Citation
The fate of mercury in municipal solid waste landfills and its potential for volatilization

Material Information

Title:
The fate of mercury in municipal solid waste landfills and its potential for volatilization
Creator:
Earle, Celia Denise Adaire, 1967-
Publication Date:
Language:
English
Physical Description:
xiv, 263 leaves : ill. ; 29 cm.

Subjects

Subjects / Keywords:
Dissertations, Academic -- Soil and Water Science -- UF ( lcsh )
Hazardous wastes -- Environmental aspects -- Florida ( lcsh )
Mercury ( lcsh )
Soil and Water Science thesis, Ph.D ( lcsh )
Alachua County ( local )
Landfills ( jstor )
Mercury ( jstor )
Methane ( jstor )
Genre:
bibliography ( marcgt )
theses ( marcgt )
non-fiction ( marcgt )

Notes

Thesis:
Thesis (Ph. D.)--University of Florida, 1997.
Bibliography:
Includes bibliographical references (leaves 242-261).
Additional Physical Form:
Also available online.
General Note:
Typescript.
General Note:
Vita.
Statement of Responsibility:
by Celia D.A Earle.

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Source Institution:
University of Florida
Holding Location:
University of Florida
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The University of Florida George A. Smathers Libraries respect the intellectual property rights of others and do not claim any copyright interest in this item. This item may be protected by copyright but is made available here under a claim of fair use (17 U.S.C. §107) for non-profit research and educational purposes. Users of this work have responsibility for determining copyright status prior to reusing, publishing or reproducing this item for purposes other than what is allowed by fair use or other copyright exemptions. Any reuse of this item in excess of fair use or other copyright exemptions requires permission of the copyright holder. The Smathers Libraries would like to learn more about this item and invite individuals or organizations to contact the RDS coordinator (ufdissertations@uflib.ufl.edu) with any additional information they can provide.
Resource Identifier:
028627104 ( ALEPH )
39526436 ( OCLC )

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Full Text











THE FATE OF MERCURY IN MUNICIPAL SOLID WASTE LANDFILLS
AND ITS POTENTIAL FOR VOLATILIZATION














By

CELIA D.A. EARLE


















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 1997















ACKNOWLEDGMENTS


I would like to thank the Lord Almighty who has been with me constantly throughout this period of time. He has given me the strength that I needed to go on even when I felt like giving up. Without Him, I could not have succeeded.

I would like to thank my parents, Jonathan and Yvonne, my brothers Kevin and Jeremy, and my grandmother, Evelyn Lindo, for their continued prayers and support throughout these trying times.

Special thanks to my wonderful chairman, Dr. R. Dean Rhue, who has steadfastly encouraged and guided me in the way that I should go. He has believed in me and has not limited me to any extent. He has provided monetary support and emotional support throughout this ordeal. He has been like a true father figure with a genuine concern for my well being and success. I truly thank God for him.

Thanks to Dr. David P. Chynoweth, who took me on at a late stage, but has been there for me and supported me. He has provided beneficial input into my research and has also ii









provided the materials and help that I needed to successfully complete the second phase of my research.

Thanks to the rest of my committee, Dr. J.J. Delfino, Dr. K.R. Reddy, Dr. L.Q. Ma, and Dr. T. Townsend, for their constructive criticisms and guidance. Their support also helped me to achieve my goals and for that I am very grateful.

Thanks to William Reve, lab manager in the Soil and Water Science Department, for being my listening ear and for assisting me in working through the glitches that occurred from time to time. Thanks to Paul Lane, lab manager in the Bioprocess lab, for his assistance in getting the second phase of my research underway. Thanks to Dr. John Owens for instructing me in the use of the gas chromatographs. Thanks to Dr. Jose Sifontes, Dr. Paul Gebert and Andreas Sifontes for their assistance with logistics of the reactor sample runs. Thanks to Wendy Stickney for being my lab assistant and constant friend during the second phase of my work, and to Maria Corchuelo who helped with the typing of the references in this document. Thanks to Clayton Clark and Adrienne Cooper for their assistance in preparing this document for submission. Thanks to Joyce Taylor and the rest of the administrative staff in the Soil and Water Science department iii








for aiding me with whatever was required during my time in the department.

Thanks go to my cousins Deirdre Lawson and Andr6 Earle; Special friend Paul Mason; friends Florette Earle, Denise Borel, Michael McCorkle, Ingrid Forbes, Ogechi Okpechi, Michael Distant, Steven Trabue, Gerco Hoogeweg, Hector Castro, Tait Chirenje, Gerald Manley; and to my other relatives and friends in the USA and Jamaica who provided love, prayers and support.































iv













TABLE OF CONTENTS

Page


ACKNOWLEDGMENTS . . . . . . . . . . . ii

LIST OF FIGURES . . . . . . . . .. . .. .... vii

LIST OF TABLES . . . . . . . . . . . . x

ABSRACT . . . . . . . . . . . . . xii

CHAPTER

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

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

Forms and Toxicity of Mercury . . . . . . . . 8
Mercury in the Atmosphere . . . . . . . 12
Mercury in Aquatic Systems . . . . . . . 15
Terrestrial System . . . . . . . . . 28
Analytical Methods . . . . . . . . . 40
Municipal Solid Waste . . . . . . . . 53
Waste Recovery and Resource Treatment . . . . 57 3 MATERIALS AND METHODS . . . . . . . . 77

Phase I-Hg in the Alachua County Landfill . . . 77 Phase II -Landfill Simulated Anaerobic Reactors . . 89 MSW Analysis . . . . . . . . . . 98
Gas Sampling and Analysis . . . . . . . 98
Volatile Fatty Acids . . . . . . . . . 100
pH Analysis . . . . . . . . . . 103
Mercury Analysis . . . . . . . . . . 103
Tygon Tubing Experiment for Mercury . . . . . 108 4 RESULTS . . . . . . . . . . . 111

Hg concentrations in the Alachua County Landfill . . 111

v









Phase II-Anaerobic Reactor Data . . . . . . 125
Phase II-Mercury Data . . . . . . . . 140

5 DISCUSSION . . . . . . . . . . . 149

Phase I Hg in the Alachua County Landfill . . . 149 Phase II . . . . . . . . . . . . 165

6 SUMMARY AND CONCLUSIONS . . . . . . . . 180

APPENDICES

A RAW MERCURY DATA FOR ALACHUA COUNTY LANDFILL RESIDUES
AND PALM BEACH COUNTY COMPOST . . . . . 190

B RAW DATA FOR ANAEROBIC REACTORS . . . . . 211

C RAW MERCURY DATA FOR ANAEROBIC REACTORS. . . ... .218

D ABBREVIATIONS . . . . . . . . . 239

REFERENCES . . . . . . . . . . . . 242

BIOGRAPHICAL SKETCH . . . . . . . . . 262


























vi















LIST OF FIGURES
Figure PaSe




3-1 Alachua County Landfill. ............................. 78

3-2 Sample Handling Procedure ............................82

3-3 Total mercury analysis of solid waste fractions as
described in EPA Method 7471 ....................85

3-4 Schematic of reactor design ..........................90

3-5 Entire setup of reactor system .......................93

3-6 Liquid collection traps behind the reactors.......... 93

3-7 Solid Waste insert into reactor ......................94

3-8 Glass fiber mesh at bottom of solid waste insert.....94 3-9 Top view of the reactor system .......................95

3-10 Close-up view of the activated carbon traps......... 95

3-11 Modification of EPA Method 7471 for determination of
total mercury in sulfur-impregnated carbon.....107 3-12 Total mercury analysis in leachate samples as
described in EPA Method 7470 ...................109

4-1 Run 1 (Control) pH vs. Days... .....................128

4-2 Run 1 (100ng Hg) pH vs. Days ........................128

4-3 Run 1 (2000ng Hg) pH vs. Days .......................128


vii









4-4 Run 1 (Control) Methane vs. Days ................... 129

4-5 Run 1 (100ng Hg) %Methane vs. Days. ..................129

4-6 Run 1 (2000ng Hg) %Methane vs. Days. .................129

4-7 Run 1 (Control) Volatile Fatty Acids vs. Days....... 130 4-8 Run 1 (100ng Hg) Volatile Fatty Acids vs. Days...... 130 4-9 Run 1 (2000ng Hg) Volatile Fatty Acids vs. Days..... 130 4-10 Run 1 (Control) Methane Yield vs. Days ..............131

4-11 Run 1 (100ng Hg) Methane Yield vs. Days .............131

4-12 Run 1 (2000ng Hg) Methane Yield vs. Days ............132

4-13 Run 2 (Control) pH vs. Days .........................133

4-14 Run 2 (100ng Hg) pH vs. Days ........................133

4-15 Run 2 (1000ng Hg) pH vs. Days .......................134

4-16 Run 2 (2000ng Hg) pH vs. Days .......................134

4-17 Run 2 (Control) % Methane vs. Days. ..................135

4-18 Run 2 (100ng Hg) % Methane vs. Days. .................135

4-19 Run 2 (1000ng Hg) % Methane vs. Days. ................136

4-20 Run 2 (2000ng Hg) % Methane vs. Days ................136

4-21 Run 2 (Control) Volatile Fatty Acids vs. Days....... 137 4-22 Run 2 (100ng Hg) Volatile Fatty Acids vs. Days...... 137 4-23 Run 2 (1000ng Hg) Volatile Fatty Acids vs. Days..... 138 4-24 Run 2 (2000ng Hg) Volatile Fatty Acids vs. Days..... 138 4-25 Run 2 (Control) Methane Yield vs. Days ..............139

viii









4-26 Run 2 (100ng Hg) Methane Yield vs. Days............. 139

4-27 Run 2 (10O00ng Hg) Methane Yield vs. Days............ 140

4-28 Run 2 (2000ng Hg) Methane Yield vs. Days............ 140

4-29 Run 1 Hg content in 3 phases(control)............... 143

4-30 Run 1 Hg content in 3 phases (100ng Hg) .............143

4-31 Run 1 Hg content in 3 phases(2000ng Hg) .............143

4-32 Run 2 Hg content in 3 phases(control)............... 144

4-33 Run 2 Hg content in 3 phases(100ng Hg).............. 144

4-34 Run 2 Hg content in 3 phases(1000ng Hg) .............145

4-35 Run 2 Hg content in 3 phases(2000ng Hg) .............145

5-1 Hg conc. ranges(ng/g) in the Alachua County Landfill
& Palm Beach County compost ...................150

5-2 Mean Hg concentrations with depth for LFS1
and LFS2 .......................................156

5-3 Hg spike(ng) in simulated landfill reactors & amount
of Hg volatilized(ng) ..........................179



















ix















LIST OF TABLES

Table ge

2-1 Gas Phase Reactions of Mercury .......................16

2-2 Aqueous Phase Reactions of Mercury ...................17

2-3 An overview of the analytical methods for mercury in
aqueous samples. ................................ 50

2-4 Discard of Hg Containing Products into the MSW
Stream (tons). ...................................58

2-5 The median concentrations in MSW leachate, in
comparison with existing exposure standards.....70 4-1 Alachua County Landfill Total Mercury Concentrations
in three fractions .............................112

4-2 % Dry weight of Total Composite Sample.............. 117

4-3 Alachua County Landfill Mercury Concentrations in
the composite(weighted sum of the three
fractions). .....................................121

4-4 Palm Beach County Mercury Data (1:1 Biosolids to
Yard Waste). ....................................126

4-5 Mercury Recovered ...................................142

5-1 Mean Hg concentrations in composite samples for
LFS 1 and LFS 2 Alachua County Landfill........154 5-2 Mean Hg concentrations in composite samples for L.S. 2
and L.S. 4 Alachua County Landfill............. 159




x









5-3 Mercury means in different fractions at different
sites. ..........................................162

5-4 MSW Sample Volatile Solids Content Average by Round
and Area .......................................164


















































xi















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



THE FATE OF MERCURY IN MUNICIPAL SOLID WASTE LANDFILLS
AND ITS POTENTIAL FOR VOLATILIZATION By

Celia D.A. Earle

December 1997


Chairperson: R. Dean Rhue
Major Department: Soil and Water Science

Mercury is conveyed into landfills primarily via batteries and a variety of other mercury-containing devices and lamps. Mercury has not been the focal point of metals research in landfills, but the first Phase of this study showed that mercury exists in the Alachua County Landfill at concentrations ranging from 32.8 ng Hg/g to greater than 16,800 ng Hg/g. However, over half of the samples had concentrations of 150 ng/g or less. While Hg concentrations in the landfill samples as well as compost samples (1:1 yard


xii








waste to biosolids) from Palm Beach County were generally above background levels for surface soils in Florida, they were two to three orders of magnitude lower than clean-up goals currently used by the Florida Department of Environmental Protection and federal regulations governing land application of sewage sludge as described in 40 CFR Part 503.

Phase II of this study investigated the fate of mercury in simulated landfills and found that the bulk of the Hg added was found in the solid waste. Sixteen out of 18 leachate samples did not have detectable levels of Hg. The percentage of Hg volatilized during anaerobic digestion reanged from over 30% at the lowest Hg level (100 ng Hg per 60 g solid waste) to about 3% at the highest Hg level (2000ng Hg per 60 g solid waste. Evidence was obtained that forms of Hg other than elemental or divalent Hg were volatilized. These other forms were thought to be organic Hg compounds, quite possibly including dimethylmercury.

Sulfur-impregnated, activated charcoal was used to trap Hg volatilized during anaerobic digestion. A modification of EPA Method 7471 successfully recovered >98% of the Hg trapped by the charcoal.

xiii








This study has implicated landfills as a potential source of mercury to the atmosphere. Further research should focus on quantifying the amount of mercury being emitted into the atmosphere from landfills and identifying those stages during anaerobic digestion of MSW that Hg is most likely to volatilize.




































xiv














CHAPTER 1
INTRODUCTION



The overall objective of this project was to determine the fate of mercury (Hg) in a municipal solid waste (MSW) landfill and to establish its potential for impacting atmospheric pollution. Mercury is one of the most toxic heavy metals in existence. It is present in the atmosphere, aquatic systems, terrestrial systems, and biota that exist within these systems. Mercury is unique in its ability to exist as a liquid at room temperature. It also exists in nature in gaseous, liquid, and solid phases.

IStudies have confirmed that anthropogenic activities have increased the input of Hg to the environment. It is estimated that Hg from these activities constitute about half of the Hg entering the environment (Fitzgerald and Clarkson, 1991). Pirrone et al. (1996) have stated that Hg emissions in developed countries increased at a rate of 4.5-5.5% per year until 1989 and have remained somewhat constant since. In developing countries, emissions have steadily climbed at a

1








2

rate of 2.7-4.5% per year. Mercury is present in fossil fuels, and to a lesser degree, coal. The vast amounts of fossil fuels and coal burned annually represent an important source of Hg released into the biosphere (Mitra, 1986).

Over the years, there has been interest in utilizing solid waste (SW) residues for land application. However, little or no data exist in the literature with respect to Hg levels in MSW landfills. In Florida an attempt is being made to set standards for heavy metals in these residues, and there is need for more data to allow for the establishment of threshold levels in SW. Currently, stabilized MSW residues are tested for heavy metals using USEPA 503 regulations that were developed for sewage sludge. Mercury ranks as number three on the EPA's toxic substances list, behind lead and arsenic. Since Hg is one of the most harmful heavy metals in the environment, the fate of Hg in MSW should be known. Successful completion of this research may assist in the establishment of sampling protocols and regulations specifically for Hg contained in SW residues.

Several Hg-containing components enter the SW stream on a daily basis, and eventually make their way to MSW landfills. Fluorescent lights, Hg-vapor lamps, arc lamps,








3

mirror coating, amalgam, and electrical apparatus are important sources of Hg in MSW. Although laws were passed in 1988 prohibiting the disposal of batteries in landfills, batteries no doubt still find their way into them (Bureau of Solid and Hazardous waste, 1995; Tchobanoglous, 1993; Steinwachs, 1990).

Research has established that Hg binds strongly to organic matter. There is a significant proportion of organic material in SW, and it is expected that the majority of the Hg existing in a landfill will be bound to it. Bacteria commonly found in landfills influence the transformations of Hg that is bound to organic matter. Studies have confirmed that many of these bacteria contribute to methylmercury production. Thus, volatilization of dimethylmercury and elemental Hg from landfills could make a significant contribution to atmospheric Hg emissions. It has also been shown that inorganic mercury (Hg2) can volatilize at the higher temperatures found during anaerobic digestion further contributing to atmospheric emissions.

Since there is need to understand the fate of Hg in a landfill, a study was conducted with the following objectives:

(1) Determine Hg concentrations in SW samples collected at









4

various depths over a five-year period from the Alachua County SW Landfill in Florida.

(2) Evaluate the potential for Hg present in landfills

to be volatilized and released to the atmosphere.

These objectives were accomplished in two phases. In Phase I, total Hg levels in residues from the Alachua County landfill were determined. This landfill has been in existence since the early 1970's and has been the primary recipient of MSW from the city of Gainesville since then. Hundreds of residue samples from the Alachua County landfill had been collected and stored during previous studies (Miller et al., 1996; Miller et al., 1994). These samples were analyzed for total Hg to generate a large database for Hg in MSW residues from a landfill. The results for this landfill should be representative of landfills created by many other municipalities around the Southeastern United States.

In Phase II, the fate of Hg occurring in a landfill was studied using laboratory-scale anaerobic digester to which various amounts of Hg were added. In this phase, the distribution of Hg between leachate, solid, and gas phases was determined after a period of anaerobic digestion. Conditions occurring in these small-scale digesters are considered to be









5

representative of those that occur in landfills similar to the one in Alachua County.















CHAPTER 2
LITERATURE REVIEW


The earth's crust has been shown to contain low concentrations of mercury. Its main deposits occur in the form of cinnabar, or mercuric sulfide (HgS), which has an average mercury content of 0.1 to 0.4% (Mitra, 1986). In certain locations in the world such as Spain, Yugoslavia, Italy, Peru, Mexico and the American Continent, mercury is found in impregnated schist or slate and as geodes of liquid mercury. Mercury is also found in nature as the oxide and the selenide and in combination with a number of minerals such as quartz, dolomite, chalcedony, calcite, and pyrite. Mercury is present in fossil fuels, and to a lesser degree, coal. The vast amount of coal burned annually (between 1.4 x 101 and 2.72 x 101 g/yr) represents an important source of mercury released into the biosphere (Mitra, 1986).

A large quantity of mercury in the environment is derived from industrially produced mercury amounting to approximately 10,161 metric tons per year. Industries which are suspected


6









7

as being responsible for the dispersion of mercury are MSW incinerators, medical waste incinerators, fossil fuel burning, chlor-alkali industries, mining and extraction of mercury from cinnabar, amalgamation, electrical equipment, paper pulp, fungicides, instrumentation, crematoria, degassing of latex paint, and cement manufacturing (Pirrone et al., 1996). Exposure to mercury in the home, hospital, or laboratory can occur by the release of mercury from broken thermometers (Stewart and Bettany, 1982; Mitra, 1986).

In the past, an estimated 90,718 kilograms of mercury was used in the USA each year in dentistry for the preparation of dental amalgam restorations (Mitra, 1986). In the early '90s, it was reported that the ingestion of elemental mercury (Hgo) from dental amalgams caused renal damage in laboratory animals (Barkay, 1992). Kunkel et al. (1996) studied the fate of mercury in dental amalgams and determined that soluble mercury was never detected in a treatment plant headworks probably due to the fact that it became insoluble in the sludge floc. Mercury levels and discharge in wastewater from dental clinics in Denmark was investigated by Arenholt-Bindslev and Larsen (1996). Clinics without amalgam separators had a mean value of 270 mg Hg/dentist/day while those equipped with amalgam









8

separators had a mean value of 35 mg Hg/dentist/day. Several hundred grams of mercury/clinic may be discharged with wastewater each year.

Wilhelm et al. (1996) studied biological monitoring of mercury vapor exposure in dental students by scalp hair analysis in comparison to blood and urine. It was concluded that hair may be used as an indicator of internal uptake of mercury provided that the hair was not externally exposed to mercury vapor.

Forms and Toxicity of Mercury

Mercury can be divided into three main categories. These are elemental mercury (Hgo), inorganic mercury (Hg2, and its compounds) and organic mercury (for example methylmercury

(CH3Hg ) phenylmercury (C6HsHg*) ethylmercury (CH3CH2Hg) These affect human health in one way or another. $Metallic mercury and inorganic mercury compounds generally attack the liver and kidney, but they normally do not remain in the body long enough (24 hours for mercuric mercury) to accumulate to serious levels (Hammond, 1971) .1 The excretion of inorganic mercury occurs through the kidneys, liver (as bile), intestinal mucosa, sweat glands, and salivary glands.

Inhalation has been shown to be the main route of









9

exposure for mercury vapor. The lungs are able to absorb the metal with nearly 100 percent efficiency. Mercury vapor is considered hazardous at concentrations greater than 0.05 mg/M3. Once mercury vapor gets into the bloodstream, it is oxidized to Hg2" (Barkay, 1992). Methylmercury and mercury vapor pose the greatest threat to human health./

Methylmercury is the most toxic form of mercury. Microbes methylate inorganic mercury, generally under anaerobic conditions, to produce methylmercury (Mitra, 1986). Fish and shellfish provide the major source of intake by humans. Methylmercury is rapidly absorbed from the gastrointestinal tract of humans and attacks the central nervous system. It is only slowly excreted. The main route of excretion in humans is via the feces, in which the rate is about ten times higher than that in urine.

SPoisoning from methylmercury is characterized by sensory disorders, concentric constriction of the visual fields, impairment of hearing, symptoms from the autonomic nervous system, decreasing physical coordination, loss of memory as well as "mental disturbances" often referred to as Minamata disease. In the most severe case in adults, specific anatomical areas of the brain are damaged. The damage is









10

irreversible because neuronal cells are destroyed (Fitzgerald and Clarkson, 1991). Pregnant women are spared from poisoning because the methylmercury quickly crosses the placental barrier to accumulate in their unborn children instead, thus causing teratogenesis and damage of their central nervous systems (D'itri and D'itri, 1977).

Special precautions must be taken in the laboratory during the handling of mercury, especially the organic mercury forms. The death of the Dartmouth professor, Dr. Karen Wetterhahn, in 1997 highlights the fact that latex gloves do not adequately protect the skin from absorbing dimethylmercury.

The most notable methylmercury case on record occurred in Minamata, Japan during the 1950's. Initially, there were deaths of aquatic and terrestrial organisms. Then the impoverished fishermen and their families who ate large amounts of fish from Minamata Bay developed the strange symptoms mentioned above. It was concluded that these symptoms were caused by people eating fish and shellfish contaminated by methylmercury, the source of which was Hg that was discharged in the wastewater from the Chisso plant. This plant manufactured acetaldehyde and the resulting waste which








11

contained inorganic mercury had been continuously dumped into Minamata Bay since the 1930's. In the sediment, the inorganic mercury was transformed to methylmercury and it eventually accumulated in the fish (Fujiki and Tajima, 1992; D'itri, 1991).

