cell density (cells per ml) in the stock culture


Based on the number of assay flasks and a volume of 1-ml of seed per

flask, the algae cell inoculum was prepared. As an example, if the assay

required 18 flasks containing 50 ml of leachate per flask and the stock algal

culture density was 106 cells/ml, then the required volume of stock culture to

produce 18 ml of algae inoculum is 9 ml. Combining a 9-mi portion of the algae

stock culture (500,000 cells/ml) with a 9-mi portion of PAAP without EDTA

provided the required 18-ml portion of algae inoculum needed for the assay.

The erlenmeyer flasks containing the leachate, or its dilution, and the algal

seed were placed under the light unit (previously described) at 25C. Constant

illumination (400 40 ft-c) was maintained for 96 hours, and the flasks were

rotated and mixed daily by swirling manually. At the conclusion of each assay,

the algae cell density in each flask was measured by algal cell counts using a

hemacytometer and a phase/contrast microscope. Growth inhibition was

determined by comparing the number of algae cells in the leachate containing

flasks to the number in the control flasks. The leachate concentration that

produced a 50 % inhibition (ICso) of algal growth was determined by graphing the

cell density in each flask versus the leachate concentration.









Prepare leachate dilutions with
MHW






Transfer 10 neonates to assay cup; add 20
ml of sample or its dilution






Expose neonates to leachate
for 48 hours







Observe neonates for
death /immobilization


Figure 3-3. Flowchart for the C. dubia assay

Ceiodaphnia dubia and Daphnia pulex

The acute toxicity of the MSW landfill leachates were evaluated with C.

dubia and D. pulex.in the standard 48-hour acute toxicity protocols (USEPA,

1993a; APHA, 1999). Basically, the assay protocols were identical for the two

species of aquatic invertebrates. Before the leachates were evaluated for

toxicity, they were pre-filtered (Whatman, GF/B) to remove large particles. The C.

dubia and D. pulex assays follow similar conditions. Prior to the start of each








aquatic invertebrate assay, the neonates (< 24hrs) were separated from the adult

daphnids and fed a mixture of 7 ml YCT/liter and 7 ml algae/liter. After feeding,

10 neonates were transferred to each test container (30-ml plastic cups) using a

small wide-mouth plastic pipette to minimize the transfer of culture water (Figure

3-3). The leachate dilutions were prepared with MHW at a 0.5 dilution factor, and

the leachate or its dilution was added at 20-ml volumes to triplicate cups

containing the 10 neonates. Containers filled with MHW were used as the

negative controls. The test containers were placed in a water bath at 20 2 C

for 48 hours, with a loose covering to allow light penetration and prevent settling

of air particles. Neonates were exposed to ambient lighting and were not fed

during the assay.

After 48-hours, the invertebrate test containers were placed on a light

table for the determination of viable organisms. The light table was designed to

sit on the laboratory bench (constructed by Martin Dolley). It was composed of a

wooden box with a plexiglass top and three fluorescent lights. The fluorescent

lights were mounted inside the box and below the plexiglass top, so as to

illuminate the work surface. Test containers were swirled gently and neonates

with the power to swim away from the center of the container were counted as

dead/immobilized. Mortality greater than 10% in the controls negated the assay

results.









Add MOAS to leachate for final
concentration of 2% NaCI





Prepare serial dilutions of leachate in with
Microtox diluent





Aliquot 50 g1 of bacterial reagent into
cuvettes; measure initial bioluminescence



JD-

Add 450 i1 of leachate, or its dilution, to
cuvettes containing bacterial reagent;
mix





Measure final bioluminescence after
15-minute exposure


Figure 3-4. Flowchart for the MicrotoxT assay








MicrotoxT

The MicrotoxT toxicity analyzer is a commercially available toxicity system

that measures toxic effects by changes in bacterial (Vibrio fischen)

bioluminescence (Beckman Instruments, 1982). The assay kit includes a diluent,

the freeze-dried Microtox bacterial-reagent, a reconstitution solution, and an

osmotic adjusting solution. Each leachate was assayed in duplicate and each

assay also included a duplicate control (DDI water) (Figure 3-4). A preliminary

investigation indicated that a 15-minute exposure produced the highest

sensitivity, which agrees with the reports of other researchers (Plotkin and Ram,

1984).

The Microtoxm analyzer combines a pre-cooling well for storage of the

reconstituted bacteria reagent during the assay, an incubator well block to cool

the cuvettes containing the sampless, and a turret containing the photomultiplier,

which quantifies the bacterial light output. Before the assay, the analyzer is

brought to thermal equilibrium and the performance calibrated. Clean glass

cuvettes are added to the incubator well block and the pre-cooling well. The

bacterial reagent is rehydrated with 1 ml of reconstitution solution and transferred

to the cuvette in the pre-cooling well. The incubator well block has a grid pattern,

with columns 1 to 5 and rows A to C, row A is designated for the preparation of

the sample dilutions and rows B and C for sample testing in duplicate. Column 1

is dedicated to assay blanks and contains only the bacterial reagent and the

Microtox diluent.








The leachate dilutions are prepared from right to left in columns 5 to 2,

and well A5 corresponded to the highest leachate concentration. The dilution

series is prepared by first adding 1000 pl of the Microtox diluent to the cuvettes in

row A of columns 5 to 2. Then 1000 pl of the leachate is added to columns 4 and

5 of row A and mixed by repeatedly pipetting 500 pl of the mixture and aspirating.

A 1000-pl aliquot from the cuvette in column 4 was transferred to the cuvette in

column 3, mixed, and then a 1000-1 aliquot from column 3 was transferred to

column 2. After column 2 was mixed and aspirated, a 1000-1d aliquot was

discarded. The cuvettes were allowed to cool for approximately 10 minutes.

Following the addition of the reconstituted Microtox bacterial reagent (50 pl) to

wells B1 through B5 and C1 through C5, the cuvettes were allowed to reach

thermal equilibrium, approximately 15 minutes. The luminescence of the

bacterial reagent in each cuvette was determined by placing the cuvette in the

turret and turning the handle to the read position. The luminescence output from

the bacterial reagent was read on the digital panel meter (DPM) at the front of the

analyzer.

The pre-exposure bacterial luminescence was measured beginning with

the cuvette in row B1, then C1, B2, C2, B3, C3, B4, C4, B5, and C5. Next, a

450-pl aliquot of the sample or its dilution was transferred from the cuvettes in

row A to the duplicate cuvettes in rows B and C and mixed, beginning with

column 5. After 15 minutes, the cuvette was placed in the turret and the

luminescence output from the bacteria exposed to the sample or its dilution was

read on the DPM. The loss of bioluminescence is described by gamma (r), which








is measured as the ratio of light lost to light remaining following leachate

exposure.

Data Analysis

The toxicity assay results were expressed as the concentration of leachate

that produced a 50 % effect in the bioassay. The endpoints of the assays were

different and included inhibition of bioluminescence (Microtox as ECso), lethality

or death (C. dubia as LC5o), and growth inhibition (P. subcapitata as IC50). The

results of the 96-hr P. subcapitata assays were determined by graphical

interpolation. The LC5ofor the C. dubia assay was determined using the USEPA

data analysis software (USEPA, 1994b). Bioassay results were presented with

one standard deviation for the C. dubia and P. subcapitata assays. MicrotoxT

test results were determined by least square regression analysis of the natural

log of the sample concentration vs. the natural log of gamma (ratio of light lost to

light remaining). Results were presented as the concentration of leachate

causing 50% inhibition of bioluminescence (gamma). The Microtox assay was

performed in duplicate; therefore, determinations of standard deviations were not

valid. Data was evaluated by least square regression, student's t-test, or the F-

test (Excel, Microsoft 2000), as appropriate. The ECso, ICso, and LC50 results

were transformed to toxicity units (TU); according to the following;

100
TU (unitless) =00
EC, or IC, or LCs,

When appropriate, data were log transformed, for evaluation of linear

relationships.