In the biosphere, elemental mercury has a high vapor pressure (Henry's constant, H=0.3) and a low aqueous solubility of 6 x 10-6 g/100 ml water at 250C. Mercuric ion has a low vapor pressure (H < 10-') and a high aqueous solubility of 7.4 g/100 ml cold water (for HgCl,2) It forms covalent bonds, has an affinity for thiol groups, and tends to form strong bonds with inorganic and organic ligands. Organic mercury consists of highly stable C-Hg bonds, is very soluble in both water and hydrocarbons, and undergoes rapid transport through membranes. The vapor pressure of methylmercury (CH3Hg*) is quite low (H< 10 but the vapor pressure of dimethylmercury ((CH3)2Hg)) is high (H < 0.3) (Barkay, 1992).

Mercury has three oxidation states: 0, +1, and +2. Cationic mercury binds tightly to iron (Fe), aluminum (Al), and manganese oxides (Mn02), hydroxides (OH), clay particles and silica. It also has a great affinity to sulfur (HgS), SHgroups, and chlorine (HgC12).









12


Mercury in the Atmosphere

In the atmosphere, mercury may exist in the gaseous phase, the aqueous phase, and the solid particulate phase. Components of the gaseous phase are elemental mercury (Hg0), mercuric chloride (HgCl,2), mercuric hydroxide ((Hg(OH)2), monoalkyl derivatives, such as methylmercuric chloride

(CH3HgCl), and dialkyl derivatives, such as dimethylmercury ((CH3)2Hg). The aqueous phase typically includes HgC12, Hg(OH)2, and sulfites (So3) ; and in the solid particulate phase, there are mercuric oxide (HgO) and mercuric sulfide (HgS) (Seigneur et al., 1994; Schroeder et al. 1991; Nriagu and Davidson, 1986). Concentrations of Hg species in the atmospheric environment are quite varied. Typical gas phase concentrations for Hg range from 2-5 ng/m3 and liquid phase concentrations range from 6-27 x 10- ng/L. In the case of inorganic mercury (Hg2,), the typical gas phase concentration is in the range 0.09-0.19 ng/m3 and the typical liquid phase concentration is in the range 3.5-13.3 ng/L (Seigneur et al., 1994).

It is estimated that 95% of atmospheric Hg exists in the form Hgo and has an atmospheric residence time of 0.7-2.0









13

years (Gustin et al., 1996; Petersen et al., 1995; Nater and Grigal, 1992). Seigneur et al. (1994) have recently shown that global atmospheric budgets of Hg indicate a Hgo half-life on the order of one year.

Slemr and Langer (1992) have found increases in atmospheric Hg concentrations from 1970 to 1990 of about 1.5% per year in the Northern Hemisphere and about 1.2% per year in the Southern Hemisphere. Swain et al. (1992) have determined approximately a 2% increase in Hg deposition rates in Minnesota and Wisconsin.

Atmospheric mercury sources

Natural sources of atmospheric Hg include windblown dust, volcanogenic particles, forest wildfires, and seasalt. Nriagu (1989) stated that volcanoes and fumaroles are responsible for 50% of the Hg that is released naturally, and soil-derived dusts, forest fires, and seasalt sprays account for less than 10% of Hg released from natural sources. Natural degassing rates of Hg on regional or global scales have been estimated to range from 0.02 to 0.03 4g/m2 h (Lindberg, 1986).

Annual global emissions of mercury to the atmosphere have been reported as 0.4 x 106 kg/yr from natural sources and 110 x 106 kg/yr from anthropogenic sources (Schroeder et al, 1992;









14

Nriagu and Davidson, 1986). Major anthropogenic sources include fossil fuel combustion, particularly coal combustion, production of non-ferrous metals, refuse incineration (e.g. MSW incineration), and fuel wood combustion. Pacyna (1987) has implicated coal combustion as the largest single source of atmospheric Hg pollution, accounting for 58% of the anthropogenic input of Hg to the atmosphere. More than 90% of Hg in coal is released as Hgo. Refuse incineration accounts for 33-40% of atmospheric mercuury pollution (Pirrone et al., 1996).

Atmospheric SDecies and Reactions

Many chemical species that are involved in atmospheric Hg chemistry. These will be listed followed by their typical concentrations in brackets: Hg (0.25-0.6 ppt); Br2, HBr

(20ppt); I2, HI (0.3 ppt); HC1 (0.01-2.0 ppb); 03 (0.02-0.40 ppm); OH (0.04-0.4 ppt); NO2 (<=1 ppb-0.5 ppm); SO2 (0.003-0.1 ppm); H2S (0.0-0.01 ppm); N20 (300 ppb); NH3 (0.015-100 ppb); H202 (<1ppb); 02 (2.09 x 10s ppm); and C12 (0.0-8.0 ppb) (Seigneur et al., 1994).

Two types of reactions occur in the atmosphere, namely gas-phase reactions and aqueous phase reactions. Most gasphase reactions involve Hgo reacting with atmospheric species








15

such as 03, 02, C12, H2S and H202.

Table 2-1 shows gas phase reactions of Hg in the atmosphere (Seigneur et al., 1994). Aqueous phase reactions occur in rainwater, fogwater, cloudwater, and other forms of precipitation. Table 2-2 shows aqueous phase reactions of Hg in the atmosphere.

j Mercury in Aquatic Systems

Most aquatic system studies have focused on Hg in lakes (Angstrom et al., 1994; Lange et al., 1993; Winfrey and Rudd, 1990; Steffan et al., 1988). Some authors have researched mercury in oceans and estuarine waters (Fabris et al., 1994; Langston, 1982; Mantoura et al., 1978; Lindberg and Harris, 1974). Mercury in rivers has been investigated by a few authors (Ebinghaus and Wilken, 1993; Bubb et al., 1991; Mantoura et al., 1978).

In freshwater, chlorides (ClI-) and hydroxides (OH-) are the main species existing, but sulfide (S2I-) may also be present. Elemental Hg and dimethylmercury (DMHg) are oxidized in the atmosphere to water-soluble species, such as Hg2 and CH3Hg* which are then deposited onto a wetland (Fitzgerald and Clarkson, 1991). When soluble MMHg enters the aquatic system, it is quickly accumulated by most aquatic biota (Clarkson et








16

Table 2-1. Gas Phase Reactions of Mercury Equilibrium or
Rate Parameter
Reaction (cm3 /molecule s)


(1) Hgo(g)+03(g)-->Hg(II) (g) <8.0 x 10-1

(2) Hg"(g)+C12(g)-->HgC12(g) <4.1 x 10-16

(3) Hg (g)+Br2(g)-->HgBr2(g) <4.1 x 10-1

(4) Hg (g)+I12(g)-->HgI2(g) <2.7 x 10-16

(5) Hg (g)+H202(g) - >Hg(OH)2(g) <4.1 x 10-16

(6) Hgo(g)+2NO,2(g)-->Hg(N2)2(s,g) 3.3 x i0-3

(7) Hg" (g) +HC1 (g) -->products 1.0 x 10-9

(8) 2Hgo)(g) +02 (g)-->2HgO(s,g) <1.0 x 10-23

(9) Hg (g) +S02 (g) -->products <6.0 x 10-1"

(10)Hg0(g)+H2S(g)-->products <6.0 x 10-17

(11)Hg'(g)+N20(g)-->products <2.0 x 10-17

(12)Hg (g)+NH3 (g) -->products <1.0 x 10-7

(13)Hgo (g)+2HI (g) -->HgI2(g)+ H2(g) 2.7 X 10-41

(14) HgC12 (g) -->products slow

(15) Hg (OH) 2 (g) -->products not available



Adapted from Seigneur et al. (1994)









17

Table 2-2. Aqueous Phase Reactions of Mercury Equilibrium or
Reaction Rate Parameter


(1) Hg22 <- >Hgo (aq)+Hg2 2.9 x 10-9M

(2) Hg (aq)+03(aq)-->Hg(II) (aq)+ 4.7 x 107M-Is'
02 (aq)

(3) Hgo (aq)+2HClO(aq)+2H+2e- not available
-- >Hg2 +2C1-+2H20 (1)

(4) Hgo (aq)+Peracetic acid or not available
m-chloro-peroxybenzoic acid(aq)
-->Hg2 or Hg*

(5) Hg (aq)+H202(aq)-->HgO(s)+ 6.0 M-Is-I
Hg2++H20 (1)

(6) Hg,2++03 (aq)-->Hg2* 9.5 x 106M-Is-1

(7) Hg22++H202 (aq) - >Hg2+ <16 M-'s(8) Hg(SO3)22-_ ->Hg 1.0 x 10-4 s-1

(9) HgSO3 (aq) + S03 2-<-->Hg (SO3) 2- 2.5 x 1011 M(10) Hg2+S032-<- >HgS03 (aq) 5.0x1012 M-1

(11) HgSO3 (aq) -->Hgo (aq) +S042- 0.6 s-1

(12)HgCl2(aq) -->Hg2++2Cl- not available

(13)Hg(OH) 2 (aq)-- >Hg (aq) not available

(14)HgC12(s)<-->HgCl2(aq) 0.27 M

(15) HgC12 (aq) <-->Hg2++2Cl 10-14 M2

(16)HgC12(aq)+2C1-<-->HgCI4- 70.8 M-2









18

Table 2-2. Aqueous Phase Reactions of Mercury (continued) Equilibrium or
Reaction Rate Parameter


(17)Hg(OH)2(s)<-->Hg(OH)2(aq) 3.5 x 10-4 M

(18)Hg(OH)2 (aq) <-->Hg2++20H- 10-22 M2 Adapted from Seigneur et al. (1994)








19

al., 1984). ?Mercury bioaccumulates in aquatic organisms by three processes, namely: (1) from the water via respiration (e.g. over the gills), (2) by absorption of water from the body surface, and (3) by ingestion of food (D'Itri, 1991). The uptake of Hg through the aquatic food chain is an extremely important route of bioaccumulation. Mercury tends to bioaccumulate and biomagnify as it moves up a food chain Methylation

Methylation of Hg is a major step in biogeochemical cycling. Sediments are a part of the aquatic system, and are an active site for Hg methylation, being the primary sink for Hg released into the environment. In the sediment, a variety of microbes transform the inorganic Hg into MMHg and DMHg.

Various studies have determined specific microorganisms that are responsible for the methylation of Hg in the environment. Wood et al. (1968) were the first to show that extracts of methanogenic bacteria (strict anaerobes) methylated Hg. )Other studies implicated the bacteria Neurospora spp. (Landner, 1971), and Clostridium cochlearium (Yamada and Tonomura, 1972) in the methylation process, as well as a number of gram-negative and gram positive cocci








20

isolated from river sediment (Hamdy and Noyes, 1975). Vonk and Sijpestein (1973) isolated Pseudomonas and Bacillus from soil, as well as culture strains of Mycobacterium, Escherichia coli, Bacillus megaterium, and fungi. Gilmour et al. (1992) determined that sulfate-reducing bacteria were important in the methylation of Hg. Therefore, it can be stated that the bacteria which methylate Hg are facultative aerobes and anaerobes and sulfate-reducing bacteria.

Mercury that is buried in deeper layers of sediment is not available for methylation unless it is disturbed. Physical occurrences, such as shifting bottom currents, plus small worms, can stir the mercury-rich layers to a depth of two centimeters. Larger freshwater organisms, such as mussels can stir up the sediment to a depth of at least nine centimeters (D'itri and D'itri, 1977). More MMHg is present in water nearest to surface layers of sediment where facultative microbes usually live.

Three methylating coenzymes in biological systems are stated to participate in enzymatic methylation. These are (1) S-adenosylmethionine,(2) N5-methyltetrahydrofolate derivatives, and (3) methylcobalamin (D'itri et al., 1978).








21

Shapiro and Schlenk (1965) and Ridley et al. (1977) discovered that S-adenosylmethionine and N5-methyltetrahydrofolate could not transfer methyl groups to mercuric ions because they were only able to transfer a methyl group as CH3, a carbonium ion. Methylcobalamin is able to transfer a methyl group to an inorganic mercuric ion. It can transfer groups as a carbanion (CH3-) and a methylradical (CH3-) to produce MMHg and DMHg under aerobic and anaerobic conditions (Wood, 1974).

When MMHg is taken up by fish, it moves to the red blood cells and then to fatty tissues where it may be retained for up to two years. Fish and shellfish provide the major source of intake by humans.

In humans, MMHg is rapidly absorbed by the gastrointestinal tract and attacks the central nervous system. When MMHg combines with tissue it is stable and is very slowly degraded or excreted from the body (Mitra, 1986; Barkay, 1992). The main route of excretion in humans is via the feces, in which the rate is about ten times that in the urine (Mitra, 1986). Studies with radioactive MMHg have demonstrated that it is retained in the human body with a half life of about 70 days. Thus, toxic amounts can be accumulated even with low dose rates (Hammond, 1971).









22

Demethylation

Methylmercury is slowly degraded because of the high stability of the carbon-mercury (C-Hg) bond. The low polarity coupled with low affinity of Hg to oxygen decreases the chances for hydrolytic cleavage of the strong bond. The means by which MHg is degraded is by the slow process of protolytic attack on the C-Hg bond. Clarkson et al. (1984) reported that the demethylation of MHg in the water phase is carried out by a variety of microorganisms in two enzymically mediated states: (1) hydrolase enzymes cleave the C-Hg bond releasing the methyl group, and (2) reductase enzymes convert ionic Hg to Hgo which can diffuse into the atmosphere. The equation below shows the transformation: hydrolase reductase CH3-Hg--------> CH4 + Hg2 -----------> Hgo



//The enzyme, organomercurial lyase, which is produced by bacteria in sediments, soils, and waters, can speed up this reaction 108 fold (Barkay, 1992). Thus, other authors have implicated a two step demethylation procedure involving cleaving of the C-Hg bond by organomercurial lyase to produce CH4 and ionic Hg, followed by reduction of the ionic Hg to Hgo









23

by the enzyme, mercuric reductase (Barkay, 1992; Oremland et al., 1991; Nakamura et ial., 1990). The equation below shows this transformation:


organomercurial mercuric
lyase reductase
CH3-Hg----------- > CH4 + Hg2-------------- > Hg"



Furthermore, it was stated that Hg volatilization has been found to be chromosomally encoded in specific bacterial genera, namely Staphylococcus aureus and Bacillus spp. (Nakamura et al., 1990).

Lakes

/ The primary source of Hg in lakes is from atmospheric deposition. Atmospheric deposition to lakes occurs mainly as inorganic Hg, even though there are small amounts of MHg (Winfrey and Rudd, 1990). In aerobic waters, Hg2 will complex with inorganic ligands, such as chlorides and hydroxides, bind with dissolved organic carbon (DOC) or attach to particulate matter. Ionic Hg can also be microbially reduced to form Hg. In anaerobic zones, ionic Hg can be converted to MHg or complex with sulfides and precipitate as mercuric sulfide (HgS). Particle-bound Hg reacts with sulfides and is








24

converted to insoluble HgS which then settles out of the water column onto the sediment. In the presence of hydrogen sulfide (H2S), MHg will form methylmercuric sulfide (CH3Hg)2S which decomposes to HgS and DMHg ((CH3)2Hg) which tends to volatilize (Winfrey and Rudd, 1990; and Iverfeldt and Lindqvist, 1986). / Gilmour et al. (1992) found the highest MHg concentrations near the sediment-water interface and in shallow sediments. Sulfate-reduction rates were also highest at the sediment-water interface, thus confirming that sulfatereducing bacteria are important in the methylation of Hg. jClarkson et al. (1984) stated that inorganic Hg methylation rates by microorganisms are pH-dependent with the greatest amount of MHg being formed at pH levels lower than 6. Higher pHs yield DMHg which is volatilized. Bloom et al. (1991) studied the impact of acidification on the MHg cycle of remote seepage lakes. Methylmercury concentrations were measured in the water of five pristine lakes which had a pH range of 4.6 to 7.2. In general, it was found that MHg in lake water tends to increase as pH decreases. Methylmercury partitioning was weakly related to pH.

Miskimmin et al. (1992) studied the effects of pH on the Hg methylation and demethylation rates in lake water. The









25

results showed that a reduction in pH from 7.0 to 5.0 yielded large increases in net methylation rates at both low and high dissolved organic carbon (DOC) concentrations. Rates of microbial activity, which were represented by rates of respiration, had the least effect on net MHg production rates in the pH range 5.0 to 7.0.

Steffan et al. (1988) studied the effects of acidification on Hg methylation, demethylation, and volatilization in sediments from an acid susceptible lake. Sulfuric acid (H2S04), hydrochloric acid (HC1) and nitric acid (HNO3) were used to acidify the sediment. Methylation was inhibited over 65% when H2SO04 and HC1 were used to reduce the pH from 6.5 to 4.5, There was almost complete inhibition of methylation when HNO3 was used to bring the pH's to 5.5, 4.5, and 3.5. Demethylation was greatest at pH < 4.4, but was not affected by pH's between 4.4 and 8.0. Volatilization was less than 2% of methylation activity and was not significantly affected at the various pH levels.

Winfrey and Rudd (1990) determined that decreased pH stimulates MHg production at the sediment-water interface and probably in the aerobic water column. It was also shown that decreased pH also decreases loss of volatile Hg from lake









26

water and increases Hg binding to particulates in water. These factors enhance the bioavailability of Hg for methylation, therefore methylation rates may be increased. In anoxic subsurface sediments, pH decreases the methylation rates, implying that formation in the water column and at the sediment-water interface may be most significant in acidified lakes.

Xun et al. (1987) reported that in the pH range 4.5-8.5, there was an inverse relationship between Hg methylation and pH. Regnell (1994) did not find significant differences in methyl203Hg in water at pH 5.8 and 6.6 when radiolabeled23HgCl2 was added to water overlying sediment in freshwater systems. The report also confirmed the findings of other authors that anoxic conditions enhanced MHg production in water (Olson and Cooper, 1976).

Mantoura et al. (1978) reported that in freshwater, more than 90% of Hg had been found to be complexed by humic materials.

Xu and Allard (1991) studied the effects of fulvic acid on the speciation and mobility of Hg in aqueous systems. It was determined that at pH levels below the point of zero charge, the humic substances adsorb and increase the uptake of








27

trace metals from the solution phase. In the case of fulvic acid taking part in the adsorption of Hg on an oxide (alumina), the presence of the fulvic acid enhanced the Hg adsorption in the pH range of 2.5 9.5. In general, fulvic acid decreases Hg mobility under both acidic and basic conditions.

Mierle and Ingram (1991) investigated the role of humic substances in the mobilization of Hg from watersheds. Their results indicated that humic matter controls the solubility and watershed export of Hg deposited in precipitation. Minagawa and Takizawa (1980) determined very low levels of inorganic and organic Hg in natural waters by CVAAS after preconcentration on a chelating resin. In analyzing lake and river waters, it was found that 35-60% of the Hg present was in the form of organic compounds or in association with organic matter.

Hakanson (1980) established that higher MMHg levels were found in fish from waters with low productivity, low pH, and high water and sediment load. The reduction of sewage sludge in a lake was shown to increase MMHg content of fish because bioproductivity is decreased and pH lowered. The Hg cycle and fish in the Adirondack lakes was investigated. There was an









28

observed pattern of increasing total Hg and MMHg with decreasing pH. Another correlation found was that lakes with higher dissolved organic carbon (DOC) had higher concentrations of both total Hg and MMHg. There were also positive correlations between fish Hg and MMHg concentrations in the water column.

In the case of wastewater, conventional methods for the removal of Hg2" include sulfide precipitation, ion exchange, alum and iron coagulation and adsorption on activated carbon. Namasivayam and Senthilkumar (1997) investigated the recycling of industrial solid waste for the removal of Hg2' by adsorption process. The report focused on the feasibility of Fe(III)/Cr(III) hydroxide for the adsorption of Hg2+ from aqueous solution.

Terrestrial System

Soil

Mercury enters the soil via the disposal of sewage sludge, rainfall, dry fall-out from the atmosphere and the use of mercury-based pesticides. A more recent addition would be land application of compost to condition the soil or fertilize crops or vegetation grown. Compost is solid waste that has undergone biological decomposition of the biodegradable








29

organic matter under controlled mesophilic and thermophilic temperature conditions, and has been stabilized to a degree which is potentially beneficial to plant growth. This stabilized material is used or sold for use as a soil amendment, artificial top soil, or other similar applications. The stabilized compost can easily and safely be stored, handled and used in an environmentally acceptable manner.

7 Mercury in the soil occurs in several forms. These are

(1) dissolved (free ion or soluble complex); (2) nonspecifically adsorbed (binding mostly due to electrostatic forces); (3) specifically adsorbed (strong binding due to covalent or coordinated forces); (4) chelated (bound to organic substances); and (5) precipitated (as sulfide, carbonate, hydroxide, phosphate, etc.) (Schuster, 1991).

Adsorption of Hg in soils is dependent on the chemical form of Hg, soil pH, amount and chemical nature of inorganic and organic soil colloids, type of exchangeable cations and redox potential (Adriano, 1986).

Coarse gravel has a lower capacity to bind Hg than

finer soil materials, clay > silt > sand in the order of Hg complexation abilities. Farrah and Pickering (1978) investigated the uptake of Hg by three clay minerals and found









30

illite > montmorillonite > kaolinite. Reimers and Krenkel (1974) found similar results in that illite and montmorillonite had faster Hg uptake than kaolinite. Under acidic conditions, ion-exchange is the most important form of sorption to clay minerals, while Hg adsorption is assumed to occur mainly in the form of Hg(OH)2 at higher pH values.

Numerous studies have reported that Hg shows a great affinity for organic matter in soils, peats, and sediments (Newton et al., 1976; Wallace et al., 1982; Schuster, 1991). The binding is strong, but reversible. At low Hg concentrations, soil organic matter is responsible for most of the Hg sorption. At higher concentrations, mineral components take part in the sorption.

Results suggest that inorganic colloids contribute to the adsorption of organomercurials, whereas inorganic Hg compounds tend to bind strongly to soil organic matter. It was also stated that organic components were even more relevant in mercury adsorption at higher Hg concentrations. This is because organic matter has a larger adsorption capacity for Hg than mineral colloids (Schuster, 1991).

Wilken (1992) stated that most of the water soluble Hg species in soil were complexed by organic material with a









31

molecular weight greater than 500 daltons and particle sizes smaller than 1.2 gm. Mercury is bound to humic matter by exchangeable cationic and non-cationic sites (Aggarwal and Desai, 1980). About one-third of the total binding capacity of the soil humus is used for cation-exchange processes, and about two-thirds of the available binding sites serve for metal complexation.

Johansson et al. (1991) studied Hg in Swedish forest soils and waters and determined that the transport and distribution of Hg in forest soils is positively correlated to the transport of organic matter.

The maximum sorption of Hg for several soil types occurs in the pH range 4.75-6.50 (Lodenius, 1990). In acidic soil (pH < 4.5 to 5), organic material is the only sorbent for inorganic Hg. In neutral soils, iron oxides and clay minerals play a part in Hg sorption. With decreasing pH, the mobility of Hg increases in soils that are low in organic matter (Schuster, 1991).

Chlorides occur in all natural soil and water systems and it is regarded as one of the most mobile and persistent complexing agents for heavy metals (Schuster, 1991; Adriano, 1986). Barrow and Cox (1992) investigated the effects of pH









32

and chloride concentrations on the sorption of Hg. They concluded that in the absence of chloride, there were only infinitesimal effects of pH on sorption between 4 and 6. Sorption decreased at higher pH. At low pH, the addition of chloride decreased sorption, but at high pH, chloride additions had little effect on sorption.

Another study dealing with complex formation on Hg (II) adsorption by bentonite also found that chloride ions reduced Hg (II) adsorption, especially at low pH's. Maximum adsorption occurred in the pH range 4.5 to 5.5, regardless of initial Hg concentration.

The leaching of Hg from peat soils, such as those found in freshwater wetlands, decreases with decreasing pH (Lodenius, 1990). Lodenius et al. (1987) studied the sorption and leaching of Hg in peat soil using small additions of labeled Hg in peat lysimeters. Most of the added Hg was bound to the uppper most layer of the peat columns. Additions of chloride and sterilant did not affect the leaching of Hg. When the peat soil was allowed to dry completely, the leaching of Hg was greatest. This may have been the result of Hg that was attached to suspended organic material passing quickly through the cracks that had formed in the soil. Also, greater









33

amounts of artificial rainfall resulted in increasing amounts of Hg being leached out.

Low organic matter acidic soils increase Hg mobility, and removal by leaching is more likely in acidic soils. Fang (1981) studied the sorption of Hg vapor in dry and moist soil columns. He found increased sorption with increasing soil moisture until a maximum was achieved near the maximal water holding capacity. Schnitzer and Kerndorff (1981) determined that humic substances tend to increase the solubility of Hg by the formation of water-soluble complexes.

Comparing peat to sandy soils, the retention of mercury was much stronger and the volatilization smaller in peat than in sandy soils due to the tight bonding of Hg with organic matter (MacLean, 1974). In mineral soils, where Hg adsorption is dominated by iron oxides and clay minerals, there is promotion of evaporation as Hgo. Humic acid enhances the reduction of Hg2" to Hgo and thereby increases volatilization losses to the atmosphere (Rogers, 1977; Lia et al., 1982).

SBacteria and other microorganisms play a major part in Hg volatilization. They are able to reduce Hg2" to Hgo, which is volatile. They are also able to methylate Hg to DMHg, which is volatile, under alkaline conditions.








34

Some bacteria adsorb Hg on the outside of their cell walls where it converts directly into a vapor. Escherichia coli absorb Hg into their systems, then the Hg2" may combine with the cytoplasm and convert into another compound or the volatile Hg. Pseudomonas aeruginosa proteus and at least two other microorganisms are known to convert Hg2 into Hg' (D'itri and D'itri, 1977).

/i"IMercury volatilization tends to increase with increasing temperature. Rogers and MacFarlane (1978) evaluated the volatilization rates of Hg in clay and sand and determined that the volatilization rate of Hg from clay was greater at higher temperatures, but less of the total Hg was volatilized. Increasing Hg concentration resulted in an increased volatilization of Hg from the sand and clay. Autoclaving bother soils drastically decreased the volatilization rates. After inoculation of the sterile soils with non-sterile soils, volatilization rates increased, thus indicating that it was microbially mediated. Another study found Hg to be translocated in soil horizons, and in some cases, lost from surface soil. This was attributed to evaporation to the atmosphere (Dudas and Pawluk, 1976).









35

j Sediments

Accumulation rates for Hg in sediments have been

estimated at between 10 and 50 4g Hg/m2 y (Andersson et al., 1990). Sediments are the primary sink for Hg released into the environment. Sediments consist of an oxidized zone and a reduced zone. In the oxidized zone, particulate Hg is desorbed to release Hg2" and methylated biotically and abiotically to produce MMHg (Zhang and Planas, 1994). In the oxidized zone, Hg2" can be readsorbed to form particulatemercury. The MMHg and particulate-Hg can be deposited into the reduced zone of sediments. In this zone, Hg2" is reduced to Hg* and ultimately Hgo which is then volatilized. Microbes aid in this reduction. In the reduced zone, ionic or Hg2 can react with sulfides that are present to form HgS.

Methylmercury in this zone also reacts with sulfides (S2-) to form DMHg and HgS (Mitra, 1986). Mercury reactions that take place in the oxdized and reduced zones of sediment are given

below;


(a) Oxidized zone reactions:
(desorption)
(1) particulate-Hg2* ----------> Hg2"

(adsorption)









36

(methylation)
(2) Hg2 -------------> CH3Hg*





(b) Reduced zone reactions:

(reduction) (reduction)
(1) Hg2 ---------> Hg ----------> Hg



+S1
(2) Hg2 ---------- > HgS


+S2
(3) CH3Hg' ---------> (CH3)2Hg + HgS



Methanogens, along with sulfate reducers, have been implicated in the methylation of Hg (Adriano, 1986; Wood, 1972). Oremland et al. (1991) investigated the involvement of methanogens and sulfate reducers in oxidative demethylation. Under anaerobic conditions, results with inhibitors showed partial involvement of both sulfate reducers and methanogens. Sulfate reducers dominate estuarine sediments. Products of anaerobic demethylation were mainly carbon dioxide (CO,) and methane (CH4). Methane was the only product resulting from aerobic demethylation in estuarine sediments, suggesting that the procedure for demethylation followed the organomercurial








37

pathway. Final results of this study led to the conclusion that both aerobes and anaerobes demethylate Hg in sediments, but either group may dominate in a specific sediment type.

'An earlier study by Spanger et al. (1973) looked at MMHg and inorganic Hg in lake sediments. Bacterial isolates rapidly degraded MMHg to CH4 and Hgo. Heaton and Laitinen (1974) investigated the electrochemical reduction of MMHg. This occurs in two one-electron steps with the first electron resulting in the formation of a methylmercuric radical on the electrode and the second electron resulting in the reduction of methylmercuric compound to CH4 and Hgo. This also confirms the role played by methanogens in Hg methylation. Langston (1982) showed that humic and fulvic acids provided favorable binding sites that accounted for 4 to 32% of the total Hg measured in surface sediments. Evans et al. (1984) and Revis et al. (1990) found that the majority of extractable Hg was in the humic-fulvic acid and organic-sulfide fractions. Abiological methylation of Hg by humic acid has been confirmed in vitro (Nagase et al., 1984).

Rekolainen et al. (1986) reported on the effect of airborne Hg and peatland drainage on sediment Hg contents in some Finnish forest lakes. The sediment Hg content correlated









38

directly with the organic matter content of the sediment and the pH of the water.

Di-Giulio and Ryan (1987) studied Hg in soils, sediments, and clams from a North Carolina peatland. After selective extractions of peat and sediment samples, it was concluded that the majority of Hg was associated with organic matter associated fractions, particularly humic/fulvic bound and organic-sulfide bound.

An important effect of pH is to mobilize MMHg sorbed on sediments. A reduction in water pH shifts MMHg from the sediments to the water phase, regardless of the type of sediment (Miller and Akagi, 1979).

SDuarte et al. (1991) studied Hg desorption from contaminated sediments. Results showed that the amount of Hg desorbed from the sediment was inversely correlated with pH and ionic strength. The effect of pH on the leaching of Hg from contaminated sediments was greatest at pH values less than 7. Concerning ionic strength, the most Hg leaching occurred at ionic strength values less than 0.4 mol/dm3.

In anaerobic environments, such as sediments, or under mildly reducing conditions, there is a tendency for Hg to be precipitated as the sulfide (HgS). Mercuric sulfide is very








39

insoluble (log Ks = -50.02 at 500C) (IUPAC, 1982), but can be transformed under low Eh and high pH conditions to the soluble form, HgS22*, or to Hgo (Barkay, 1992;, Wilken, 1992; Schuster, 1991). The insolubility of HgS makes it resistant to methylation, but under aerobic conditions, the sulfur in HgS may be oxidized to sulfate, after which the Hg2+ can undergo methylation (Adriano, 1986).

Weber (1993) stated that sulfate-reducing bacteria contribute to MMHg production. A low sulfate concentration was needed for sulfate reducing bacteria to produce MMHg. Kerry et al. (1991) determined that methylation of Hg in sediments of an acid stressed lake was mainly due to the activity of sulfate reducing bacteria. The methylation rate did not correlate with the concentration of sulfate in the system. They speculated that formation of insoluble HgS reduced Hg availability, thus reducing MMHg production.

j Revis et al. (1991) investigated the immobilization of Hg in soil, sediment, sludge, and water by sulfate-reducing bacteria. It was determined that the addition of calcium sulfate (CaSO4) which slowly moves to the sulfur-reducing bacteria, producing H2S and eventually producing HgS,








40

prevented the transformation of Hg to organic methylated forms.

Gilmour et al. (1992) reported that additions of sulfate to anoxic sediments yielded an increased microbial production of MMHg from added inorganic Hg. There were positive correlations between sediment depth profiles of bacterial sulfate reduction and Hg methylation. It was also noted that specific inhibition of sulfate-reducing bacteria blocked MMHg production at all depths in the sediment profile. Sulfatereducing rates and MMHg concentrations were highest near the sediment-water interface and in shallow sediments. These results show that sulfate-reducing bacteria are definite mediators of Hg methylation.

Munthe et al. (1991) investigated the aqueous reduction of Hg2" by sulfite. This involved the formation of an unstable intermediate HgSO3, which underwent decomposition to produce Hg', which in turn was reduced to Hgo. The concentration of sulfite inversely determined the overall rate of the reaction.

Analytical Methods

Atmosphere

Lindberg et al. (1992) investigated atmosphere exchange of Hg in a forest. In this report, Hg was collected in a









41

series of gold traps, wet digested and then analyzed by longpath flameless atomic absorption. The detection limit for Hg in air ranged from 0.01 to 0.5 ng/m3.

Licata et al. (1994) reviewed the testing of Hg in flue gases both in the USA and in Germany. In the USA, the most common method of analysis was CVAAS with SnCl2 in a HC1 solution as the reducing agent. In Germany, the most common method of analysis was also CVAAS but with NaBH4 as the reducing agent.

Prestbo and Bloom (1995) investigated a Hg speciation adsorption (MESA) method for combustion flue gas. The sampling system for gas phase Hg species utilized a series of heated, solid phase adsorbent traps. Mercury (II) and monomethylmercury (MMHg) were the flue gas oxidized species that were adsorbed by a potassium chloride (KC1) impregnated soda lime sorbent. An iodated carbon sorbent was used to collect Hgo after passing through the KCl/soda lime sorbent. CVAFS was used for final determination of Hg from these sorbents.

Chu and Porcella (1995) investigated Hg stack emissions from US electric utility power plants. It was determined that Hg emissions from total electric utilities was on the order of








42

40 tons/yr. It was also determined that Hg emissions were not consistently captured by conventional air pollution control technologies including fabric filters, electrostatic precipitators, and flue gas desulfurization systems. The Hg in the emissions was analyzed by instrument neutron activation analysis (INAA). The MESA method was also used in this study, but neither this method nor the INAA was validated for Hg speciation.

Balogh and Liang (1995) investigated Hg pathways in municipal wastewater treatment plants that used incineration to dispose of the solid material. Iodated carbon traps were used to collect Hg in the incinerator exhaust gas. Subsequently, the traps were digested with acids and final analysis was done by CVAFS.

Keeler et al. (1995) studied particulate Hg in the atmosphere. The analytical technique performed on Hg(p) extracted from glass fiber and other types of filters from glass fiber dual-amalgamation preconcentration and CVAFS detection. Teflon filters have been detected by instrumental neutron activation analysis (INAA).









43

Aquatic Systems

Minagawa and Takizawa (1980) determined very low levels of inorganic and organic Hg in natural waters by CVAAS after preconcentration on a chelating resin. A column of dithiocarbamate-treated resin was used to simultaneously collect inorganic and organic Hg at ng/L concentrations and subsequently quantitatively eluted with slightly acidic thiourea solution. Alkaline SnCl2 solution is used to reduce inorganic Hg to Hg vapor. Mercury vapor was generated from inorganic and organic Hg with a CdCl2-SnC12 solution. The range of determination was 0.2-5,000 ppt for 20-L water samples.

Determination of inorganic and organic Hg compounds by high performance liquid chromatography (HPLC)-inductively coupled plasma (ICP) emission spectrometry with cold vapor generation was investigated by Krull and Bushee (1986). It was not necessary to derivatize samples before analysis by HPLC. The conventional propylene spray chamber of the ICP was replaced by an all glass chamber. Detection limits ranged from 32 to 62 ppb of Hg for four Hg compounds. This represented three to four orders of magnitude enhancement over detection limits without cold vapor generation.








44

Xiankun et al. (1990) investigated the effect of the Huanghe river runoff on the occurrence, transportation, and speciation of Hg in the Huanghe Estuary and the adjacent sea. Mercury content in 150 mL filtered water samples was determined using two stage gold amalgamation coupled with CVAAS.

Lee and Iverfeldt (1991) measured Hg in run-off, lake and rain waters using an oxidative treatment with BrCl prior to reduction by SnCl2. A gold trap was used to preconcentrate the Hg after its volatilization from a quartz glass reduction vessel. Nitrogen gas was used for the purging. A double amalgamation helium dc-plasma atomic emission method was used for analysis of the gold traps.

Baeyens and Belgium (1992) determined total Hg by acidification of the sample to pH 1, pretreatment with a strong oxidant, BrCl, followed by reduction of the BrCl with NH20H-HCI prior to SnCl reduction, air purging of Hgo and collection on a Au-column. Atomic fluorescence was used for analysis of these samples.

Smith (1993) determined natural levels of Hg in water samples by preconcentration onto gold traps followed by electrothermal heating and purging of the traps with argon








45

directly into the torch of the ICPMS. The detection limit was

0.2 ng/L using a 200 mL sample.

Garcia et al. (1994) determined inorganic and MMHg in sea-water by an online preconcentration, solvent extraction and total Hg determination by CVAAS. A detection limit of 16 ng/L of Hg was observed for sample volumes of 25 mL.

Bloom et al (1995) reported on the results of the international aqueous Hg speciation intercomparison exercise. Twenty three labs were involved with 18 utilizing BrCI oxidation, gold trapping and CVAFS for total Hg. Four other labs used CVAAS or wet chemistry and the results were similar. Sixteen labs provided MMHg results, but 15 of them used various combinations of aqueous phase ethylation, GC separation, and CVAFS detection.

Emteborg et al. (1995) used a dithiocarbamate resin for the sampling and determination of Hg species in humic-rich natural waters. Filtration was used to collect the dithiocarbamate resin and transferred to a column which was placed into a closed flow injection system where acidic thiourea solution was used to elute the Hg species. The separation was performed using a gas chromatograph equipped with a non-polar capillary column and detection utilized









46

atomic emission spectrometry at 253.7 nm after excitation in a microwave-induced helium plasma.

Paquette and Helz (1995) reported on the solubility of cinnabar (Red HgS) and implications for Hg speciation in sulfidic waters. In this paper, the Hg analysis was performed by CVAAS on a homemade stannous chloride reduction, gold amalgamation apparatus attached to a Perkin-Elmer Model 2380 Spectrophotometer. A detection limit of 0.3 ppb Hg was noted.

Saouter et al. (1995) developed and field-validated a microcosm to simulate the Hg in a contaminated pond. The water samples were treated and preserved with BrCl and total Hg determination was accomplished using CVAFS.

Specific analytical techniques for organic Hg in aqueous samples have also been investigated. One of the earlier papers by Fitzgerald and Lyons (1973) dealt with trapping the Hg after its reduction while purging with nitrogen gas and concentrating it on a packed column immersed in a liquid nitrogen bath. After the completion of purging, the column was removed from the cold trap, heated, and the gas phase concentration of the eluted Hg was measured on a Coleman Hg analyzer (MAS-50). A detection limit of 0.0017 4g/L was obtained.









47

Schintu et al. (1987) developed a practical isolation technique for MMHg in natural waters. Mercury compounds were extracted quantitatively from six different sources of water with 5 mL of a 50 ppm dithizone-chloroform solution. This method provided a high recovery for both organic as well as inorganic Hg from an aqueous medium, prior to their determination by gold trap CVAAS.

Bloom and Watras (1989) developed a technique to quantify MMHg and DMHg in several rainfall and lake water samples by aqueous phase ethylation to the volatile dialkyl analogs followed by cryogenic gas chromatographic separation. Mercury-specific detection by CVAFS provided a 0.1 pg as Hg limit.

Lansens et al. (1990) determined MMHg in natural waters by headspace (HS) gas chromatography (GC) with microwaveinduced plasma (MIP) detection after preconcentration on a resin containing dithiocarbamate groups. The chelating resin, Sumilate Q-10, showed a high affinity for both organic and inorganic Hg.

Lee and Hultberg (1989) determined MMHg in some Swedish surface waters. Methylmercury was preconcentrated from 10-20 L of water on a sulfhydryl cotton fiber (SCF) adsorbent,









48

packed in a column and eluted with a small volume of 2M hydrochloric acid (HC1). The eluate was extracted with benzene and then analyzed for MMHg with the GC/ECD method.

Rapsomanikis and Craig (1991) investigated the speciation of Hg and MHg compounds in aqueous samples by gas chromatography-atomic absorption spectrometry after ethylation with sodium tetraethylborate. The absolute detection limit for CH3HgCl was 167 pg.

Sarzanini et al. (1992) simultaneously determined methyl, phenyl-, ethyl-, and inorganic Hg by CVAAS with on-line chromatographic separation. Reversed-phase chromatography on an ODS column and elution with an acetonitrile-water-ammonium tetramethylenedithiocarbamate buffered mixture was investigated with or without ammonium tetramethylenedithiocarbamate precomplexation. The detection limits for methyl, ethyl, phenyl, and inorganic Hg were 100 ng/mL, 50 ng/mL, 300 ng/mL, and 8 ng/mL respectively with direct injection (100 4L sample). For the on-line preconcentration procedure, the detection limits were 0.5 ng/mL, 0.09 ng/mL, 0.5 ng/mL, and 0.015 ng/mL respectively in a 100 mL sample.









49

Quevauviller et al. (1992) studied the occurrence of methylated tin and DMHg compounds in a mangrove core from Sepetiba Bay, Brazil. Mercury compounds were determined by derivatization with NaBH4, cryogenic trapping in a chromatographic column and using an electrothermally heated quartz furnace by AA using an EDL source for final detection.

Gilmour and Bloom (1995) reported on a case study of Hg and MHg dynamics in a Hg-contaminated municipal wastewater treatment plant. Methylmercury was determined after distillation by ethylation, isothermal GC and CVAF detection. Filippelli et al. (1992) investigated MMHg determination as volatile MHg hydride by purge and trap gas chromatography in line with fourier transform infrared spectroscopy. The detection limit of this method was 0.15 4g.

Table 2-3 provides a summary of the various analytical methods that the different authors used for the determination of mercury in aqueous samples.

Terrestrial System

Revis et al. (1990) determined MMHg in soil by developing a method for assessing acceptable limits. Water, acid, copper sulfate (CuSO4), sodium bromide (NaBr), and toluene were all added to a soil sample. The toluene phase was collected and








50



Table 2-3. An overview of the analytical methods for
mercury in aqueous samples.

Analytical Methods Species Sources

Gold Trap CVAAS Hg(Tot) Xiankun et al.(1990)

GC-HS-MIP CH3Hg Lansens et al.(1990)

GC-ECD CH3Hg Lee & Hultberg(1990)

He-dc plasma AED Hg(Tot) Lee & Iverfeldt(1991)

GC-QFAAS CH3Hg Rapsomanikis & Craig (1991)

Gold Trap CVAFS Hg(Tot) Baeyens (1992)

GC-CVAAS CH3Hg Sarzanini et ai.(1992) Gold Trap with ICPMS Hg(Tot) Smith (1993) CVAAS Hg(Tot) CH3Hg Garcia et al. (1994) CVAFS Hg(Tot) Saouter et al.(1995) GC-CVAFS Hg(II), CH3Hg Gilmour & Bloom(1995) Gold Trap CVAAS Hg(Tot) Paquette & Helz (1995)


GC = gas chromatography
CVAAS = cold vapor atomic absorption spectrometry CVAFS = cold vapor atomic fluoresence spectrometry GC-HS-MIP = headspace gas chromatography with microwave
induced plasma detection
QFAAS = quartz furnace atomic absorption spectrometry ICPMS = inductively coupled plasma mass spectrometry AED = atomic emission detector ECD = electron capture detector









51

mixed with an equal volume of sodium thiosulfate (Na2S203) in ethanol. Potassium iodide (KI) and benzene were eventually added to the previous mixture. The benzene phase, which contains the MMHg was then analyzed on the GC. The detection limit for MMHg was 30 ppb.

Hempel et ial. (1992) determined organic Hg species in soils by high-performance liquid chromatography (HPLC) with ultraviolet (UV) detection. The simultaneous separation and quantification of nine organic Hg compounds was performed. This separation was done on octadecylsilane columns by gradient elution with a methanol-water mixture ranging from 30 to 50% v/v. The detection limits for the various compounds were in the range 70-95.1 pg/dm3.

Evans et al. (1984) determined organic Hg in peat, sediment, and biological samples. The sediment aspect involved the use of acetone as an extractant, KBr and CuSO4 in sulfuric acid and finally toluene as the last extractant. Mercury analysis was done using a graphite furnace AAS with the detection limit for MMHg set at 25 ng/g.

Mikac and Picer (1985) investigated Hg distribution in a polluted marine area. The methylmercury determination in the sediment samples involved HC1 hydrolysis of the wet sample and









52

extraction of CH3HgC1 into benzene. This extract was then dried with Na2SO, and analyzed by GC/ECD.

Sakamoto et al. (1992) reported on the differential determination of organic Hg, Hg (II) oxide, and Hg (II) sulfide in sediments by CVAAS. All of these forms were removed from sediment by extracting with a solvent such as chloroform followed by a thiosulfate addition and subsequent analysis by CVAAS after digestion.

Engstrom et al. (1994) studied atmospheric Hg deposition to lakes and watersheds. Sediment cores were obtained from various systems and the samples were digested with a strong acid-permanganate-persulfate digestion technique. CVAAS was used for the total Hg analysis of these samples.

Zhang and Planas (1994) reported on biotic and abiotic Hg methylation and demethylation in sediments. Methylmercury was extracted from sediment samples using CuSO4/NaBr/H2SO4 and toluene and analysis was performed by GC.

Rood et al. (1995) investigated Hg accumulation trends in Florida Everglades and Savannas Marsh flooded soils. The Hg in the soil and sediment samples underwent a strong digestion using an acid-permanganate-persulfate combination with final analysis by CVAAS.









53

Municipal Solid Waste

Municipal solid waste consists of community refuse, including garbage, rubbish and trash (Nathanson, 1986). Average composition of the MSW stream in the USA includes paper and paperboard (41.0%), glass (8.2%), metals (8.7%), plastics (6.5%), rubber and leather (8.1%), food wastes (7.9%), yard wastes (17.9%) and other miscellaneous materials (1.6%) (Tchobanoglous, 1993; Concern, Inc., 1988). Predictions of increases in MSW (MSW) up to the year 2000 indicate increases in paper and paperboard, rubber and leather goods, plastics, yard waste, and metals. There are decreasing trends for food waste and glass (Tchobanoglous, 1993; Concern, Inc., 1988). In 1987-1988, 80% of MSW generated in the USA was landfilled, 10% was incinerated and 10% was recycled. Other amounts were dumped illegally on land and in the ocean. MSW is regulated under Subtitle D of the Resource Conservation and Recovery Act (RCRA) in 40 CFR Part 257 (USEPA, 1988).