Table 3-3. Physical and chemical characteristics of MSW landfill leachates at six
sites in Florida
Site Site Site Site Site Site
Parameter 1 2 3 4 5 6
7.5 7.0 7.2 7.5 7.7 7.6
pH (7.3-7.6) (6.5-7.4) (6.8-7.5) (7.3-7.8) (7.5-7.9) (7.3-7.8)
Temperature 35 26 27 26 23 32
(C) (32-38) (25-29) (26-27) (24-28) (19-28) (31-33)
Conductivity 14.1 6.2 5.2 7.6 8.3 9.6
(mS/cm) (13.2-15.2) (3.1-8.4) (2.6-9.5) (6.5-8.6) (3.2-12.1) (1.0-14.2)
Alkalinity 6213 2407 1494 3238 2503 5500
(mg/L as (6075- (1725- (1000- (3125- (1200- (250-8775)
CaCO3) 6625) 2975) 2050) 3350) 4050)
CBODa 140 22 15 66 42 NMc
(mg/L) (89-204) (14-30) (13-21) (55-77) (13-66)
CODb 1850 636 351 1165 857 1245
(mg/L) (107- (522-827) (242-440) (902-1616) (416-1208) 13960)
Sulfide 0.3 20 19.7 47 2.2 3580
(g/IL) (0.08-0.8) (17-23) (15-24) (42-56) (0.1-15.8) 5
Al 0.22 <0.2 0.4 0.2 3.5 10.8
(mg/L) (0.07-0.4) (0.17-0.47) (0.17-0.32) (0.1-4.5) (4.2-23.7)
Cu <0.07 <0.07 <0.07 <0.07 <0.07 <0.07
(mgIL)
Cd 0.01 <0.015
(mCd <0.015 <0.015 <0.015 <0.015 (0.006-
(mg/L) 0.02)
Pb <0.06 <0.06 <0.06 <0.06 0.06 <0.06
(mrgL) (0.04"0.1)
Zn 0.04 <0.03 <0.03 <0.03 0.05 0.45
(mg/L) (0.020.06) (0.03-0.07) (0.050.63)
As 0.16 <0.09 <0.09 <0.09 0.12 0.14
(mg/L) (0.1-0.25) (0.03-0.2) (0.1-0.2)
Cr 0.05 <005 <0.05 <0.05 0.07 0.18
(mg/L) (0.03-0.08) (0.02-0.1) (0.05-0.39)
Ba <0.05 0.05 0.07 0.09 0.1 0.1
(mg/L) (0.03-0.1) (0.03-0.1) (0.08-0.11) (0.03-0.2) (0.08-0.15)
Fe 7.13 12.5 4.7 7.0 7.5 50.4
(mg/L) (5.9-9.5) (9.3-21.4) (2.2-9.2) (3.7-9.9) (3.9-13.6) (3.2-86.5)
Na 1495 735.7 390 1001.5 1039 1663
(mg/L) (1402-
(m1) ( (464-928) (352-445) (989-1014) (554-1890) (76-2670)
K 555 343 202 440 201 644
(mg/L) (520-589) (234-414) (184-220) (431-444) (93-316) (45-1011)
Results are shown as the mean and (range). Abbreviations: "BOD,biological oxygen demand;
bCOD, chemical oxygen demand; CNM, not measured.








Results and Discussion

Chemical Analysis of MSW Leachates

The physical and chemical characteristics of the MSW landfill leachates

are summarized in Table 3-3. Although the mean pH values were near neutral,

with a range from 7.0 to 7.7, lower pH values were measured in some samples.

The toxicity and bioavailability of some leachate toxicants, especially heavy

metals, are pH dependent (Schubauer-Berigan et al., 1993). The alkalinity of the

leachates was highly variable and ranged from 1,494 mg/L as CaC03 at site 3 to

6,213 mg/L as CaCO3 at site 1. These concentrations are typical for landfill

leachates in the early phases of waste stabilization (Kjeldsen et al., 2002).