In Florida, for the 12 month period ending June 30, 1993, 21.5 million tons of MSW were generated (up from 20.3 million tons in 1992). Of that amount, 9.9 million tons were landfilled (down from 10.4 million tons in 1992) 6.6 million tons were recycled (up from 5.4 million tons in 1992) and 5









54

million tons were combusted (up from 4.5 million tons in 1992)( Florida Department of Commerce, 1994). This is an excellent indication, at least within the State of Florida, that landfilling is decreasing and recycling is on the rise.

The MSW stream contains hazardous household and commercial toxic products of which heavy metals are included. The presence and mobility of heavy metals in MSW have been a source of concern for quite some time. MSW is landfilled, incinerated, or composted and the environmental issues with heavy metals focus on groundwater contamination, air pollution, and uptake by crops and other animals, and eventually humans. Hazardous wastes are a subset of MSW and can cause serious illness, injury, or death. They also present a problem to the environment if improperly transported or disposed (Nathanson, 1986). Heavy metals can be classified as hazardous wastes due to their various characteristics as stated in 40 CFR 261 subpart C which lists the four characteristics of hazardous waste as ignitability, corrosivity, reactivity, and extraction procedures (EP) toxicity. On the other hand, there are many items that are excluded from the definition of hazardous wastes in EPA regulations, of which the first item listed is household









55

waste. Many heavy metals found in MSW are actually products of household waste (Congress of the United States, 1983). Types of Heavy Metals

Many different types of heavy metals are found in the MSW stream. It is important to note that aluminum and tin are not considered when referring to heavy metals. Included in the classification of heavy metals that exist are antimony (Sb), arsenic (As), cadmium (Cd), copper (Cu), chromium (Cr), iron

(Fe) lead (Pb) manganese (Mn) nickel (Ni) zinc (Zn) mercury (Hg), selenium (Se), silver (Ag), and tin (Sn) Maximum concentration levels (in mg/l) as determined with the TCLP method for some of the heavy metals are 5.0 for arsenic, 100.0 for barium, 1.0 for cadmium, 5.0 for chromium, 5.0 for lead, 0.2 for mercury, 1.0 for selenium, and 5.0 for silver (Congress of the United States, 1983).

Cadmium, copper, nickel, zinc and molybdenum have been identified as the heavy metals with the greatest potential to accumulate in plants. Cadmium can easily accumulate in crops at concentrations which could increase the dietary intake of the metal without causing crop phytotoxicity. Lead, mercury, arsenic and selenium are not considered to be much of a threat to plants because they have low solubility in slightly acid or









56

neutral, well-aerated soils (Louisiana Cooperative Extension Service, 1991).

Sources

The major sources of heavy metals in MSW are lead acid batteries, household batteries (auto and flashlight), consumer electronics, plastics, used motor oil and transmission fluid, magazine ink, stain, varnish, and sealant (Bretz, 1990; USEPA, 1988). Fluorescent lamps, high pressure sodium lamps, and high intensity discharge lamps contain mercury. To date, the lamps still exceed the regulatory level of 0.2 mg/L for mercury and also exceed the regulatory level of 5.0 mg/L for lead when TCLP are performed (USACHPPM, 1997). Outdated light bulbs contain around 20 to 80 milligrams of liquid mercury. The breaking of one bulb can poison 50 cubic meters of breathing air (Kovalov, 1994).

Household batteries are major contributors of heavy metals in the solid waste stream. Each year, over two billion batteries are deposited in SW facilities in the United States (USEPA, 1997) Therefore, in a city of 500,000 people, almost 9,000 pounds of Hg would potentially enter the air, earth, and water (Jade Mountain, Inc., 1997). The different types of battery systems may contain Hg, cadmium, nickel, zinc,









57

manganese, and lithium. Although each of these metals have negative health and environmental effects, Hg and cadmium are of greatest concern. It has been estimated that up to half of the Hg that is utilized in the United States is used in batteries. Batteries account for about 54% of cadmium in the solid waste stream. Fluorescent lights, Hg-vapor lamps, arc lamps, light switches, mirror coating, amalgam, and electrical apparatus are other sources of Hg in MSW, but to a much lesser degree than that of batteries (Bureau of Solid and Hazardous Waste, 1995; Tchobanoglous, 1993; Steinwachs, 1990). Table 24 identifies the discarding of Hg containing products into the MSW stream.

Waste Recovery and Resource Treatment Combustion

Combustion is defined as the chemical reaction of oxygen with organic materials, to produce oxidized compounds accompanied by the emission of light and rapid generation of heat (Tchobanoglous, 1993). MSW combustion (incineration) has two main functions, (1) reduction in the volume of waste subject to final disposal, and (2) recovery of energy. Combustion facilities are referred to as waste-to-energy (WTE ) units which are capable of producing steam and electricity









58


Table 2-4. Discard of Hg Containing Products into the
MSW Stream (tons)


Product United States Florida 1989 2000 1989 2000 Household Batteries 621.10 98.50 32.30 5.12 Mercury Light Switches 0.40 1.90 0.02 0.10 Electric Lighting 26.70 40.90 1.39 2.13 Paint Residues 18.20 0.50 0.95 0.03 Fever Thermometers 16.30 16.80 0.85 0.87 Thermostats 11.20 10.30 0.58 0.54 Pigments 10.00 1.50 0.52 0.08 Dental Uses 4.00 2.30 0.21 0.12 Special Paper Coating 1.00 0.00 0.05 0.00




Adapted from Bureau of Solid and Hazardous Waste, 1995.









59

and can be used in conjunction with source reduction, recycling and composting programs (USEPA, 1989). For MSW, excess air is used to ensure complete combustion of the organic fraction. This is represented by the equation: Organic matter + excess air ~ N2 + CO2 + H2 0 + 02 + ash + heat

The endproducts of combustion include hot combustion gases and noncombustible residues. The heavy metals are normally found in the noncombustible residue known as ash. Heavy metals emissions from municipal waste combustion (MWC), particularly Hg, are cause for concern because significant amounts of Hg are released through incinerator stack emissions. Mercury can vaporize at very low temperatures (40-500C), therefore it is easily emitted into the environment after which it can be converted into various forms, one of which is methylmercury which bioaccumulates in the food chain (Steinwachs, 1990). In 1986, EPA issued operational guidance on control technology for MWC. The final regulations were issued in December 1990 to reduce deleterious emissions. Batteries are supposed to be separated from MSW prior to incineration to reduce the heavy metal emissions. In the past, dead batteries in the waste









60

stream were the principal source of heavy metals in the fluegas stream (Bretz, 1990).

Metzger and Braun (1987) investigated in-situ Hg speciation in flue gas by liquid and solid sorption systems. Solid sorbents included iodated activated carbon and also gold or silver amalgams. Liquid sorbents included nitric acid/peroxydisulfate solutions. A condensation/absorption approach was used to identify Hg within the sorbents.

In Europe, Hg in gas emissions from waste incinerators consists of 20% of elemental Hg, 60% of divalent inorganic Hg compounds, and 20% of particulate Hg (Petersen et al., 1995). In Europe, the Hg content in MSW ranged from 0.3 to 9 g/t. In North America, the Hg content ranged from 0.36 to 5.8 g/t. Medical waste incinerators emit an average of about 30 g Hg/t in developed countries and 10-20 g Hg/t in developing countries (Pirrone et al., 1996).

Krishnan et al. (1997) investigated Hg control in MWCs and coal-fired utilities. The report showed the ability of activated carbons PC-100 and FGD to capture Hgo and HgC12 emitted from these two systems. The precursor for PC-100 is bituminous coal and for FGD, it is lignite. The PC-100









61

captured more Hgo than FGD at both temperatures for MWC and coal-fired simulations.

Carroll et al. (1995) investigated Hg emissions from a hazardous waste incinerator equipped with a State-of-the-Art wet scrubber. Scrubber collection efficiency for Hg averaged 87% which was lower than expected. Gowin et al. (1993) studied the ability of triple-reverse-burned coal char (TRB char) to remove Hg vapor from the gasification reactor and determined that it was an effective sorbent for this purpose.

The main problem with MWC stems from the production of residual ash. The inorganic, noncombustible portion of the waste stream and the uncombusted organic matter are the constituents in the ash. Heavy metals remaining in incinerator ash are in a more leachable form. Exposure to these heavy metals occur in two ways. First of all, respirable particles may disperse into the air from the stack or during transport and handling. Next, leaching of the metals may occur at disposal, contaminating ground and surface water (Bretz, 1990). Fly ash and bottom ash often contain high levels of heavy metals that require special handling and burial in ash landfills. Standards for the handling, processing, disposal and recycling of MSW combustor ash in









62

Florida are found in Chapter 17-702, F.A.C., and were adopted by the Environmental Regulation Commission in June, 1990 (Department of Environmental Regulation, 1991).

Heavy metal concentrations and percentage distributions in soil samples confirm a strong tendency of heavy metals to accumulate in solid effluents and concentrate in fly-ash. This is one possible explanation of why there is a high amount of metals emitted per ton of burned waste (Morselli et al., 1992). In the past, ash was just disposed of in a landfill specifically created for ash disposal called a monofill, but ash reuse is currently on the rise.

Ash recycling involves the use of ash in soil cement for road sub-base material, the use of ash as aggregate in road asphalt, and the construction of blocks from the ash for artificial reef construction (Department of Environmental Regulation, 1991).

Composting (Aerobic)

In MSW composting, preprocessing is performed to isolate the compostable portion, i.e. yard wastes, food wastes, and organic fractions such as paper. These materials constitute about 30 to 60 percent of the MSW stream. Separation of the compostable portion is generally performed using a rotating









63

screen called a trommel. Once these are separated, they are usually shredded to reduce the particle size and moisture may be added to aid the composting process (USEPA, 1989). The main requirement for compost is that it should be suitable for agricultural use as an organic soil conditioner. Therefore, physical, chemical and biological stability, nonphytotoxicity and balance among mineral elements are the primary characteristics for compost to be useful to the soil and for crops (de Bertoldi et al., 1990).

Heavy metals do not degrade, but tend to be concentrated during the composting process. Metals of greatest importance are those which bioaccumulate, resulting in long or short-term toxic effects to organisms in the environment. Those most commonly regulated include Cd, Hg, Pb, Ni, and Zn. To date, most MSW composts have achieved regulatory limits for most metals, with the exception of lead, but there is an interest in attaining lower levels. The earlier that sorting occurs during the collection and composting process, the lower the heavy metal content in the finished compost. Source separation of a few contaminants such as lead-acid batteries and television picture tubes will definitely decrease the









64

heavy metal content of MSW compost (Richard and Woodbury, 1994).

One report determining how composting affects heavy metal content stated that there was a great occurrence of water soluble heavy metals, except for Cd. The concentrations of heavy metals were generally lower in the samples taken at the end of the composting period in the correctly prepared compost. In correctly produced compost maturation increases the humic acids with respect to the fulvic acids which does not occur in incorrectly produced compost. This is an indication of the process of humification of organic matter leading to the unavailability of heavy metals as the heavy metals would now be complexed to the humic material, hence reaching the soil in a complexed, less mobile form. Therefore, there would be a decrease in the leaching of heavy metals. About one-third of the total content of heavy metals present in compost was reportedly bound to the alkali soluble organic matter. (Canarutto et al., 1991).

Heavy metal concentrations in the finished compost determine its usage. Lower heavy metal content allows the compost to be used for practical applications, such as landscaping. Higher metal concentrations in compost prohibit









65

its usage because of the potential for plant uptake, or leaching into the environment (Steinwachs, 1990). Land application rate of heavy metal containing composts are dependent on the concentration of the metal, the pH of the soil and the cation exchange capacity of the soil. The decrease in concentration of metals in the soil compost mixtures results from leaching, with very little as a result of plant uptake.

In aerobic biologically stabilized solid waste residues, one of the main forms of Hg will likely be mercuric oxide (HgO), along with other possible forms of Hg such as HgC12 which is likely to occur in the absence of sulfides. Chloride is normally a weak complexing agent, but it strongly complexes with Hg2" and Hg22+ (Pohland et al., 1981). HgO may be dissolved and mobilized by acid rain percolating through residue-amended soil, increasing the chances of Hg being taken up by plants, as well as the possibility of leaching. Landfilling

In anaerobic environments, such as sediments or landfills, or under mildly reducing conditions, there is a tendency for Hg to be precipitated as the sulfide (HgS). Mercuric sulfide is highly insoluble (log K. = -50.02) (IUPAC,









66

1982), but can be transformed under low Eh and high pH
to 2+
conditions to the soluble form, HgS2 or to the free metal (Barkay, 1992;, Wilken, 1992; Schuster, 1991). The insolubility of HgS makes it resistant to methylation, but under aerobic conditions, HgS may be oxidized to the sulfate form which can then undergo methylation (Adriano, 1986).

Gases found in landfills include methane (CH4), carbon dioxide (CO2) ammonia (NH3), carbon monoxide (CO), hydrogen sulfide (H2S), nitrogen (N2), and oxygen (02) (Tchobanoglous, 1993). When considering the movement of Hg in the landfill and the transformations that occur, it can be hypothesized that dimethylmercury and elemental Hg may actually be emitted into the atmosphere. Anaerobic bacteria may act on the available Hg in the landfill and transform it to either monomethylmercury or dimethylmercury. The vapor pressure of monomethylmercury is quite low (H < 10-) while the vapor pressure of dimethylmercury is quite high (H < 0.3) (Barkay, 1992). As a result, this form of Hg may also be given off from a landfill in the same manner as methane or carbon dioxide which are the two most prevalent gases resulting from the anaerobic processes occurring in a landfill.









67

Under warm or hot conditions, Hg can be transformed to elemental Hg which is the second most volatile form of Hg (H=0.3), next to dimethylmercury. If this transformation does in fact occur in a landfill, then without proper control mechanisms, elemental Hg would be emitted into the atmosphere.

Heavy metals dissolved in aqueous systems exist in combination with other chemical species in the form of complexes. Metal ions, such as Hg2", combine with non-metallic compounds known as ligands by means of coordinate-covalent bonds. In anaerobic residues, the species of Hg found will likely be mercuric sulfide (HgS). Cinnabar or mercuric sulfide is about the most stable Hg mineral that occurs in nature. This Hg mineral forms under reducing conditions where sulfate (S042-) has been reduced to sulfide (S2-) (Mitra, 1986). Gambrell et al. (1978) showed that HgS is unavailable to plants. Therefore it may be feasible to use anaerobically stabilized solid waste residues for land application. If Hg is in the HgS form, there should be minor, if any, adverse impacts on crops, and the groundwater should remain uncontaminated as HgS is quite immobile and unlikely to leach.

The dumping of trash is becoming a less viable solution since the siting of landfills is extremely difficult as a









68

result of public opposition. The public invokes the NIMBY (not in my back yard) syndrome due to the odors and the potential contamination to surface and groundwater that may occur due to leaching. However, as technology improves and liners become more reliable, landfills are meeting federal and state standards, which should actually ease the public's minds when it comes to the siting of landfills. Some communities that are faced with diminishing landfill space send their waste to other states, a practice sanctioned by the Supreme Court ruling in 1978 (Philadelphia v. New Jersey) that a state may not refuse waste from another state. This poses a problem for those communities that are trying to extend the life of their landfills by recycling or reducing their own wastes when the space is filled by other counties or states (Concern, Inc., 1988).

The Solid Waste Management Act (SWMA), Senate Bill No. 1192, was passed by the Florida Legislature and became effective on October 1, 1988. Under this Act, certain items were banned from landfills. These are used motor oil which became effective October 1, 1988, lead acid batteries, which became effective on January 1, 1989, white goods which became effective on January 1, 1990, and yard trash which became









69

effective January 1, 1992. Exclusion of these items from the landfill will reduce the amount of hazardous materials that can potentially contaminate groundwater or the surrounding land (Earle et al., 1991), especially heavy metals.

Landfill leachates often contain high concentrations of toxic heavy metals. Many of these metals can form strong complexes with biomolecules, therefore, even in small amounts their presence can have adverse effects on both plants and animals (Bolton and Evans, 1991). Some probable ligands in landfill leachates and residues would be chlorides, sulfates, phosphates, and organics (Pohland et al., 1981). Table 2-5 shows the median concentrations in MSW leachate, in comparison with existing exposure standards. It was interesting to note that on an overall basis, the USEPA (1988) report had lower levels than those given in the Congress of the United States (1989) report. The fact that certain items containing heavy metals were banned in 1988 and 1989 would have led one to believe that the heavy metal contents would have decreased. Obviously, items such as batteries are still finding their way into landfills even though there is now an emphasis on rerouting them away from landfills and recycling the relevant










70

Table 2-5. The median concentrations in MSW leachate, in
comparison with existing exposure standards. Metals Median concentration Exposure (ppm) standards (ppm)

Antimony 0.066 0.01 Arsenic 0.042 0.05 Barium 0.853 1.0 Beryllium 0.006 0.2 Cadmium 0.022 0.01 Chromium 0.175 0.05 Copper 0.168 0.012 Iron 221 1000 Lead 0.162 0.05 Manganese 9.59 0.05 Mercury 0.002 0.002 Nickel 0.326 0.07 Selenium 0.012 0.01 Silver 0.021 0.05 Thallium 0.175 0.04 Zinc 8.32 0.110



(USEPA, 1988 and Congress of the United States, 1989).









71

metals. Landfill leachates contain many organic and inorganic ligands, particularly chloride (Cl-), which help to determine the forms of metals in solution. High concentrations of inorganic ligands, primarily chloride (Cl-), and high dissolved organic carbon (DOC) concentrations affect the speciation of metals in landfill leachates. Complexation reactions in solution lower the positive charge associated with the metal and therefore may affect the mobility of leachate metals in the soil and underlying sediments (Bolton and Evans, 1991). Therefore, methods have to be developed to decrease the amount of heavy metals in leachate.

Leachate recycling is one method to decrease the amount of heavy metals present. One report showed that leachate recycling did not result in an increase in heavy metal concentrations, but actually removed some of the heavy metals in the solid waste columns. The explanation provided was that under anaerobic conditions that exist in a landfill, heavy metals precipitate as sulfides (Pohland et al., 1979). Anaerobic Digestion

The main steps in anaerobic digestion of MSW are: (1) Pretreatment to remove undesirable material, upgrade and homogenize the feedstock for digestion and to protect









72

downstream treatment processes; (2) Anaerobic digestion to produce biogas for energy and to deodorize, stabilize, and disinfect the digestate product; (3) Post-treatment to complete the stabilization and disinfection of the digestate, to remove residual inert undesirable material (glass and plastics) and produce a refined product of suitable moisture content, particle size and physical structure for the proposed end-use. The end uses are energy in the form of biogas, typically around 100-200 ml biogas per ton of organic MSW digested, also solid and liquid by-products which can be used as compost products and have a value as a fertilizer or soil improver (IEA Bioenergy, 1994).

The majority of the literature on anaerobic digestion deals with zinc (Zn2+), iron (Fe3) cadmium (Cd2) copper (Cu2*) chromium (Cr3* Cr6) Nickel (Ni3 ) and lead (Pb3) The focus of most of these articles was the inhibition of anaerobic digestion by these heavy metals. It was determined that concentrations and species of these metals were indicative of their toxicity and the ability of microorganisms (bacteria) to adjust to their effects. Temperature, pH, hydraulic retention time, and the ratio of the toxic substance concentration to the bacterial mass concentration are









73

determinants in the inhibitory concentrations of these heavy metals. Some inhibitory concentrations given were Cd2" (180 ppm), Fe3+ (1750 ppm), Cu2 (170) ppm, Cr6+ (450 ppm), Cr 3 (530 ppm), and Ni3" (250 ppm) (Chynoweth and Pullammanappallil, 1996; IEA Bioenergy, 1994; Peiffer, 1993; Lin, 1993; Lin 1992; Mueller and Steiner, 1992).

Determination of the heavy metal content after the digestion process is the principal determining factor in the feasibility of using the liquid or solid digestate as compost. Post-treatment follows the actual anaerobic digestion process which allows for two to four weeks of maturation for the solid fraction whereas the liquid fraction may be directly applied onto farmland as slurry, if it is of good quality and meets all set regulations. Review of the literature provided no information about the presence of Hg. Hence, it is obvious that Hg has rarely been included in past studies.

In an anaerobic environment, microorganisms that do not need oxygen are called anaerobes. There are two types of anaerobes, obligate and facultative. Obligate anaerobes cannot use oxygen at all, and tend to be poisoned by it. Facultative anaerobes are able to use oxygen when it is present, but can use other sources when oxygen is absent. The









74

other sources that replace oxygen are nitrate (NO3-), sulfate (S042-), and carbon dioxide (C02). When these compounds are used in the energy-generating process they are reduced. Nitrate is reduced to nitrogen gas (N2), sulfate is reduced to hydrogen sulfide (H2S), and carbon dioxide is reduced to methane (CH4). Bacteria that use nitrate are called denitrifying bacteria and those that use sulfate are called sulfate reducers, and those that use carbon dioxide are called methanogenic bacteria (methanogens). Denitrifying bacteria are facultative anaerobes and sulfate reducers and methanogenic bacteria are called obligate anaerobes (Brock and Madigan, 1988; Brock and Brock, 1978).

In an MSW anaerobic digester, there is a consortium of microorganisms that play a part in methane fermentation. Large-molecular weight compounds such as polysaccharides, proteins, and fats are converted to methane by these microorganisms. For conversion of a typical polysaccharide, this consortium includes cellulolytic bacteria which cleave high molecular weight cellulose molecule into cellobiose and free glucose, fermentative anaerobes which ferment the glucose to a variety of products including acetate, propionate, butyrate, hydrogen (H2), and CO2. Methanogenic bacteria









75

consume any H2 that is produced. Also, certain methanogens are capable of converting acetate to methane (Brock and Madigan, 1988).

Other organisms that convert complex materials to methane are H2-producing fatty acid-oxidizing bacteria which use fatty acids or alcohols as energy sources and grow very well in the presence of H2-consumers such as methanogens and sulfate reducers. The association between hydrogen producers and hydrogen consumers is known as interspecies hydrogen transfer which is important in the anaerobic digestion process. Syntrophomonas and Syntrophobacter are H2-producers which generate acetate, CO2, and H2. Hydrogen-consuming acetogenic bacteria (Acetobacterium) consume hydrogen and produce acetate for methanogenesis. During methanogenesis, methanogens reduce the acetate, CO2 and H2 to methane and carbon dioxide which are the main end products of anaerobic digestion (Chynoweth and Pullammanappallil, 1996; Brock and Madigan, 1988).

This literature review has included all three phases of the environment, namely air, earth, and water, and the many Hg interactions that take place in the environment. This literature review showed that data of Hg in landfills and MSW









76

is extremely limited. However, data on metal concentrations in landfill leachate is well documented. The fact that Hgcontaining devices are still finding their way into landfills in spite of more stringent recycling efforts, is the primary reason why more information is needed on the fate of Hg in landfills. The literature on anaerobic degradation processes strongly suggests that transformation of inorganic and elemental Hg to volatile organic forms is highly likely to occur in landfills.















CHAPTER 3
MATERIALS AND METHODS



Phase I-Hq in the Alachua County Landfill



Samples that were analyzed for total Hg in this study had been obtained from the Alachua County Landfill during a previous study (Miller et al., 1996; Miller et al., 1994). Five areas were sampled and were designated as LFS1, LFS2, LFS3, LFS4, and LFS5 (Figure 3-1). LFS1 is a leachate recycle area (Infiltration Pond 1), LFS2 is a control area for leachate recycling, LFS3 is a 4.5 hectares closed landfill, LFS4 is a 12 hectare closed landfill, and LFS5 is a leachate recycle area with horizontal injection. The active lined landfill (11 hectares) at this site began receiving MSW in 1988. Leachate recirculation was performed on a portion of this lined landfill (LFS1) which is adjacent to the leachate recycle area. LFS2 is actually a portion of the lined landfill that was used as a control area where there was no



77













LFS1: Leachate Recycle Area (Infiltration Pond 1) LFS2: Control Area LFS3: 0.05 km' Closed Landfill LFS4: 0.12 km' Closed Landfill LFS5: Leachate Recycle Area (Horizontal Injection)
4.5 hectare unlincd capp Cd landfill unic LFS 4

LFSl

'73 '85 11 hectare lined landfill unic
LFS 3


12 hectare unlined
LFS5 '85 '88 capped landfill unit



'88 Present






200 fi






Figure 3-1. Alachua County Landfill (Lee, 1996)




OD









79

leachate recirculation occurring. LFS3 received waste from 1985 to 1988 and LFS4 received waste from 1973 to 1985 and both were in a portion of the landfill that was unlined.

Solid waste samples were collected from LFS 1 both prior to and following the filling of the infiltration pond with leachate. Three boreholes were normally augured at each site with a 10.16 cm truck mounted open flight power auger that penetrated through the landfill until it reached the desired depth. In general, four samples were collected per hole at 3.05 m (10 feet) intervals. The maximum depth of the boreholes ranged from 9.14 13.72 m. The top portions of some samples (0 1.52 m) were discarded because they were mostly cover soil. Table 4-1 shows the depths of each collected sample. Additional information on sampling procedures can be found in Miller et al. (1996)

After collection, the samples were placed into polyethylene bags and brought back to the Bioprocess Engineering Research Laboratory of the Agricultural and Biological Engineering Department at The University of Florida and stored in freezers until processed. Processing of these samples involved separation and characterization of waste components, moisture content, and solids composition









80

determinations and biochemical methane potential assays. The samples were dried at 1050C, grossly picked, and then passed through a 6 mm screen. The material that remained on the 6 mm screen was finely picked and ground to 0.76 mm using an Urschell Mill (Comitrol Model 3600). The material that passed the 6 mm screen was then passed through a #40 screen (0.42 mm). The material that remained on the #40 screen was finely picked and then ground to 0.76 mm using the Urschell Mill. For the purpose of total Hg analyses, the samples consisted of three fractions. One fraction was that remaining on the 6 mm screen, a second fraction was that passing the #40 screen, and the third fraction was that remaining on the #40 screen. These are shown in Table 4-1. These three fractions represented the entire sample taken from the landfill after gross picking and fine picking. In the study by Miller et al. (1996) the SW from the landfill was also characterized as paper, cardboard, food waste, glass and stone, metal, wood, plastics and rubber, and fabrics. Gross and fine picking removed glass and stones, metal, wood, plastics, rubber, and fabrics, leaving essentially paper, cardboard, and food wastes.









81

The elaborate separation scheme used by Miller et al. (1996) was developed to obtain the fraction of SW that is directly responsible for methane production, namely the volatile organic fraction. This is shown in Figure 3-2. For Phase I of this study, the three fractions described above were analyzed to evaluate total Hg concentrations existing in the landfill and to give an idea of the Hg distribution that occurred in such an environment. Mercury was being evaluated because there has been little research concerning Hg in biologically stabilized SW residues. The fact that these samples are volatile organic fractions are important concerning Hg because Hg binds strongly to organic material, therefore the majority of the Hg in the landfill should be bound to any existing organic matter present.

Total Hg in these samples was measured on a ng/g dry weight basis to give an idea of the Hg levels existing in the Alachua County Landfill. These Hg concentrations could have been converted back to a wet weight basis, but it was not deemed necessary since it is not scientifically incorrect to report on a dry weight basis. Furthermore, if the landfill is ever reclaimed and the SW applied to land, application rates will be computed on a dry weight basis. It must also be noted














82





















S,-rnrluf (b i cen




) Jngile
r IL inn tI









crena6 (6 m m creen I~ftti





T


ine picking & S-mrle characltrtu ion






Cite orad sionbtodeg pari


......... .. ....... ...... ..- -...--


.-rrr nm g (No 0J .creen) B: PAper, cardboard S
food waste
C: CLUss & slor ?sf. MetW Reoniiing.sin Paon n V: NN ood on through No.40 P: Plastics & rubber .No, 0 No 40 screer r rn F; Fahrks sctrr





Firk picking








Flrne hiodeg flt F innotbsourg( "<2,






Lbrd for

G rindirg to 0.03" jize...... .... ....... .. r ly e ..

u~ .-. mnl..H.n.li n --- --Procedure- (Lee--1996)





Figure 3-2. Sample Handling Procedure (Lee, 1996)









83

that the Hg values reported in this study may be lower than the actual values because the samples had been dried at 1050C in the study of Miller et al. (1996). Elemental Hg vaporizes at around 40-500C and any elemental Hg that may have been in the original landfill samples would have been lost in the drying process.

Total Hg was determined using standards ranging from 0 to 550 ng Hg. A check sample was used in every run as a way to evaluate whether or not the system was performing adequately and also as a reference point to the validity of the other samples. A duplicate and a spike were run with each sample batch (12 samples) and a continuous calibration blank and a continuous calibration variance (a 50 ng Hg spike of a blank sample) were used throughout every sample batch to ensure that the instrument absorbances were being maintained. The acceptance criteria for the calibration curves (6 standards, 1 blank) was r2 > 0.99. The blanks had to be below the Method Detection Limit, the spikes had to be +/- 20% of the true value, and for duplicate samples, the standard error had to be < 20%. This QA/QC was used in a previous Hg study (Delfino et al., 1993). Results of the QA/QC are given in Appendix A.









84

Total Hg in the SW fractions was determined using the digestion procedure described in EPA Method 7471 for the determination of Hg in soil and sediment. Following the degestion procedure, Hg was analyzed by cold vapor atomic absorption spectrophotometry (USEPA, 1986). Two grams of sample were weighed to the nearest 0.0001 g on an enclosed Mettler analytical balance. The sample was transferred quantitatively to an acid rinsed 300 mL BOD bottle with a 10 mL double distilled deionized (DDDI) water rinse. The digestion involved 2.5 mL of trace metal grade concentrated nitric acid and 5 mL of trace metal grade concentrated sulfuric acid. The sample was heated at 950C for two minutes, then 15 mL of potassium permanganate (50 g/L), and 8 mL of ammonium peroxydisulfate (50 g/L) were added to the digestion mixture. The sample was heated at 950C for one to two hours. An additional 15 mL of potassium permanganate was added if the color disappeared within fifteen minutes of the initial addition. Upon completion of digestion, samples were cooled and decolorized by the addition of 6 mL of hydroxylamine hydrochloride solution (120 g hydroxylamine sulfate, and 120 g sodium chloride per liter of deionized water solution). Figure 3-3 shows a schematic of the procedure.










85








2 g dry solid waste sample +
10 mL DDDI in BOD bottle


Y



Add 5 mL sulfuric acid and
2.5 mL nitric acid


YI




Add 15 mL potassium
permanganate





Add 8 mL ammonium
peroxydisulfate


SHeat at 95 C for 1 hour Cool and decolorize samples
with 6 mL hydroxylamine hydrochloride solution






Measure elemental mercury
absorbance on CVAAS





Figure 3-3. Total mercury analysis of solid waste
fractions as described in EPA Method
7471.









86

Trace metal grade nitric and sulfuric acid are oxidizers. Potassium permanganate is another strong oxidizer that was added to eliminate possible interference from sulfide. Ammonium peroxydisulfate is another added oxidizer to ensure the complete oxidation of the sample. Heating the sample in a water bath, along with the strong oxidizers, converts all of the Hg to Hg2".

Each digested SW sample was transferred to a plastic reaction vessel fitted for a Perkin Elmer MHS-10 cold vapor unit. The sample was purged with high purity nitrogen gas. Stannous chloride was used to reduce Hg2" to the elemental state, Hgo. Elemental Hg was then swept into an open ended quartz tube (1 cm diameter) with a 16 cm cell path length. The Hg was quantified by CVAAS using a Perkin Elmer Model 3030 Atomic Absorption Spectrophotometer (X = 253.6 nm, SBW = 0.7 nm) with Hg hollow cathode lamp (I = 6 MA). The standard calibration curve working range (0 to 550 ng) gave an absorbance range of 0.017 to 0.260.

Total Hg (ng/g) was determined using the equation derived from the regression analysis of the Hg standard curve. From the equation:




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THE FATE OF MERCURY IN MUNICIPAL SOLID WASTE LANDFILLS AND ITS POTENTIAL FOR VOLATILIZATION By CELIA D.A. EARLE 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 1997

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ACKNOWLEDGMENTS I would like to thank the Lord Almighty who has been with me constantly throughout this period of time. He has given me the strength that I needed to go on even when I felt like giving up. Without Him, I could not have succeeded. I would like to thank my parents, Jonathan and Yvonne, my brothers Kevin and Jeremy, and my grandmother, Evelyn Lindo, for their continued prayers and support throughout these trying times. Special thanks to my wonderful chairman. Dr. R. Dean Rhue, who has steadfastly encouraged and guided me in the way that I should go. He has believed in me and has not limited me to any extent He has provided monetary support and emotional support throughout this ordeal. He has been like a true father figure with a genuine concern for my well being and success. I truly thank God for him. Thanks to Dr. David P. Chynoweth, who took me on at a late stage, but has been there for me and supported me. He has provided beneficial input into my research and has also ii

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provided the materials and help that I needed to successfully complete the second phase of my research. Thanks to the rest of my committee, Dr. J.J. Delfino, Dr. K.R. Reddy, Dr. L.Q. Ma, and Dr. T. Townsend, for their constructive criticisms and guidance. Their support also helped me to achieve my goals and for that I am very grateful Thanks to William Reve, lab manager in the Soil and Water Science Department, for being my listening ear and for assisting me in working through the glitches that occurred from time to time. Thanks to Paul Lane, lab manager in the Bioprocess lab, for his assistance in getting the second phase of my research underway. Thanks to Dr. John Owens for instructing me in the use of the gas chromatographs Thanks to Dr. Jose Sifontes, Dr. Paul Gebert and Andreas Sifontes for their assistance with logistics of the reactor sample runs. Thanks to Wendy Stickney for being my lab assistant and constant friend during the second phase of my work, and to Maria Corchuelo who helped with the typing of the references in this document Thanks to Clayton Clark and Adrienne Cooper for their assistance in preparing this document for submission. Thanks to Joyce Taylor and the rest of the administrative staff in the Soil and Water Science department 111

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for aiding me with whatever was required during my time in the department Thanks go to my cousins Deirdre Lawson and Andre Earle; Special friend Paul Mason; friends Florette Earle, Denise Borel, Michael McCorkle, Ingrid Forbes, Ogechi Okpechi, Michael Distant, Steven Trabue, Gerco Hoogeweg, Hector Castro, Tait Chirenje, Gerald Manley; and to my other relatives and friends in the USA and Jamaica who provided love, prayers and support IV

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TABLE OF CONTENTS Page ACKNOWLEDGMENTS ii LIST OF FIGURES vii LIST OF TABLES x ABSRACT xii CHAPTER 1 INTRODUCTION 1 2 LITERATURE REVIEW 6 Forms and Toxicity of Mercury 8 Mercury in the Atmosphere 12 Mercury in Aquatic Systems 15 Terrestrial System 28 Analytical Methods 40 Municipal Solid Waste 53 Waste Recovery and Resource Treatment 57 3 MATERIALS AND METHODS 77 Phase I-Hg in the Alachua County Landfill 77 Phase II -Landfill Simulated Anaerobic Reactors .... 89 MSW Analysis 98 Gas Sampling and Analysis 98 Volatile Fatty Acids 100 pH Analysis 103 Mercury Analysis 103 Tygon Tubing Experiment for Mercury 108 4 RESULTS Ill Hg concentrations in the Alachua County Landfill .... Ill V

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Phase II-Anaerobic Reactor Data 125 Phase 1 1 -Mercury Data 14 DISCUSSION 14 9 Phase I Hg in the Alachua County Landfill 149 Phase II 165 SUMMARY AND CONCLUSIONS 180 APPENDICES A RAW MERCURY DATA FOR ALACHUA COUNTY LANDFILL RESIDUES AND PALM BEACH COUNTY COMPOST 190 B RAW DATA FOR ANAEROBIC REACTORS 211 C RAW MERCURY DATA FOR ANAEROBIC REACTORS 218 D ABBREVIATIONS 23 9 REFERENCES 242 BIOGRAPHICAL SKETCH 262 VI

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LIST OF FIGURES Figure Pa^e 3-1 Alachua County Landfill 78 3-2 Sample Handling Procedure 82 3-3 Total mercury analysis of solid waste fractions as described in EPA Method 7471 85 3-4 Schematic of reactor design 90 3-5 Entire setup of reactor system 93 3-6 Liquid collection traps behind the reactors 93 3-7 Solid Waste insert into reactor 94 3-8 Glass fiber mesh at bottom of solid waste insert 94 3-9 Top view of the reactor system 95 3-10 Close-up view of the activated carbon traps 95 3-11 Modification of EPA Method 7471 for determination of total mercury in sulfurimpregnated carbon 107 3-12 Total mercury analysis in leachate samples as described in EPA Method 7470 109 4-1 Run 1 (Control) pH vs. Days 12 8 4-2 Run 1 (lOOng Hg) pH vs. Days 128 4-3 Run 1 (2000ng Hg) pH vs. Days 128 Vll

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4-4 Run 1 (Control) %Methane vs. Days 129 4-5 Run 1 (lOOng Hg) %Methane vs. Days 129 4-6 Run 1 (2000ng Hg) %Methane vs. Days 129 4-7 Run 1 (Control) Volatile Fatty Acids vs. Days 130 4-8 Run 1 (lOOng Hg) Volatile Fatty Acids vs. Days 130 4-9 Run 1 (2000ng Hg) Volatile Fatty Acids vs. Days 130 4-10 Run 1 (Control) Methane Yield vs. Days 131 4-11 Run 1 (lOOng Hg) Methane Yield vs. Days 131 4-12 Run 1 (2 000ng Hg) Methane Yield vs. Days 13 2 4-13 Run 2 (Control) pH vs. Days 13 3 4-14 Run 2 (lOOng Hg) pH vs. Days 133 4-15 Run 2 (lOOOng Hg) pH vs. Days 134 4-16 Run 2 (2000ng Hg) pH vs. Days 134 4-17 Run 2 (Control) % Methane vs. Days 13 5 4-18 Run 2 (lOOng Hg) % Methane vs. Days 13 5 4-19 Run 2 (lOOOng Hg) % Methane vs. Days 136 4-20 Run 2 (2000ng Hg) % Methane vs. Days 136 4-21 Run 2 (Control) Volatile Fatty Acids vs. Days 137 4-22 Run 2 (lOOng Hg) Volatile Fatty Acids vs. Days 137 4-23 Run 2 (lOOOng Hg) Volatile Fatty Acids vs. Days 138 4-24 Run 2 (2000ng Hg) Volatile Fatty Acids vs. Days 138 4-25 Run 2 (Control) Methane Yield vs. Days 139 Vlll

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4-26 Run 2 (lOOng Hg) Methane Yield vs. Days 139 4-27 Run 2 (lOOOng Hg) Methane Yield vs. Days 140 4-28 Run 2 (2 000ng Hg) Methane Yield vs. Days 14 4-29 Run 1 Hg content in 3 phases (control) 143 4-30 Run 1 Hg content in 3 phases (lOOng Hg) 143 4-31 Run 1 Hg content in 3 phases (2000ng Hg) 143 4-32 Run 2 Hg content in 3 phases (control) 144 4-33 Run 2 Hg content in 3 phases (lOOng Hg) 144 4-34 Run 2 Hg content in 3 phases (lOOOng Hg) 145 4-35 Run 2 Hg content in 3 phases (2000ng Hg) 145 5-1 Hg cone, ranges (ng/g) in the Alachua County Landfill Sc Palm Beach County compost 150 5-2 Mean Hg concentrations with depth for LFSl and LFS2 156 5-3 Hg spike (ng) in simulated landfill reactors & amount of Hg volatilized (ng) 179 IX

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LIST OF TABLES Table Page 2-1 Gas Phase Reactions of Mercury 16 2-2 Aqueous Phase Reactions of Mercury 17 2-3 An overview of the analytical methods for mercury in aqueous samples 5 2-4 Discard of Hg Containing Products into the MSW Stream (tons) 58 2-5 The median concentrations in MSW leachate, in comparison with existing exposure standards 70 4-1 Alachua County Landfill Total Mercury Concentrations in three fractions 112 4-2 % Dry weight of Total Composite Sample 117 4-3 Alachua County Landfill Mercury Concentrations in the composite (weighted sum of the three fractions) 121 4-4 Palm Beach County Mercury Data (1:1 Biosolids to Yard Waste) 126 4-5 Mercury Recovered 142 5-1 Mean Hg concentrations in composite samples for LFS 1 and LPS 2 Alachua County Landfill 154 5-2 Mean Hg concentrations in composite samples for L.S. 2 and L.S. 4 Alachua County Landfill 159

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5-3 Mercury means in different fractions at different sites 162 5-4 MSW Sample Volatile Solids Content Average by Round and Area 164 XI

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Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy THE FATE OF MERCURY IN MUNICIPAL SOLID WASTE LANDFILLS AND ITS POTENTIAL FOR VOLATILIZATION By Celia D.A. Earle December 1997 Chairperson: R. Dean Rhue Major Department: Soil and Water Science Mercury is conveyed into landfills primarily via batteries and a variety of other mercury-containing devices and lamps. Mercury has not been the focal point of metals research in landfills, but the first Phase of this study showed that mercury exists in the Alachua County Landfill at concentrations ranging from 32.8 ng Hg/g to greater than 16,800 ng Hg/g. However, over half of the samples had concentrations of 150 ng/g or less. While Hg concentrations in the landfill samples as well as compost samples (1:1 yard xii

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1 waste to biosolids) from Palm Beach County were generally above background levels for surface soils in Florida, they were two to three orders of magnitude lower than clean-up goals currently used by the Florida Department of Environmental Protection and federal regulations governing land application of sewage sludge as described in 4 CFR Part 503. Phase II of this study investigated the fate of mercury in simulated landfills and found that the bulk of the Hg added was found in the solid waste. Sixteen out of 18 leachate samples did not have detectable levels of Hg. The percentage of Hg volatilized during anaerobic digestion reanged from over 30% at the lowest Hg level (100 ng Hg per 60 g solid waste) to about 3% at the highest Hg level (2000ng Hg per 60 g solid waste. Evidence was obtained that forms of Hg other than elemental or divalent Hg were volatilized. These other forms were thought to be organic Hg compounds, quite possibly including dimethylmercury Sulfur-impregnated, activated charcoal was used to trap Hg volatilized during anaerobic digestion. A modification of EPA Method 7471 successfully recovered >98% of the Hg trapped by the charcoal xiii

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This study has implicated landfills as a potential source of mercury to the atmosphere Further research should focus on quantifying the amount of mercury being emitted into the atmosphere from landfills and identifying those stages during anaerobic digestion of MSW that Hg is most likely to volatilize XIV

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CHAPTER 1 INTRODUCTION The overall objective of this project was to determine the fate of mercury (Hg) in a municipal solid waste (MSW) landfill and to establish its potential for impacting atmospheric pollution. Mercury is one of the most toxic heavy metals in existence. It is present in the atmosphere, aquatic systems, terrestrial systems, and biota that exist within these systems. Mercury is unique in its ability to exist as a liquid at room temperature. It also exists in nature in gaseous, liquid, and solid phases. /Studies have confirmed that anthropogenic activities have increased the input of Hg to the environment. It is estimated that Hg from these activities constitute about half of the Hg entering the environment (Fitzgerald and Clarkson, 1991) Pirrone et al. (1996) have stated that Hg emissions in developed countries increased at a rate of 4.5-5.5% per year until 1989 and have remained somewhat constant since. In developing countries, emissions have steadily climbed at a

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2 rate of 2.7-4.5% per year. Mercury is present in fossil fuels, and to a lesser degree, coal. The vast amounts of fossil fuels and coal burned annually represent an important source of Hg released into the biosphere (Mitra, 1986) Over the years, there has been interest in utilizing solid waste (SW) residues for land application. However, little or no data exist in the literature with respect to Hg levels in MSW landfills. In Florida an attempt is being made to set standards for heavy metals in these residues, and there is need for more data to allow for the establishment of threshold levels in SW. Currently, stabilized MSW residues are tested for heavy metals using USEPA 503 regulations that were developed for sewage sludge. Mercury ranks as number three on the EPA's toxic substances list, behind lead and arsenic. Since Hg is one of the most harmful heavy metals in the environment, the fate of Hg in MSW should be known. Successful completion of this research may assist in the establishment of sampling protocols and regulations specifically for Hg contained in SW residues. Several Hg-containing components enter the SW stream on a daily basis, and eventually make their way to MSW landfills. Fluorescent lights, Hg-vapor lamps, arc lamps,

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mirror coating, amalgam, and electrical apparatus are important sources of Hg in MSW. Although laws were passed in 1988 prohibiting the disposal of batteries in landfills, batteries no doubt still find their way into them (Bureau of Solid and Hazardous waste, 1995; Tchobanoglous 1993; Steinwachs, 1990) X Research has established that Hg binds strongly to organic matter. There is a significant proportion of organic material in SW, and it is expected that the majority of the Hg existing in a landfill will be bound to it. Bacteria commonly found in landfills influence the transformations of Hg that is bound to organic matter. Studies have confirmed that many of these bacteria contribute to methylmercury production. Thus, volatilization of dimethylmercury and elemental Hg from landfills could make a significant contribution to atmospheric J, Hg emissions. It has also been shown that inorganic mercury (Hg^"") can volatilize at the higher temperatures found during anaerobic digestion further contributing to atmospheric emissions Since there is need to understand the fate of Hg in a landfill, a study was conducted with the following objectives: (1) Determine Hg concentrations in SW samples collected at

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^i various depths over a five-year period from the Alachua County SW Landfill in Florida. (2) Evaluate the potential for Hg present in landfills to be volatilized and released to the atmosphere. These objectives were accomplished in two phases. In Phase I, total Hg levels in residues from the Alachua County landfill were determined. This landfill has been in existence since the early 1970 's and has been the primary recipient of MSW from the city of Gainesville since then. Hundreds of residue samples from the Alachua County landfill had been collected and stored during previous studies (Miller et al 1996; Miller et al 1994) These samples were analyzed for total Hg to generate a large database for Hg in MSW residues from a landfill. The results for this landfill should be representative of landfills created by many other municipalities around the Southeastern United States. In Phase II, the fate of Hg occurring in a landfill was studied using laboratory-scale anaerobic digester to which various amounts of Hg were added. In this phase, the distribution of Hg between leachate, solid, and gas phases was determined after a period of anaerobic digestion. Conditions occurring in these small-scale digesters are considered to be

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representative of those that occur in landfills similar to the one in Alachua County.