Conductivity measures ionized molecules in solution, including both

cationic and anionic species. In the MSW landfill leachates, the range of

conductivity values was wide from a low of 5.2 mS/cm at site 3 to a high of 14.1

mS/cm at site 1. Specifically, the mean sulfide concentrations varied widely from

a low of 0.31 ig/L at site 1 to a high of 3,580 pg/L at site 6. Similar mean sulfide

concentrations of 20 pg/L at site 2 and 19.7 pg/L at site 3 were reported. Since

the leachates from sites 2 and 3 were collected from the same landfill and the

waste compositions were comparable, then similar sulfide levels were expected.

This was also true for most of the other chemical characteristics of the sites 2

and 3 leachates, with a few exceptions. The levels of alkalinity in the site 2

leachates ranged from 1725 to 2975 mg/L as CaCO3, while in the site 3

leachates the range was from 1000 to 2050 mg/L as CaCO3. Alkalinity

concentrations generally decrease with increasing landfill age, as the buffering








capacity is consumed by the production of organic acids. Differences in the

BOD/COD ratios were slight, but suggested a higher degree of organic matter

degradation in the leachates from site 3. The BOD/COD ratio was 0.03 in the

leachates from site 2 and 0.04 in the leachates from site 3. The lower

concentration or inorganic cations in the site 3 leachates was due to the "wash-

out" effect typical of older leachates (Kjeldsen et al., 2002). While organic

components in the leachates are degraded by biological activity, inorganic

constituents decrease over time with increased rates of leachate production

(Cameron and Koch, 1980; Chian and DeWalle, 1976). Some variations in the

chemical composition of MSW landfill leachates are expected based on age,

waste degradation and site-specific factors (Ragle et al., 1995).

Generally, the concentrations of heavy metals in the MSW landfill

leachates were below analytical detection limits (Table 3-3). However, some

leachates contained elevated concentrations of heavy metals. Aluminum was

detected in most of the leachates at least once over the 6-month investigation. In

the leachates from site 6, aluminum was detected in each of the leachate

samples collected with a range from 4.2 to 23.7 mg/L. The heavy metals copper

and cadmium were not detected (detection limits of 0.07 mg/L and 0.02 mg/L,

respectively) in any of the leachates analyzed. While lead concentrations were

less than 0.06 mg/L in the leachates from sites 1, 2, 3, 4, and 6, in the leachates

from site 5 lead concentrations ranged from <0.06 to 0.1 mg/L.

The leachates from site 6 contained zinc concentrations that ranged from

0.05 to 0.63 mg/L. Zinc concentrations in the leachates from site 1 exceeded the









2000 50



40
1500

1300



o20 E
10oo





500
10


N N
0 __ i i4 0
Feb March April May June July


Figure 3-5. Concentrations of total (NH4+NH3) (bars) and un-ionized ammonia
(NH3) (diamonds) in MSW landfill leachates from site 1 over a 6-month
sampling interval. NM indicates sample not measured

detection limit of 0.03 mg/L in five of the six-leachate samples analyzed and

ranged from < 0.03 to 0.06 mg/L. The highest barium levels were measured in

the leachates from sites 2 and 3, although neither exceeded 0.1 mg/L. Similar

concentrations of barium were reported in the leachates from sites 4, 5, and 6,

but barium concentrations were below the detection limit in the leachates from

site 1. In relation to mean arsenic levels in the leachates, 0.16, 0.12, and 0.14

mg/L were identified at sites 1, 5, and 6, respectively. A similar patten was

shown with chromium, but in this case the leachates from site 6 displayed the

highest concentrations with a range from 0.05 to 0.39 mg/L.









500 5



400 4



300 3E
o E-

EC
200 2.5


100 1


0 _0
Feb March April May June July
Figure 3-6. Concentrations of total (NH4++NH3) (bars) and un-ionized ammonia
(NH3) (diamonds) in MSW landfill leachates from site 2 over a 6-month
sampling interval. NM indicates sample not measured

The concentrations of iron in the MSW landfill leachates were high and

site-specific. The lowest mean iron concentrations were reported for the

leachates from site 3 with a mean of 4.7 mg/L, while with the leachates from site

6 mean concentration of 50.4 mg/L was reported. High iron concentrations in the

MSW landfill leachates directly impact the bioavailability of heavy metals via

influences on the formation of insoluble precipitates. Metallo-sulfide precipitates

reduce metal toxicity; however, at high concentrations iron may out-compete the

toxic metals for binding sites on the sulfide molecule (Bozkurt et al., 2000).