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CHAPTER 2 LITERATURE REVIEW The earth's crust has been shown to contain low concentrations of mercury. Its main deposits occur in the form of cinnabar, or mercuric sulfide (HgS) which has an average mercury content of 0.1 to 0.4% (Mitra, 1986) In certain locations in the world such as Spain, Yugoslavia, Italy, Peru, Mexico and the American Continent, mercury is found in impregnated schist or slate and as geodes of liquid mercury. Mercury is also found in nature as the oxide and the selenide and in combination with a number of minerals such as quartz, dolomite, chalcedony, calcite, and pyrite. Mercury is present in fossil fuels, and to a lesser degree, coal. The vast amount of coal burned annually (between 1.4 x 10^ and 2.72 X 10^ g/yr) represents an important source of mercury released into the biosphere (Mitra, 1986) ... A large quantity of mercury in the environment is derived from industrially produced mercury amounting to approximately 10,161 metric tons per year. Industries which are suspected

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7 as being responsible for the dispersion of mercury are MSW incinerators, medical waste incinerators, fossil fuel burning, chlor-alkali industries, mining and extraction of mercury from cinnabar, amalgamation, electrical equipment, paper pulp, fungicides, instrumentation, crematoria, degassing of latex paint, and cement manufacturing (Pirrone et al 1996) Exposure to mercury in the home, hospital, or laboratory can occur by the release of mercury from broken thermometers (Stewart and Bettany, 1982; Mitra, 1986). In the past, an estimated 90,718 kilograms of mercury was used in the USA each year in dentistry for the preparation of dental amalgam restorations (Mitra, 1986) In the early '90s, it was reported that the ingestion of elemental mercury (Hg) from dental amalgams caused renal damage in laboratory animals (Barkay, 1992) Kunkel et al. (1996) studied the fate of mercury in dental amalgams and determined that soluble mercury was never detected in a treatment plant headworks probably due to the fact that it became insoluble in the sludge floe. Mercury levels and discharge in wastewater from dental clinics in Denmark was investigated by Arenholt-Bindslev and Larsen (1996) Clinics without amalgam separators had a mean value of 270 mg Hg/dentist/day while those equipped with amalgam i

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separators had a mean value of 3 5 mg Hg/dentist/day. Several hundred grams of mercury/clinic may be discharged with wastewater each year. Wilhelm et al. (1996) studied biological monitoring of mercury vapor exposure in dental students by scalp hair analysis in comparison to blood and urine. It was concluded that hair may be used as an indicator of internal uptake of mercury provided that the hair was not externally exposed to mercury vapor. i Forms and Toxicity of Mercury I Mercury can be divided into three main categories These are elemental mercury (Hg) inorganic mercury (Hg^*, and its compounds) and organic mercury (for example methylmercury (CHjHg*) phenylmercury (CgHgHgO ethylmercury (CHjCHjHg*) These affect human health in one way or another. ^Metallic "' mercury and inorganic mercury compounds generally attack the liver and kidney, but they normally do not remain in the body long enough (24 hours for mercuric mercury) to accumulate to serious levels (Hammond, 1971).; The excretion of inorganic mercury occurs through the kidneys, liver (as bile) intestinal mucosa, sweat glands, and salivary glands. Inhalation has been shown to be the main route of

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exposure for mercury vapor. The lungs are able to absorb the metal with nearly 100 percent efficiency. Mercury vapor is considered hazardous at concentrations greater than 0.05 mg/m-^ Once mercury vapor gets into the bloodstream, it is oxidized to Hg-^* (Barkay, 1992) Methylmercury and mercury vapor pose the greatest threat to human health. ^ Methylmercury is the most toxic form of mercury. Microbes methylate inorganic mercury, generally under anaerobic conditions, to produce methylmercury (Mitra, 1986) Fish and shellfish provide the major source of intake by humans Methylmercury is rapidly absorbed from the gastrointestinal tract of humans and attacks the central nervous system. It is only slowly excreted. The main route of excretion in humans is via the feces, in which the rate is about ten times higher than that in urine. j Poisoning from methylmercury is characterized by sensory disorders, concentric constriction of the visual fields, impairment of hearing, symptoms from the autonomic nervous system, decreasing physical coordination, loss of memory as well as "mental disturbances" often referred to as Minamata disease. In the most severe case in adults, specific anatomical areas of the brain are damaged. The damage is

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10 irreversible because neuronal cells are destroyed (Fitzgerald and Clarkson, 1991) Pregnant women are spared from poisoning because the methylmercury quickly crosses the placental barrier to accumulate in their unborn children instead, thus causing teratogenesis and damage of their central nervous systems (D'itri and D'itri, 1977). •I Special precautions must be taken in the laboratory during the handling of mercury, especially the organic mercury forms. The death of the Dartmouth professor. Dr. Karen Wetterhahn, in 1997 highlights the fact that latex gloves do not adequately protect the skin from absorbing dimethylmercury The most notable methylmercury case on record occurred in Minamata, Japan during the 1950 's. Initially, there were deaths of aquatic and terrestrial organisms. Then the impoverished fishermen and their families who ate large amounts of fish from Minamata Bay developed the strange symptoms mentioned above. It was concluded that these symptoms were caused by people eating fish and shellfish contaminated by methylmercury, the source of which was Hg that was discharged in the wastewater from the Chisso plant. This plant manufactured acetaldehyde and the resulting waste which

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11 contained inorganic mercury had been continuously dumped into Minamata Bay since the 1930 's. In the sediment, the inorganic mercury was transformed to methylmercury and it eventually accumulated in the fish (Fujiki and Tajima, 1992; D'itri, 1991) In the biosphere, elemental mercury has a high vapor pressure (Henry's constant, H=0.3) and a low aqueous solubility of 6 x 10"^ g/100 ml water at 25C. Mercuric ion has a low vapor pressure (H < 10"^) and a high aqueous solubility of 7.4 g/lOO ml cold water (for HgClj) It forms covalent bonds, has an affinity for thiol groups, and tends to form strong bonds with inorganic and organic ligands. Organic mercury consists of highly stable C-Hg bonds, is very soluble in both water and hydrocarbons, and undergoes rapid transport through membranes The vapor pressure of methylmercury (CHjHg*) is quite low (H< 10 '} but the vapor pressure of dimethylmercury ((CH3)2Hg)) is high (H < 0.3) (Barkay, 1992). Mercury has three oxidation states: 0, +1, and +2. Cationic mercury binds tightly to iron (Fe) aluminum (Al) and manganese oxides (MnOj) hydroxides (OH) clay particles and silica. It also has a great affinity to sulfur (HgS) SHgroups, and chlorine (HgClj)

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12 Mercury in the Atmosphere In the atmosphere, mercury may exist in the gaseous phase, the aqueous phase, and the solid particulate phase. Components of the gaseous phase are elemental mercury (Hg) mercuric chloride (HgClj) mercuric hydroxide ((HglOH);), monoalkyl derivatives, such as methylmercuric chloride (CH3HgCl) and dialkyl derivatives, such as dimethylmercury ((CH3)2Hg). The aqueous phase typically includes HgCl^, Hg(0H)2, and sulfites (SO3) ; and in the solid particulate phase, there are mercuric oxide (HgO) and mercuric sulfide (HgS) (Seigneur et al 1994; Schroeder et al. 1991; Nriagu and Davidson, 1986) Concentrations of Hg species in the atmospheric environment are quite varied. Typical gas phase concentrations for Hg range from 2-5 ng/m^ and liquid phase concentrations range from 6-27 x 10"^ ng/L. In the case of inorganic mercury (Hg^*) the typical gas phase concentration is in the range 0.09-0.19 ng/m^ and the typical liquid phase concentration is in the range 3.5-13.3 ng/L (Seigneur et al. 1994) It is estimated that 95% of atmospheric Hg exists in the form Hg and has an atmospheric residence time of 0.7-2.0

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13 years (Gustin et al 1996; Petersen et al 1995; Nater and Grigal, 1992) Seigneur et al. (1994) have recently shown that global atmospheric budgets of Hg indicate a Hg half-life on the order of one year. J Slemr and Langer (1992) have found increases in atmospheric Hg concentrations from 1970 to 1990 of about 1.5% per year in the Northern Hemisphere and about 1.2% per year in the Southern Hemisphere. Swain et al. (1992) have determined approximately a 2% increase in Hg deposition rates in • J Minnesota and Wisconsin. Atmospheric mercury sources Natural sources of atmospheric Hg include windblown dust, volcanogenic particles, forest wildfires, and seasalt Nriagu i (1989) stated that volcanoes and fumaroles are responsible for I 50% of the Hg that is released naturally, and soil-derived dusts, forest fires, and seasalt sprays account for less than 10% of Hg released from natural sources. Natural degassing |; rates of Hg on regional or global scales have been estimated to range from 0.02 to 0.03 /^g/m^ h (Lindberg, 1986). Annual global emissions of mercury to the atmosphere have been reported as 0.4 x lO*" kg/yr from natural sources and 110 X 10*^ kg/yr from anthropogenic sources (Schroeder et al, 1992;

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5 II 14 Nriagu and Davidson, 1985) Major anthropogenic sources include fossil fuel combustion, particularly coal combustion, production of non-ferrous metals, refuse incineration (e.g. MSW incineration) and fuel wood combustion. Pacyna (1987) has implicated coal combustion as the largest single source of atmospheric Hg pollution, accounting for 58% of the anthropogenic input of Hg to the atmosphere. More than 90% of Hg in coal is released as Hg. Refuse incineration accounts for 33-40% of atmospheric mercuury pollution (Pirrone et al 1996) Atmospheric Species and Reactions Many chemical species that are involved in atmospheric Hg chemistry. These will be listed followed by their typical concentrations in brackets: Hg (0.25-0.6 ppt) ; Br2, HBr (2 0ppt) ; I2, HI (0.3 ppt); HCl (0.01-2.0 ppb) ; O3 (0.02-0.40 ppm) ; OH (0.04-0.4 ppt); NO. (<=1 ppb-0.5 ppm) ; SO2 (0.003-0.1 ppm) ; HjS (0.0-0.01 ppm); NjO (300 ppb); NH3 (0.015-100 ppb); H2O2 (
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15 such as O3, O2, CI2, H2S and H2O2 Table 2-1 shows gas phase reactions of Hg in the atmosphere (Seigneur et al 1994). Aqueous phase reactions occur in rainwater, fogwater, cloudwater, and other forms of precipitation. Table 2-2 shows aqueous phase reactions of Hg in the atmosphere. -./ Mercury in Aquatic Systems ,-Most aquatic system studies have focused on Hg in lakes (Angstrom et al. 1994; Lange et al 1993; Winfrey and Rudd, 1990; Steffan et al. 1988). Some authors have researched mercury in oceans and estuarine waters (Fabris et al. 1994; Langston, 1982; Mantoura et al 1978; Lindberg and Harris, 1974) Mercury in rivers has been investigated by a few authors (Ebinghaus and Wilken, 1993; Bubb et al. 1991; Mantoura et al. 1978) In freshwater, chlorides (Cl") and hydroxides (OH') are the main species existing, but sulfide (S^^) may also be present. Elemental Hg and dimethylmercury (DMHg) are oxidized in the atmosphere to water-soluble species, such as Hg^* and CH3Hg* which are then deposited onto a wetland (Fitzgerald and Clarkson, 1991) When soluble MMHg enters the aquatic system, it is quickly accumulated by most aquatic biota (Clarkson et

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16 Table 2-1. Gas Phase Reactions of Mercury Reaction Equilibrium or Rate Parameter (cm^/molecule s] (1) Hg ( 2 ) Hg (3) Hg { 4 ) Hg (5) Hg (6) Hg (7) Hg (8) 2Hg ( 9 ) Hg (10)Hg (ll)Hg (12)Hg g)+03(g) -->Hg(II) (g) g)+Cl2(g)-->HgCl2(g) g) +Br2 (g) -->HgBr2 (g) g)+l2{g) -->Hgl2(g) g)+H202(g) -->Hg(0H)2(g) g)+2N02(g) -->Hg(N02)2(s,g) g) +HC1 (g) -->products '(g) +02(g) -->2HgO(s,g) g) +SO2 (g) ->products g) +H2S (g) -->products g) +N2O (g) -->products g) +NH3 (g) >products 13)Hg(g)+2HI (g) -->Hgl2(g)+ Hj (g) ;i4) HgCl2 (g) -->products ;i5) Hg (OH) 2 (g) -->products <8.0 X 10"^5 <4.1 X 10"^^ <4.1 X 10"^^ <2.7 X 10-" <4.1 X 10"" 3 .3 X 10-35 1.0 X 10-" <1.0 X 10-" <6 X 10-1^ <6.0 X 10-^^ <2.0 X 10"^" <1.0 X 10-^^ 2.7 X 10-*i slow not available Adapted from Seigneur et aJ. (1994)

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17 Table 2-2. Aqueous Phase Reactions of MercuryEquilibrium or Reaction Rate Parameter (1) Hgj^" <-->Hg(aq) +Hg2" 2.9 x lO'^M (2) Hg(aq) +03(aq) -->Hg(II) (aq) + 4.7 x lO'^M-^s"" 02(aq) (3) Hg(aq) +2HClO(aq) +2H* + 2enot available -->Hg2* + 2Cl+ 2H20(l) (4) Hg (aq) +Peracetic acid or not available m-chloro-peroxybenzoic acid(aq) -->Hg2" or Hg* (5) Hg(aq) +H202(aq) -->HgO(s) + 6.0 M'^s"^ Hg^VHjOd) (6) Hg2^*+03(aq) -->Hg2* 9.5 x 10M-^s"' (7) Hg22*+H202(aq) -->Hg2^ <16 M-^s"^ (8) Hg(S03)2'--->Hg 1.0 x 10"* s'^ (9) HgS03(aq)+ S032-<>Hg (SO3) 2^' 2.5 x 10" M"^ (10)Hg=" + SO32"<-->HgSO3(aq) 5.0x10^^ M"^ (ll)HgS03(aq) -->Hg(aq)+S0420.6 s"^ (12)HgCl2 (aq) -->Hg2* + 2Clnot available (13)Hg(OH)2(aq) -->Hg(aq) not available (14)HgCl2(s)<-->HgCl2(aq) 0.27 M (15)HgCl2 (aq) <-->Hg2* + 2Cl10"^' M^ (16)HgCl2(aq) +2Cl-<-->HgCl470.8 M^^

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18 Table 2-2. Aqueous Phase Rea ctions of Mercury (continued) Equilibrium or Reaction Rate Parameter ;i7)Hg(OH)2(s)<-->Hg(OH)2(aq) 3.5 x 10"^ M :i8)Hg(OH)2(aq)<-->Hg2^ + 20H10"" M^ Adapted from Seigneur et al. (1994;

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19 al 1984). >;^Mercury bioaccumulates in aquatic organisms by three processes, namely: (1) from the water via respiration (e.g. over the gills), (2) by absorption of water from the body surface, and (3) by ingestion of food (D'ltri, 1991). The uptake of Hg through the aquatic food chain is an extremely important route of bioaccumulation. Mercury tends to bioaccumulate and biomagnify as it moves up a food chain. ^ Methylation Methylation of Hg is a major step in biogeochemical cycling. Sediments are a part of the aquatic system, and are an active site for Hg methylation, being the primary sink for Hg released into the environment. In the sediment, a variety of microbes transform the inorganic Hg into MMHg and DMHg. Various studies have determined specific microorganisms that are responsible for the methylation of Hg in the environment. Wood et al. (1968) were the first to show that extracts of methanogenic bacteria (strict anaerobes) methylated Hg .^. Other studies implicated the bacteria Neurospora. spp (Landner, 1971), and Clostridium cochlearium (Yamada and Tonomura, 1972) in the methylation process, as well as a number of gram-negative and gram positive cocci

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20 isolated from river sediment (Hamdy and Noyes, 1975) Vonk and Sijpestein (1973) isolated Pseudomonas and Bacillus from soil, as well as culture strains of Mycojbacterium, Escherichia coli, Bacillus megaterium, and fungi. Gilmour et al(1992) determined that sulfate-reducing bacteria were important in the methylation of Hg. Therefore, it can be stated that the bacteria which methylate Hg are facultative aerobes and anaerobes and sulfate-reducing bacteria. ''Mercury that is buried in deeper layers of sediment is not available for methylation unless it is disturbed. Physical occurrences, such as shifting bottom currents, plus small worms, can stir the mercury-rich layers to a depth of two centimeters. Larger freshwater organisms, such as mussels can stir up the sediment to a depth of at least nine centimeters (D'itri and D'itri, 1977). More MMHg is present in water nearest to surface layers of sediment where facultative microbes usually live. ... Three methylating coenzymes in biological systems are stated to participate in enzymatic methylation. These are (1) S-adenosylmethionine (2) N^ -methyl tetrahydrofolate derivatives, and (3) methylcobalamin (D'itri et al., 1978).

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21 Shapiro and Schlenk (1965) and Ridley et al. (1977) discovered that S-adenosylmethionine and N^-methyltetrahydrof olate could not transfer methyl groups to mercuric ions because they were only able to transfer a methyl group as Cll^% a carbonium ion. Methylcobalamin is able to transfer a methyl group to an inorganic mercuric ion. It can transfer groups as a carbanion (CH3-) and a methylradical (CH3) to produce MMHg and DMHg under aerobic and anaerobic conditions (Wood, 1974) When MMHg is taJcen up by fish, it moves to the red blood cells and then to fatty tissues where it may be retained for up to two years. Fish and shellfish provide the major source of intake by humans In humans, MMHg is rapidly absorbed by the gastrointestinal tract and attacks the central nervous system. When MMHg combines with tissue it is stable and is very slowly degraded or excreted from the body (Mitra, 1986; Barkay, 1992) The main route of excretion in humans is via the feces, in which the rate is about ten times that in the urine (Mitra, 1986) Studies with radioactive MMHg have demonstrated that it is retained in the human body with a half life of about 70 days. Thus, toxic amounts can be accumulated even with low dose rates (Hammond, 1971)

PAGE 36

22 Demethylation Methylmercury is slowly degraded because of the high stability of the carbon-mercury (C-Hg) bond. The low polarity coupled with low affinity of Hg to oxygen decreases the chances for hydrolytic cleavage of the strong bond. The means by which MHg is degraded is by the slow process of protolytic attack on the C-Hg bond. : Clarkson et al. (1984) reported that the demethylation of MHg in the water phase is carried out by a variety of microorganisms in two enzymically mediated states: (1) hydrolase enzymes cleave the C-Hg bond releasing the methyl group, and (2) reductase enzymes convert ionic Hg to Hg which can diffuse into the atmosphere. The equation below shows the transformation: hydrolase reductase CH3-Hg > CH4 + Hg2* > Hg /The enzyme, organomercurial lyase, which is produced by bacteria in sediments, soils, and waters, can speed up this reaction 10^ fold (Barkay, 1992) Thus, other authors have implicated a two step demethylation procedure involving cleaving of the C-Hg bond by organomercurial lyase to produce CH4 and ionic Hg, followed by reduction of the ionic Hg to Hg

PAGE 37

23 by the enzyme, mercuric reductase (Barkay, 1992; Oremland et al. 1991; Nakamura et al. 1990). The equation below shows this transformation: organomercurial mercuric lyase reductase CH3-Hg > CH4 + Hg2* > Hg= Furthermore, it was stated that Hg volatilization has been found to be chromosomally encoded in specific bacterial genera, namely Staphylococcus aureus and Bacillus spp (Nakamura et al. 1990) Lakes ./' The primary source of Hg in lakes is from atmospheric deposition. Atmospheric deposition to lakes occurs mainly as inorganic Hg, even though there are small amounts of MHg (Winfrey and Rudd, 1990) In aerobic waters, Hg^'' will complex with inorganic ligands, such as chlorides and hydroxides, bind with dissolved organic carbon (DOC) or attach to particulate matter. Ionic Hg can also be microbially reduced to form Hg. In anaerobic zones, ionic Hg can be converted to MHg or complex with sulfides and precipitate as mercuric sulfide (HgS) Particle-bound Hg reacts with sulfides and is

PAGE 38

24 converted to insoluble HgS which then settles out of the water column onto the sediment. In the presence of hydrogen sulfide (HjS) MHg will form methylmercuric sulfide (CH3Hg)2S which decomposes to HgS and DMHg {(CH3)2Hg) which tends to volatilize (Winfrey and Rudd, 1990; and Iverfeldt and Lindqvist, 1986) / Gilmour et al. (19 92) found the highest MHg concentrations near the sediment -water interface and in shallow sediments. Sulf ate-reduction rates were also highest at the sediment -water interface, thus confirming that sulfatereducing bacteria are important in the methylation of Hg y/ciarkson et al. (1984) stated that inorganic Hg methylation rates by microorganisms are pH-dependent with the greatest amount of MHg being formed at pH levels lower than 6 Higher pHs yield DMHg which is volatilized. Bloom et al. (1991) studied the impact of acidification on the MHg cycle of remote seepage lakes. Methylmercury concentrations were measured in the water of five pristine lakes which had a pH range of 4 6 to 7.2. In general, it was found that MHg in lake water tends to increase as pH decreases. Methylmercury partitioning was weakly related to pH. "^ ^ Miskimmin et al. (1992) studied the effects of pH on the Hg methylation and demethylation rates in lake water. The

PAGE 39

25 results showed that a reduction in pH from 7.0 to 5.0 yielded large increases in net methylation rates at both low and high dissolved organic carbon (DOC) concentrations. Rates of microbial activity, which were represented by rates of respiration, had the least effect on net MHg production rates in the pH range 5.0 to 7.0. Steffan et al. (1988) studied the effects of acidification on Hg methylation, demethylation, and volatilization in sediments from an acid susceptible lake. Sulfuric acid (HjSO^) hydrochloric acid (HCl) and nitric acid (HNO3) were used to acidify the sediment. Methylation was inhibited over 65% when H2SO4 and HCl were used to reduce the pH from 6.5 to 4.5, There was almost complete inhibition of methylation when HNO3 was used to bring the pH's to 5.5, 4.5, and 3.5. Demethylation was greatest at pH < 4.4, but was not affected by pH's between 4.4 and 8.0. Volatilization was less than 2% of methylation activity and was not significantly affected at the various pH levels. Winfrey and Rudd (1990) determined that decreased pH stimulates MHg production at the sediment -water interface and probably in the aerobic water column. It was also shown that decreased pH also decreases loss of volatile Hg from lake

PAGE 40

26 water and increases Hg binding to particulates in water. These factors enhance the bioavailability of Hg for methylation, therefore methylation rates may be increased. In anoxic subsurface sediments, pH decreases the methylation rates, implying that formation in the water column and at the sediment -water interface may be most significant in acidified lakes Xun et al. (1987) reported that in the pH range 4.5-8.5, there was an inverse relationship between Hg methylation and pH. Regnell (1994) did not find significant differences in methyl^"Hg in water at pH 5.8 and 6 6 when radiolabeled ^" HgClj was added to water overlying sediment in freshwater systems The report also confirmed the findings of other authors that anoxic conditions enhanced MHg production in water (Olson and Cooper, 1976) Mantoura et al. (1978) reported that in freshwater, more than 9 0% of Hg had been found to be complexed by humic materials Xu and Allard (1991) studied the effects of fulvic acid on the speciation and mobility of Hg in aqueous systems. It was determined that at pH levels below the point of zero charge, the humic substances adsorb and increase the uptake of

PAGE 41

27 trace metals from the solution phase. In the case of fulvic acid taking part in the adsorption of Hg on an oxide (alumina) the presence of the fulvic acid enhanced the Hg adsorption in the pH range of 2 5 9 5 In general, fulvic acid decreases Hg mobility under both acidic and basic conditions Mierle and Ingram (1991) investigated the role of humic substances in the mobilization of Hg from watersheds. Their results indicated that humic matter controls the solubility and watershed export of Hg deposited in precipitation. Minagawa and Takizawa (1980) determined very low levels of inorganic and organic Hg in natural waters by CVAAS after preconcentration on a chelating resin. In analyzing lake and river waters, it was found that 35-60% of the Hg present was in the form of organic compounds or in association with organic matter. Hakanson (1980) established that higher MMHg levels were found in fish from waters with low productivity, low pH, and high water and sediment load. The reduction of sewage sludge in a lake was shown to increase MMHg content of fish because bioproductivity is decreased and pH lowered. The Hg cycle and fish in the Adirondack lakes was investigated. There was an

PAGE 42

28 observed pattern of increasing total Hg and MMHg with decreasing pH. Another correlation found was that lakes with higher dissolved organic carbon (DOC) had higher concentrations of both total Hg and MMHg. There were also positive correlations between fish Hg and NMHg concentrations in the water column. In the case of wastewater, conventional methods for the removal of Hg^* include sulfide precipitation, ion exchange, alum and iron coagulation and adsorption on activated carbon. Namasivayam and Senthilkumar (1997) investigated the recycling of industrial solid waste for the removal of Hg^* by adsorption process. The report focused on the feasibility of Fe ( III ) /Cr ( III) hydroxide for the adsorption of Hg^* from aqueous solution. Terrestrial System Soil Mercury enters the soil via the disposal of sewage sludge, rainfall, dry fall-out from the atmosphere and the use of mercury-based pesticides. A more recent addition would be land application of compost to condition the soil or fertilize crops or vegetation grown. Compost is solid waste that has undergone biological decomposition of the biodegradable

PAGE 43

29 organic matter under controlled mesophilic and thermophilic temperature conditions, and has been stabilized to a degree which is potentially beneficial to plant growth. This stabilized material is used or sold for use as a soil amendment, artificial top soil, or other similar applications. The stabilized compost can easily and safely be stored, handled and used in an environmentally acceptable manner. Mercury in the soil occurs in several forms. These are (1) dissolved (free ion or soluble complex); (2) nonspecifically adsorbed (binding mostly due to electrostatic forces) ; (3) specifically adsorbed (strong binding due to covalent or coordinated forces) ; (4) chelated (bound to organic substances); and (5) precipitated (as sulfide, carbonate, hydroxide, phosphate, etc.) (Schuster, 1991) Adsorption of Hg in soils is dependent on the chemical form of Hg, soil pH, amount and chemical nature of inorganic and organic soil colloids, type of exchangeable cations and redox potential (Adriano, 1986) Coarse gravel has a lower capacity to bind Hg than finer soil materials, clay > silt > sand in the order of Hg complexation abilities. Farrah and Pickering (1978) investigated the uptake of Hg by three clay minerals and found

PAGE 44

30 illite > montmorillonite > kaolinite. Reimers and Krenkel (1974) found similar results in that illite and montmorillonite had faster Hg uptake than kaolinite. Under acidic conditions, ion-exchange is the most important form of sorption to clay minerals, while Hg adsorption is assumed to occur mainly in the form of Hg(0H)2 at higher pH values. yNumerous studies have reported that Hg shows a great affinity for organic matter in soils, peats, and sediments (Newton et al 1976; Wallace et al 1982; Schuster, 1991) The binding is strong, but reversible. At low Hg concentrations, soil organic matter is responsible for most of the Hg sorption. At higher concentrations, mineral components take part in the sorption. Results suggest that inorganic colloids contribute to the adsorption of organomercurials, whereas inorganic Hg compounds tend to bind strongly to soil organic matter. It was also stated that organic components were even more relevant in mercury adsorption at higher Hg concentrations. This is because organic matter has a larger adsorption capacity for Hg than mineral colloids (Schuster, 1991) Wilken (1992) stated that most of the water soluble Hg species in soil were complexed by organic material with a

PAGE 45

31 molecular weight greater than 500 daltons and particle sizes smaller than 1.2 fim. Mercury is bound to humic matter byexchangeable cat ionic and non-cationic sites (Aggarwal and Desai, 1980) About one-third of the total binding capacity of the soil humus is used for cation-exchange processes, and about two-thirds of the available binding sites serve for metal complexation. Johansson et, al. (1991) studied Hg in Swedish forest soils and waters and determined that the transport and distribution of Hg in forest soils is positively correlated to the transport of organic matter. The maximum sorption of Hg for several soil types occurs in the pH range 4.75-6.50 (Lodenius, 1990) In acidic soil (pH < 4.5 to 5) organic material is the only sorbent for inorganic Hg. In neutral soils, iron oxides and clay minerals play a part in Hg sorption. With decreasing pH, the mobility of Hg increases in soils that are low in organic matter (Schuster, 1991) Chlorides occur in all natural soil and water systems and it is regarded as one of the most mobile and persistent complexing agents for heavy metals (Schuster, 1991; Adriano, 1986) Barrow and Cox (1992) investigated the effects of pH

PAGE 46

I 32 and chloride concentrations on the sorption of Hg. They concluded that in the absence of chloride, there were only infinitesimal effects of pH on sorption between 4 and 6. Sorption decreased at higher pH. At low pH, the addition of chloride decreased sorption, but at high pH, chloride additions had little effect on sorption. Another study dealing with complex formation on Hg (II) adsorption by bentonite also found that chloride ions reduced Hg (II) adsorption, especially at low pH s Maximum adsorption occurred in the pH range 4.5 to 5.5, regardless of initial Hg concentration. The leaching of Hg from peat soils, such as those found in freshwater wetlands, decreases with decreasing pH (Lodenius, 1990) Lodenius et al. (1987) studied the sorption and leaching of Hg in peat soil using small additions of labeled Hg in peat lysimeters. Most of the added Hg was bound to the uppper most layer of the peat columns. Additions of chloride and sterilant did not affect the leaching of Hg. When the peat soil was allowed to dry completely, the leaching of Hg was greatest This may have been the result of Hg that was attached to suspended organic material passing quickly through the cracks that had formed in the soil. Also, greater

PAGE 47

33 amounts of artificial rainfall resulted in increasing amounts of Hg being leached out. Low organic matter acidic soils increase Hg mobility, and removal by leaching is more likely in acidic soils. Fang (1981) studied the sorption of Hg vapor in dry and moist soil columns. He found increased sorption with increasing soil moisture until a maximum was achieved near the maximal water holding capacity. Schnitzer and Kerndorff (1981) determined that humic substances tend to increase the solubility of Hg by the formation of water-soluble complexes. Comparing peat to sandy soils, the retention of mercury was much stronger and the volatilization smaller in peat than in sandy soils due to the tight bonding of Hg with organic matter (MacLean, 1974) In mineral soils, where Hg adsorption is dominated by iron oxides and clay minerals, there is promotion of evaporation as Hg Humic acid enhances the reduction of Hg^" to Hg and thereby increases volatilization losses to the atmosphere (Rogers, 1977; Lia et al., 1982). j Bacteria and other microorganisms play a major part in Hg volatilization. They are able to reduce Hg^* to Hg, which is volatile. They are also able to methylate Hg to DMHg, which is volatile, under alkaline conditions.

PAGE 48

34 Some bacteria adsorb Hg on the outside of their cell walls where it converts directly into a vapor. Escherichia coli absorb Hg into their systems, then the Hg^" may combine with the cytoplasm and convert into another compound or the volatile Hg. Pseudomonas aeruginosa proteus and at least two other microorganisms are known to convert Hg^* into Hg (D'itri and D' itri, 1977) /"''Mercury volatilization tends to increase with increasing temperature. Rogers and MacFarlane (1978) evaluated the volatilization rates of Hg in clay and sand and determined that the volatilization rate of Hg from clay was greater at higher temperatures, but less of the total Hg was volatilized. Increasing Hg concentration resulted in an increased volatilization of Hg from the sand and clay. Autoclaving bother soils drastically decreased the volatilization rates. After inoculation of the sterile soils with non-sterile soils, volatilization rates increased, thus indicating that it was microbially mediated. Another study found Hg to be translocated in soil horizons, and in some cases, lost from surface soil. This was attributed to evaporation to the atmosphere (Dudas and Pawluk, 1976)

PAGE 49

35 ^ Sediments Accumulation rates for Hg in sediments have been estimated at between 10 and 50 /^g Hg/m^ y (Andersson et al. 1990) Sediments are the primary sink for Hg released into the environment. Sediments consist of an oxidized zone and a reduced zone. In the oxidized zone, particulate Hg is desorbed to release Hg^-" and methylated biotically and abiotically to produce MMHg (Zhang and Planas, 1994) In the oxidized zone, Hg^* can be readsorbed to form particulatemercury. The MMHg and particulate-Hg can be deposited into the reduced zone of sediments. In this zone, Hg^* is reduced to Hg" and ultimately Hg which is then volatilized. Microbes aid in this reduction. In the reduced zone, ionic or Hg^* can react with sulfides that are present to form HgS Methylmercury in this zone also reacts with sulfides (S^") to form DMHg and HgS (Mitra, 1986) Mercury reactions that take place in the oxdized and reduced zones of sediment are given below; ;a) Oxidized zone reactions: (desorption) (1) particulate-Hg^" > Hg^" < (adsorption)

PAGE 50

36 (methylation) (2) Hg2^ > CH3Hg* (b) Reduced zone reactions: (reduction) (reduction) (1) Hg2^ > Hg* > Hg [2) Hg2" > HgS + S2(3) CHjHg* > (CH3)2Hg + HgS Methanogens along with sulfate reducers, have been implicated in the methylation of Hg (Adriano, 1986; Wood, 1972) Oremland et al. (1991) investigated the involvement of methanogens and sulfate reducers in oxidative demethylation Under anaerobic conditions, results with inhibitors showed partial involvement of both sulfate reducers and methanogens. Sulfate reducers dominate estuarine sediments. Products of anaerobic demethylation were mainly carbon dioxide (CO2) and methane (CH4) Methane was the only product resulting from aerobic demethylation in estuarine sediments, suggesting that the procedure for demethylation followed the organomercurial

PAGE 51

37 pathway. Final results of this study led to the conclusion that both aerobes and anaerobes demethylate Hg in sediments, but either group may dominate in a specific sediment type. ^''"An earlier study by Spanger et al. (1973) looked at MMHg and inorganic Hg in lake sediments. Bacterial isolates rapidly degraded MMHg to CH4 and Hg. Heaton and Laitinen (1974) investigated the electrochemical reduction of MMHg. This occurs in two one-electron steps with the first electron resulting in the formation of a methylmercuric radical on the electrode and the second electron resulting in the reduction of methylmercuric compound to CH4 and Hg. This also confirms the role played by methanogens in Hg methylation. Langston (1982) showed that humic and fulvic acids provided favorable binding sites that accounted for 4 to 32% of the total Hg measured in surface sediments. Evans et al. (1984) and Revis et al (1990) found that the majority of extractable Hg was in the humic-fulvic acid and organic-sulf ide fractions. Abiological methylation of Hg by humic acid has been confirmed in vitro (Nagase et. al. 1984) Rekolainen et al. (1986) reported on the effect of airborne Hg and peatland drainage on sediment Hg contents in some Finnish forest lakes The sediment Hg content correlated

PAGE 52

38 directly with the organic matter content of the sediment and the pH of the water. Di-Giulio and Ryan (1987) studied Hg in soils, sediments, and clams from a North Carolina peatland. After selective extractions of peat and sediment samples, it was concluded that the majority of Hg was associated with organic matter associated fractions, particularly humic/fulvic bound and organic-sulf ide bound. y An important effect of pH is to mobilize MMHg sorbed on sediments. A reduction in water pH shifts MMHg from the sediments to the water phase, regardless of the type of sediment (Miller and Akagi, 1979) -7 Duarte et al. (1991) studied Hg desorption from contaminated sediments Results showed that the amount of Hg desorbed from the sediment was inversely correlated with pH and ionic strength. The effect of pH on the leaching of Hg from contaminated sediments was greatest at pH values less than 7. Concerning ionic strength, the most Hg leaching occurred at ionic strength values less than 0.4 mol/dm^. In anaerobic environments, such as sediments, or under mildly reducing conditions, there is a tendency for Hg to be precipitated as the sulfide (HgS) Mercuric sulfide is very

PAGE 53

39 insoluble (log K^ = -50.02 at 50C) (lUPAC, 1982), but can be transformed under low Eh and high pH conditions to the soluble form, HgSj^^ or to Hg (Barkay, 1992;, Wilken, 1992; Schuster, 1991) The insolubility of HgS makes it resistant to methylation, but under aerobic conditions, the sulfur in HgS may be oxidized to sulfate, after which the Hg^* can undergo methylation (Adriano, 1986) Weber (1993) stated that sulf ate-reducing bacteria contribute to MMHg production. A low sulfate concentration was needed for sulfate reducing bacteria to produce MMHg. Kerry et al. (1991) determined that methylation of Hg in sediments of an acid stressed lake was mainly due to the activity of sulfate reducing bacteria. The methylation rate did not correlate with the concentration of sulfate in the system. They speculated that formation of insoluble HgS reduced Hg availability, thus reducing MMHg production. J Revis et al. (1991) investigated the immobilization of Hg in soil, sediment, sludge, and water by sulf ate-reducing bacteria. It was determined that the addition of calcium sulfate (CaS04) which slowly moves to the sulfur-reducing bacteria, producing HjS and eventually producing HgS,

PAGE 54

40 prevented the transformation of Hg to organic methylated forms Gilmour et al. (1992) reported that additions of sulfate to anoxic sediments yielded an increased microbial production of MMHg from added inorganic Hg There were positive correlations between sediment depth profiles of bacterial sulfate reduction and Hg methylation. It was also noted that specific inhibition of sulf ate-reducing bacteria blocked MMHg production at all depths in the sediment profile. Sulf atereducing rates and MMHg concentrations were highest near the sediment-water interface and in shallow sediments. These results show that sulf ate-reducing bacteria are definite mediators of Hg methylation. Munthe et al. (1991) investigated the aqueous reduction of Hg^* by sulfite. This involved the formation of an unstable intermediate HgSOj, which underwent decomposition to produce Hg*, which in turn was reduced to Hg. The concentration of sulfite inversely determined the overall rate of the reaction. Analytical Methods Atmosphere Lindberg et al. (1992) investigated atmosphere exchange of Hg in a forest. In this report, Hg was collected in a

PAGE 55

41 series of gold traps, wet digested and then analyzed by longpath flameless atomic absorption. The detection limit for Hg in air ranged from 0.01 to 0.5 ng/m^ Licata et al. (1994) reviewed the testing of Hg in flue gases both in the USA and in Germany. In the USA, the most common method of analysis was CVAAS with SnClj in a HCl solution as the reducing agent. In Germany, the most common method of analysis was also CVAAS but with NaBH4 as the reducing agent. Prestbo and Bloom (1995) investigated a Hg speciation adsorption (MESA) method for combustion flue gas. The sampling system for gas phase Hg species utilized a series of heated, solid phase adsorbent traps. Mercury (II) and monomethylmercury (MMHg) were the flue gas oxidized species that were adsorbed by a potassium chloride (KCl) impregnated soda lime sorbent An iodated carbon sorbent was used to collect Hg after passing through the KCl/soda lime sorbent. CVAFS was used for final determination of Hg from these sorbents Chu and Porcella (1995) investigated Hg stack emissions from US electric utility power plants. It was determined that Hg emissions from total electric utilities was on the order of

PAGE 56

42 40 tons/yr. It was also determined that Hg emissions were not consistently captured by conventional air pollution control technologies including fabric filters, electrostatic precipitators, and flue gas desulfurization systems. The Hg in the emissions was analyzed by instrument neutron activation analysis (INAA) The MESA method was also used in this study, but neither this method nor the INAA was validated for Hg speciation. Balogh and Liang (1995) investigated Hg pathways in municipal wastewater treatment plants that used incineration to dispose of the solid material lodated carbon traps were used to collect Hg in the incinerator exhaust gas. Subsequently, the traps were digested with acids and final analysis was done by CVAFS Keeler et al. (1995) studied particulate Hg in the atmosphere. The analytical technique performed on Hg(p) extracted from glass fiber and other types of filters from glass fiber dual -amalgamation preconcentration and CVAFS detection. Teflon filters have been detected by instrumental neutron activation analysis (INAA)

PAGE 57

43 Aquatic Systems Minagawa and Takizawa (198 0) determined very low levels of inorganic and organic Hg in natural waters by CVAAS after preconcentration on a chelating resin. A column of dithiocarbamate-treated resin was used to simultaneously collect inorganic and organic Hg at ng/L concentrations and subsequently quantitatively eluted with slightly acidic thiourea solution. Alkaline SnCl2 solution is used to reduce inorganic Hg to Hg vapor. Mercury vapor was generated from inorganic and organic Hg with a CdCl2-SnCl2 solution. The range of determination was 0.2-5,000 ppt for 20-L water samples Determination of inorganic and organic Hg compounds by high performance liquid chromatography (HPLC) -inductively coupled plasma (ICP) emission spectrometry with cold vapor generation was investigated by Krull and Bushee (1986) It was not necessary to derivatize samples before analysis by HPLC. The conventional propylene spray chamber of the ICP was replaced by an all glass chamber. Detection limits ranged from 32 to 62 ppb of Hg for four Hg compounds. This represented three to four orders of magnitude enhancement over detection limits without cold vapor generation.

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44 Xiankun ^ ^. (1990) investigated the effect of the Huanghe river runoff on the occurrence, transportation, and speciation of Hg in the Huanghe Estuary and the adjacent sea. Mercury content in 15 mL filtered water samples was determined using two stage gold amalgamation coupled with CVAAS. Lee and Iverfeldt (1991) measured Hg in run-off, lake and rain waters using an oxidative treatment with BrCl prior to reduction by SnClj A gold trap was used to preconcentrate the Hg after its volatilization from a quartz glass reduction vessel. Nitrogen gas was used for the purging. A double amalgamation helium dc-plasma atomic emission method was used for analysis of the gold traps. Baeyens and Belgium (1992) determined total Hg by acidification of the sample to pH 1, pretreatment with a strong oxidant, BrCl, followed by reduction of the BrCl with NHjOH-HCl prior to SnClj reduction, air purging of Hg and collection on a Au-column. Atomic fluorescence was used for analysis of these samples. Smith (1993) determined natural levels of Hg in water samples by preconcentration onto gold traps followed by electrothermal heating and purging of the traps with argon

PAGE 59

45 directly into the torch of the ICPMS. The detection limit was 0.2 ng/L using a 2 00 mL sample. Garcia et al. (1994) determined inorganic and MMHg in sea-water by an online preconcentration, solvent extraction and total Hg determination by CVAAS A detection limit of 16 ng/L of Hg was observed for sample volumes of 2 5 mL. Bloom et al (1995) reported on the results of the international aqueous Hg speciation intercomparison exercise. Twenty three labs were involved with 18 utilizing BrCl oxidation, gold trapping and CVAFS for total Hg. Four other labs used CVAAS or wet chemistry and the results were similar. Sixteen labs provided MMHg results, but 15 of them used various combinations of aqueous phase ethylation, GC separation, and CVAFS detection. Emteborg et al. (1995) used a dithiocarbamate resin for the sampling and determination of Hg species in humic-rich natural waters. Filtration was used to collect the dithiocarbamate resin and transferred to a column which was placed into a closed flow injection system where acidic thiourea solution was used to elute the Hg species. The separation was performed using a gas chromatograph equipped with a non-polar capillary column and detection utilized

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46 atomic emission spectrometry at 253.7 nm after excitation in a microwave -induced helium plasma. Paquette and Helz (1995) reported on the solubility of cinnabar (Red HgS) and implications for Hg speciation in sulfidic waters. In this paper, the Hg analysis was performed by CVAAS on a homemade stannous chloride reduction, gold amalgamation apparatus attached to a Perkin-Elmer Model 2380 Spectrophotometer. A detection limit of 3 ppb Hg was noted. Saouter et al. (1995) developed and field-validated a microcosm to simulate the Hg in a contaminated pond. The water samples were treated and preserved with BrCl and total Hg determination was accomplished using CVAFS Specific analytical techniques for organic Hg in aqueous samples have also been investigated. One of the earlier papers by Fitzgerald and Lyons (1973) dealt with trapping the Hg after its reduction while purging with nitrogen gas and concentrating it on a packed column immersed in a liquid nitrogen bath. After the completion of purging, the column was removed from the cold trap, heated, and the gas phase concentration of the eluted Hg was measured on a Coleman Hg analyzer (MAS-50) A detection limit of 0.0017 //g/L was obtained.

PAGE 61

47 Schintu ^ al. (1987) developed a practical isolation technique for MMHg in natural waters Mercury compounds were extracted quantitatively from six different sources of water with 5 mL of a 50 ppm dithizone-chlorof orm solution. This method provided a high recovery for both organic as well as inorganic Hg from an aqueous medium, prior to their determination by gold trap CVAAS Bloom and Watras (1989) developed a technique to quantify MMHg and DMHg in several rainfall and lake water samples by aqueous phase ethylation to the volatile dialkyl analogs followed by cryogenic gas chromatographic separation. Mercury-specific detection by CVAFS provided a 0.1 pg as Hg limit Lansens et al. (1990) determined MMHg in natural waters by headspace (HS) gas chromatography (GC) with microwave induced plasma (MIP) detection after preconcentration on a resin containing dithiocarbamate groups. The chelating resin, Sumilate Q-10, showed a high affinity for both organic and inorganic Hg Lee and Hultberg (1989) determined MMHg in some Swedish surface waters. Methylmercury was preconcentrated from 10-20 L of water on a sulfhydryl cotton fiber (SCF) adsorbent.

PAGE 62

^1 i 48 packed in a column and eluted with a small volume of 2M hydrochloric acid (HCl) The eluate was extracted with benzene and then analyzed for MMHg with the GC/ECD method. Rapsomanikis and Craig (1991) investigated the speciation of Hg and MHg compounds in aqueous samples by gas chromatographyatomic absorption spectrometry after ethylation with sodium tetraethylborate The absolute detection limit for CHsHgCl was 167 pg. Sarzanini et al. (1992) simultaneously determined methyl, phenyl-, ethyl-, and inorganic Hg by CVAAS with on-line chromatographic separation. Reversed-phase chromatography on an ODS column and elution with an acetonitrile -water-ammonium tetramethylenedithiocarbamate buffered mixture was investigated with or without ammonium tetramethylenedithiocarbamate precomplexation. The detection limits for methyl, ethyl, phenyl, and inorganic Hg were 10 ng/mL, 50 ng/mL, 300 ng/mL, and 8 ng/mL respectively with direct injection (100 /uh sample) For the on-line preconcentration procedure, the detection limits were 0.5 ng/mL, 0.09 ng/mL, 0.5 ng/mL, and 0.015 ng/mL respectively in a 10 mL sample.