Concentrations of other inorganic cations were also high in the MSW landfill

leachates. Sodium levels ranged from a low mean of 390 mg/L at site 3 to high









250



200 4


0












March April May June July

Figure 3-7. Concentrations of total (NH4+NH3) (bars) and un-ionized ammonia
(NH150 ) (diamonds) in MSW landfill leachates from site 3 over a 6-month








sampling interval. NM indicates sample not measured

mean of 1663 mg/L at site 6. Although the potassium concentrations were lower,

a similar pattern was demonstrated. Reported mean concentrations of

potassium in the MSW landfill leachates from sites 3 and 6 were 202 and 644
E














mg/L,March April May June Julyrespectively.

Figure 3-7. Monthly levelntrations of total (NH4+ NH3) (bars) amnd un-ionized ammonia







leachates fluctuated widely, and the concentrations were dependent on site-

specific conditions. Consistently, the highest overall total ammonia
concentration were identified in the MSW landfill leachates collected from site 3 over a 6-month
sampling interval. NM indicates sample not measured








with a range from 970 to 18601663 mg/L (Figure 3-5)at site 6. Although the potassium concentrations were lower,

total ammonia pattern was demonstrate site 2 and rangeported from 100 to 380 mg/L, theions of
potassium in the MSW landfill leachates from sites 3 and 6 were 202 and 644

mgIL, respectively.

The monthly levels of total (NH4 + NH3) ammonia in the MSW landfill

leachates fluctuated widely, and the concentrations were dependent on site-

specific conditions. Consistently, the highest overall total ammonia

concentrations were identified in the MSW landfill leachates collected from site 1,

with a range from 970 to 1860 mg/L (Figure 3-5). Although the concentrations of

total ammonia were lower at site 2 and ranged from 100 to 380 mg/L, the









400 15



300
10 .0
E














(NH3) (diamonds) in MSW landfill leachates from site 4 over a 6-month
sampling interval. NM indicates sample not measured
100




Feb March April May June

Figure 3-8. Concentrations of total (NH4*+NH3) (bars) and un-ionized ammonia
(N H3) (diamonds) in MSW landfill leachates from site 4 over a 6-month
sampling interval. NM indicates sample not measured

variability remained high (Figure 3-6). The lowest total ammonia concentrations

were displayed by the MSW leachates collected from site 3 (the capped landfill

site), with a range from 82 to 220 mg/L (Figure 3-7). In the MSW landfill

leachates collected from sites 4 and 5, total ammonia concentrations ranged

from 211 to 361 mg/L and 119 to 351.6 mg/L, respectively (Figures 3-8 and 3-9).

This contrasts with total ammonia concentrations measured in the leachates from

site 6, which ranged from 33 mg/L in July 2000 to 1957 mg/L in April of 2000

(Figure 3-10).

Ammonia speciation is dependent on both pH and temperature.

Ammonium is the dominant species at pH values < 9.3, while ammonia









400 15




300
10 .a
0

E 200

o 5

100



NM
0-- 0
Feb March April May June July
Figure 3-9. Concentrations of total (NH4+NH3) (bars) and un-ionized ammonia
(NH3) (diamonds) in MSW landfill leachates from site 5 over a 6-month
sampling interval. NM indicates not measured

dominates at pH values > 9.3 (Figures 3-5 to 3-10). As previously discussed, the

pH values were similar at the six landfill sites; however, slightly higher

temperatures were reported in the leachates from sites 1 and 6 (Table 3-3). At

the pH of the leachates, the un-ionized ammonia levels were expected to be low,

and this was true for sites 2, 3, 4, and 5 with reported concentrations of less than

6.5 mg/L. Higher total ammonia concentrations were reported in the leachates

from sites 1 and 6 with mean un-ionized ammonia levels of 23.4 and 44.4 mg/L,

respectively. Basically, the un-ionized ammonia concentrations followed the

same pattern as previously described for the total ammonia concentrations.




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