PAGE 63

49 Quevauviller et al. (1992) studied the occurrence of methylated tin and DMHg compounds in a mangrove core from Sepetiba Bay, Brazil. Mercury compounds were determined by derivatization with NaBH4, cryogenic trapping in a chromatographic column and using an electrothermally heated quartz furnace by AA using an EDL source for final detection. Gilmour and Bloom (1995) reported on a case study of Hg and MHg dynamics in a Hg-contaminated municipal wastewater treatment plant Methylmercury was determined after distillation by ethylation, isothermal GC and CVAF detection. Filippelli et al. (1992) investigated MMHg determination as volatile MHg hydride by purge and trap gas chromatography in line with fourier transform infrared spectroscopy. The detection limit of this method was 0.15 /ug. Table 2-3 provides a summary of the various analytical methods that the different authors used for the determination of mercury in aqueous samples Terrestrial System Revis et al. (1990) determined MMHg in soil by developing a method for assessing acceptable limits. Water, acid, copper sulfate (CUSO4) sodium bromide (NaBr) and toluene were all added to a soil sample. The toluene phase was collected and

PAGE 64

50 Table 2-3 An overview of the analytical methods for mercury in aqueous samples. Analytical Methods Species Sources Gold Trap CVAAS GC-HS-MIP GC-ECD He-dc plasma AED GC-QFAAS Gold Trap CVAFS Hg(Tot) Xiankun et al. (1990) CHsHg Lansens et al. (1990) CHsHg Lee & Hultberg (19 90) Hg(Tot) Lee & Iverf eldt (1991) CHsHg Rapsomanikis & Craig (1991) Hg(Tot) Baeyens (1992) GCCVAAS CHjHg Sarzanini et al. (1992) Gold Trap with ICPMS Hg(Tot) Smith (1993) CVAAS Hg(Tot), CH3Hg Garcia et al. (1994 CVAFS Hg(Tot) Saouter et ad. (1995) GC-CVAFS Hg(II), CHjHg Gilmour & Bloom(1995) Gold Trap CVAAS Hg(Tot) Paquette & Helz (1995) GC = gas chromatography CVAAS = cold vapor atomic absorption spectrometry CVAFS = cold vapor atomic fluoresence spectrometry GC-HS-MIP = headspace gas chromatography with microwave induced plasma detection QFAAS = quartz furnace atomic absorption spectrometry ICPMS inductively coupled plasma mass spectrometry AED = atomic emission detector ECD = electron capture detector

PAGE 65

51 mixed with an equal volume of sodium thiosulfate (NajSjOj) in ethanol Potassium iodide (KI) and benzene were eventuallyadded to the previous mixture. The benzene phase, which contains the MMHg was then analyzed on the GC. The detection limit for MMHg was 3 ppb. Hempel et al. (1992) determined organic Hg species in soils by high-performance liquid chromatography (HPLC) with ultraviolet (UV) detection. The simultaneous separation and quantification of nine organic Hg compounds was performed. This separation was done on octadecylsilane columns by gradient elution with a methanol -water mixture ranging from 3 to 50% v/v. The detection limits for the various compounds were in the range 70-95.1 /^g/dm^ Evans et al. (1984) determined organic Hg in peat, sediment, and biological samples. The sediment aspect involved the use of acetone as an extractant, KBr and CUSO4 in sulfuric acid and finally toluene as the last extractant. Mercury analysis was done using a graphite furnace AAS with the detection limit for MMHg set at 25 ng/g Mikac and Picer (1985) investigated Hg distribution in a polluted marine area. The methylmercury determination in the sediment samples involved HCl hydrolysis of the wet sample and

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I 52 extraction of CHjHgCl into benzene. This extract was then dried with Na2S04 and analyzed by GC/ECD. Sakamoto et al. (1992) reported on the differential determination of organic Hg, Hg (II) oxide, and Hg (II) sulfide in sediments by CVAAS All of these forms were removed from sediment by extracting with a solvent such as chloroform followed by a thiosulfate addition and subsequent analysis by CVAAS after digestion. Engstrom £t al. (1994) studied atmospheric Hg deposition to lakes and watersheds Sediment cores were obtained from various systems and the samples were digested with a strong acid-permanganate-persulf ate digestion technique. CVAAS was used for the total Hg analysis of these samples. Zhang and Planas (1994) reported on biotic and abiotic Hg methylation and demethylation in sediments. Methylmercury was extracted from sediment samples using CuS04/NaBr/H2S04 and toluene and analysis was performed by GC. Rood et al. (1995) investigated Hg accumulation trends in Florida Everglades and Savannas Marsh flooded soils. The Hg in the soil and sediment samples underwent a strong digestion using an acid-permanganate-persulf ate combination with final analysis by CVAAS.

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53 Municipal Solid Waste Municipal solid waste consists of community refuse, including garbage, rubbish and trash (Nathanson, 1986) Average composition of the MSW stream in the USA includes paper and paperboard (41.0%), glass (8.2%), metals (8.7%), plastics (6.5%), rubber and leather (8.1%), food wastes (7.9%), yard wastes (17.9%) and other miscellaneous materials (1.6%) (Tchobanoglous, 1993; Concern, Inc., 1988). Predictions of increases in MSW (MSW) up to the year 2000 indicate increases in paper and paperboard, rubber and leather goods, plastics, yard waste, and metals. There are decreasing trends for food waste and glass (Tchobanoglous, 1993; Concern, Inc., 1988). In 1987-1988, 80% of MSW generated in the USA was landfilled, 10% was incinerated and 10% was recycled. Other amounts were dumped illegally on land and in the ocean. MSW is regulated under Subtitle D of the Resource Conservation and Recovery Act (RCRA) in 4 CFR Part 257 (USEPA, 1988) In Florida, for the 12 month period ending June 30, 1993, 21.5 million tons of MSW were generated (up from 20.3 million tons in 1992). Of that amount, 9.9 million tons were landfilled (down from 10.4 million tons in 1992) 5.6 million tons were recycled (up from 5.4 million tons in 1992) and 5

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54 million tons were combusted (up from 4.5 million tons in 1992) ( Florida Department of Commerce, 1994) This is an excellent indication, at least within the State of Florida, that landfilling is decreasing and recycling is on the rise. The MSW stream contains hazardous household and commercial toxic products of which heavy metals are included. The presence and mobility of heavy metals in MSW have been a source of concern for quite some time. MSW is landfilled, incinerated, or composted and the environmental issues with heavy metals focus on groundwater contamination, air pollution, and uptake by crops and other animals, and eventually humans. Hazardous wastes are a subset of MSW and can cause serious illness, injury, or death. They also present a problem to the environment if improperly transported or disposed (Nathanson, 1986) Heavy metals can be classified as hazardous wastes due to their various characteristics as stated in 40 CFR 261 subpart C which lists the four characteristics of hazardous waste as ignitability, corrosivity, reactivity, and extraction procedures (EP) toxicity. On the other hand, there are many items that are excluded from the definition of hazardous wastes in EPA regulations, of which the first item listed is household

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55 waste. Many heavy metals found in MSW are actually products of household waste (Congress of the United States, 1983). Types of Heavy Metals Many different types of heavy metals are found in the MSW stream. It is important to note that aluminum and tin are not considered when referring to heavy metals. Included in the classification of heavy metals that exist are antimony (Sb) arsenic (As) cadmium (Cd) copper (Cu) chromium (Cr) iron (Fe) lead (Pb) manganese (Mn) nickel (Ni) zinc (Zn) mercury (Hg) selenium (Se) silver (Ag) and tin (Sn) Maximum concentration levels (in mg/1) as determined with the TCLP method for some of the heavy metals are 5.0 for arsenic, 100.0 for barium, 1.0 for cadmium, 5.0 for chromium, 5.0 for lead, 0.2 for mercury, 1.0 for selenium, and 5.0 for silver (Congress of the United States, 1983) Cadmium, copper, nickel, zinc and molybdenum have been identified as the heavy metals with the greatest potential to accumulate in plants. Cadmium can easily accumulate in crops at concentrations which could increase the dietary intake of the metal without causing crop phytotoxicity Lead, mercury, arsenic and selenium are not considered to be much of a threat to plants because they have low solubility in slightly acid or

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56 neutral, well-aerated soils (Louisiana Cooperative Extension Service, 1991) Sources The major sources of heavy metals in MSW are lead acid batteries, household batteries (auto and flashlight) consumer electronics, plastics, used motor oil and transmission fluid, magazine ink, stain, varnish, and sealant (Bretz, 1990; USEPA, 1988). Fluorescent lamps, high pressure sodium lamps, and high intensity discharge lamps contain mercury. To date, the lamps still exceed the regulatory level of 0.2 mg/L for mercury and also exceed the regulatory level of 5 mg/L for lead when TCLP are performed (USACHPPM, 1997) Outdated light bulbs contain around 20 to 80 milligrams of liquid mercury. The breaking of one bulb can poison 50 cubic meters of breathing air (Kovalov, 1994). Household batteries are major contributors of heavy metals in the solid waste stream. Each year, over two billion batteries are deposited in SW facilities in the United States (USEPA, 1997) Therefore, in a city of 500,000 people, almost 9,000 pounds of Hg would potentially enter the air, earth, and water (Jade Mountain, Inc., 1997). The different types of battery systems may contain Hg, cadmium, nickel, zinc,

PAGE 71

57 manganese, and lithium. Although each of these metals have negative health and environmental effects, Hg and cadmium are of greatest concern. It has been estimated that up to half of the Hg that is utilized in the United States is used in batteries. Batteries account for about 54% of cadmium in the solid waste stream. Fluorescent lights, Hgvapor lamps, arc lamps, light switches, mirror coating, amalgam, and electrical apparatus are other sources of Hg in MSW, but to a much lesser degree than that of batteries (Bureau of Solid and Hazardous Waste, 1995; Tchobanoglous 1993; Steinwachs, 1990) Table 24 identifies the discarding of Hg containing products into the MSW stream. Waste Recovery and Resource Treatment Combustion Combustion is defined as the chemical reaction of oxygen with organic materials, to produce oxidized compounds accompanied by the emission of light and rapid generation of heat (Tchobanoglous, 1993) MSW combustion (incineration) has two main functions, (1) reduction in the volume of waste subject to final disposal, and (2) recovery of energy. Combustion facilities are referred to as waste-to-energy (WTE ) units which are capable of producing steam and electricity

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58 Table 2-4. Discard of Hg Containing Products into the MSW Stream (tons) Product United States Florida 1989 2000 1989 2000 Household Batteries 621.10 98.50 32.30 5.12 Mercury Light Switches 0.40 1.90 0.02 0.10 Electric Lighting 26.70 40.90 1.39 2.13 Paint Residues 18.20 0.50 0.95 0.03 Fever Thermometers 16.30 16.80 0.85 0.87 Thermostats 11.20 10.30 0.58 0.54 Pigments 10.00 1.50 0.52 0.08 Dental Uses 4.00 2.30 0.21 0.12 Special Paper Coating 1.00 0.00 0.05 0.00 Adapted from Bureau of Solid and Hazardous Waste, 1995.

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59 and can be used in conjunction with source reduction, recycling and composting programs (USEPA, 198 9) For MSW, excess air is used to ensure complete combustion of the organic fraction. This is represented by the equation: Organic matter + excess air ->• Nj + COj + Hj + O2 + ash + heat The endproducts of combustion include hot combustion gases and noncombustible residues. The heavy metals are normally found in the noncombustible residue known as ash. Heavy metals emissions from municipal waste combustion (MWC) particularly Hg, are cause for concern because significant amounts of Hg are released through incinerator stack emissions. Mercury can vaporize at very low temperatures (40-50C) therefore it is easily emitted into the environment after which it can be converted into various forms, one of which is methylmercury which bioaccumulates in the food chain (Steinwachs, 1990) In 1986, EPA issued operational guidance on control technology for MWC. The final regulations were issued in December 1990 to reduce deleterious emissions. Batteries are supposed to be separated from MSW prior to incineration to reduce the heavy metal emissions. In the past, dead batteries in the waste

PAGE 74

60 stream were the principal source of heavy metals in the fluegas stream (Bretz, 1990). Metzger and Braun (1987) investigated in-situ Hg speciation in flue gas by liquid and solid sorption systems. Solid sorbents included iodated activated carbon and also gold or silver amalgams. Liquid sorbents included nitric acid/peroxydisulf ate solutions. A condensation/absorption approach was used to identify Hg within the sorbents. In Europe, Hg in gas emissions from waste incinerators consists of 20% of elemental Hg, 60% of divalent inorganic Hg compounds, and 20% of particulate Hg (Petersen et al. 1995) In Europe, the Hg content in MSW ranged from 0.3 to 9 g/t. In North America, the Hg content ranged from 0.3 6 to 5.8 g/t. Medical waste incinerators emit an average of about 30 g Hg/t in developed countries and 10-20 g Hg/t in developing countries (Pirrone et al. 1996) Krishnan et al. (1997) investigated Hg control in MWCs and coal-fired utilities. The report showed the ability of activated carbons PC-100 and FGD to capture Hg and HgClj emitted from these two systems. The precursor for PC-100 is bituminous coal and for FGD, it is lignite. The PC-100

PAGE 75

61 captured more Hg than FGD at both temperatures for MWC and coal-fired simulations. Carroll et al. (1995) investigated Hg emissions from a hazardous waste incinerator equipped with a State-of -the-Art wet scrubber. Scrubber collection efficiency for Hg averaged 87% which was lower than expected. Gowin et al. (1993) studied the ability of triple-reverse-burned coal char (TRB char) to remove Hg vapor from the gasification reactor and determined that it was an effective sorbent for this purpose. The main problem with MWC stems from the production of residual ash. The inorganic, noncombustible portion of the waste stream and the uncombusted organic matter are the constituents in the ash. Heavy metals remaining in incinerator ash are in a more leachable form. Exposure to these heavy metals occur in two ways. First of all, respirable particles may disperse into the air from the stack or during transport and handling. Next, leaching of the metals may occur at disposal, contaminating ground and surface water (Bretz, 1990) Fly ash and bottom ash often contain high levels of heavy metals that require special handling and burial in ash landfills. Standards for the handling, processing, disposal and recycling of MSW combustor ash in

PAGE 76

62 Florida are found in Chapter 17-702, F.A.C., and were adopted by the Environmental Regulation Commission in June, 1990 (Department of Environmental Regulation, 1991) Heavy metal concentrations and percentage distributions in soil samples confirm a strong tendency of heavy metals to accumulate in solid effluents and concentrate in fly-ash. This is one possible explanation of why there is a high amount of metals emitted per ton of burned waste (Morselli et ai. 1992) In the past, ash was just disposed of in a landfill specifically created for ash disposal called a monofill, but ash reuse is currently on the rise. Ash recycling involves the use of ash in soil cement for road sub-base material, the use of ash as aggregate in road asphalt, and the construction of blocks from the ash for artificial reef construction (Department of Environmental Regulation, 1991) Composting (Aerobic) In MSW composting, preprocessing is performed to isolate the compostable portion, i.e. yard wastes, food wastes, and organic fractions such as paper. These materials constitute about 30 to 60 percent of the MSW stream. Separation of the compostable portion is generally performed using a rotating

PAGE 77

63 screen called a trommel. Once these are separated, they are usually shredded to reduce the particle size and moisture may be added to aid the composting process (USEPA, 1989) The main requirement for compost is that it should be suitable for agricultural use as an organic soil conditioner. Therefore, physical, chemical and biological stability, nonphytotoxicity and balance among mineral elements are the primary characteristics for compost to be useful to the soil and for crops (de Bertoldi et al 1990) Heavy metals do not degrade, but tend to be concentrated during the composting process. Metals of greatest importance are those which bioaccumulate, resulting in long or short-term toxic effects to organisms in the environment. Those most commonly regulated include Cd, Hg, Pb, Ni and Zn. To date, most MSW composts have achieved regulatory limits for most metals, with the exception of lead, but there is an interest in attaining lower levels. The earlier that sorting occurs during the collection and composting process, the lower the heavy metal content in the finished compost. Source separation of a few contaminants such as lead-acid batteries and television picture tubes will definitely decrease the

PAGE 78

64 heavy metal content of MSW compost (Richard and Woodbury, 1994) One report determining how composting affects heavy metal content stated that there was a great occurrence of water soluble heavy metals, except for Cd. The concentrations of heavy metals were generally lower in the samples taken at the end of the composting period in the correctly prepared compost. In correctly produced compost maturation increases the humic acids with respect to the fulvic acids which does not occur in incorrectly produced compost. This is an indication of the process of humification of organic matter leading to the unavailability of heavy metals as the heavy metals would now be complexed to the humic material, hence reaching the soil in a complexed, less mobile form. Therefore, there would be a decrease in the leaching of heavy metals. About one-third of the total content of heavy metals present in compost was reportedly bound to the alkali soluble organic matter. (Canarutto et aJ. 1991) Heavy metal concentrations in the finished compost determine its usage. Lower heavy metal content allows the compost to be used for practical applications, such as landscaping. Higher metal concentrations in compost prohibit

PAGE 79

65 its usage because of the potential for plant uptake, or leaching into the environment (Steinwachs, 199 0) Land application rate of heavy metal containing composts are dependent on the concentration of the metal, the pH of the soil and the cation exchange capacity of the soil. The decrease in concentration of metals in the soil compost mixtures results from leaching, with very little as a result of plant uptake. In aerobic biologically stabilized solid waste residues, one of the main forms of Hg will likely be mercuric oxide (HgO) along with other possible forms of Hg such as HgClj which is likely to occur in the absence of sulfides. Chloride is normally a weak complexing agent, but it strongly complexes with Hg'* and Hgz^" (Pohland et al., 1981). HgO may be dissolved and mobilized by acid rain percolating through residue -amended soil, increasing the chances of Hg being taken up by plants, as well as the possibility of leaching. Landf illing In anaerobic environments, such as sediments or landfills, or under mildly reducing conditions, there is a tendency for Hg to be precipitated as the sulfide (HgS) Mercuric sulfide is highly insoluble (log K^ = -50.02) (lUPAC,

PAGE 80

66 1982) but can be transformed under low Eh and high pH conditions to the soluble form, HgSj^*, or to the free metal (Barkay, 1992;, Wilken, 1992; Schuster, 1991). The insolubility of HgS makes it resistant to methylation, but under aerobic conditions, HgS may be oxidized to the sulfate form which can then undergo methylation (Adriano, 1986) Gases found in landfills include methane (CH4) carbon dioxide (COj) ammonia (NH3) carbon monoxide (CO) hydrogen sulfide (H2S) nitrogen (N2) and oxygen (O2) (Tchobanoglous, 1993) When considering the movement of Hg in the landfill and the transformations that occur, it can be hypothesized that dimethylmercury and elemental Hg may actually be emitted into the atmosphere. Anaerobic bacteria may act on the available Hg in the landfill and transform it to either monomethylmercury or dimethylmercury. The vapor pressure of monomethylmercury is quite low (H < 10"^) while the vapor pressure of dimethylmercury is quite high (H < 0.3) (Barkay, 1992) As a result, this form of Hg may also be given off from a landfill in the same manner as methane or carbon dioxide which are the two most prevalent gases resulting from the anaerobic processes occurring in a landfill.

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67 Under warm or hot conditions, Hg can be transformed to elemental Hg which is the second most volatile form of Hg (H=0.3), next to dimethylmercury If this transformation does in fact occur in a landfill, then without proper control mechanisms, elemental Hg would be emitted into the atmosphere. Heavy metals dissolved in aqueous systems exist in combination with other chemical species in the form of complexes. Metal ions, such as Hg^% combine with non-metallic compounds known as ligands by means of coordinate -covalent bonds. In anaerobic residues, the species of Hg found will likely be mercuric sulfide (HgS) Cinnabar or mercuric sulfide is about the most stable Hg mineral that occurs in nature. This Hg mineral forms under reducing conditions where sulfate (SO42-) has been reduced to sulfide (S^") (Mitra, 1986) Gambrell et al. (1978) showed that HgS is unavailable to plants. Therefore it may be feasible to use anaerobically stabilized solid waste residues for land application. If Hg is in the HgS form, there should be minor, if any, adverse impacts on crops, and the groundwater should remain uncontaminated as HgS is quite immobile and unlikely to leach. The dumping of trash is becoming a less viable solution since the siting of landfills is extremely difficult as a

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68 result of public opposition. The public invokes the NIMBY (not in my back yard) syndrome due to the odors and the potential contamination to surface and groundwater that may occur due to leaching. However, as technology improves and liners become more reliable, landfills are meeting federal and state standards, which should actually ease the public's minds when it comes to the siting of landfills. Some communities that are faced with diminishing landfill space send their waste to other states, a practice sanctioned by the Supreme Court ruling in 1978 (Philadelphia v. New Jersey) that a state may not refuse waste from another state. This poses a problem for those communities that are trying to extend the life of their landfills by recycling or reducing their own wastes when the space is filled by other counties or states (Concern, Inc., 1988). The Solid Waste Management Act (SWMA) Senate Bill No. 1192, was passed by the Florida Legislature and became effective on October 1, 1988. Under this Act, certain items were banned from landfills. These are used motor oil which became effective October 1, 1988, lead acid batteries, which became effective on January 1, 1989, white goods which became effective on January 1, 1990, and yard trash which became

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59 effective January 1, 1992. Exclusion of these items from the landfill will reduce the amount of hazardous materials that can potentially contaminate groundwater or the surrounding land (Earle et al 1991), especially heavy metals. Landfill leachates often contain high concentrations of toxic heavy metals. Many of these metals can form strong complexes with biomolecules, therefore, even in small amounts their presence can have adverse effects on both plants and animals (Bolton and Evans, 1991) Some probable ligands in landfill leachates and residues would be chlorides, sulfates, phosphates, and organics (Pohland et al. 1981). Table 2-5 shows the median concentrations in MSW leachate, in comparison with existing exposure standards. It was interesting to note that on an overall basis, the USEPA (1988) report had lower levels than those given in the Congress of the United States (1989) report. The fact that certain items containing heavy metals were banned in 1988 and 1989 would have led one to believe that the heavy metal contents would have decreased. Obviously, items such as batteries are still finding their way into landfills even though there is now an emphasis on rerouting them away from landfills and recycling the relevant

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Table 2-5 70 The median concentrations in MSW leachate, in comparison with existing exposure standards. Metals Median concentration (ppm) Exposure standards (ppm) Antimony 0.066 0.01 Arsenic 0.042 0.05 Barium 0.853 1.0 Beryllium 0.006 0.2 Cadmium 0.022 0.01 Chromium 0.175 0.05 Copper 0.168 0.012 Iron 221 1000 Lead 0.162 0.05 Manganese 9.59 0.05 Mercury 0.002 0.002 Nickel 0.326 0.07 Selenium 0.012 0.01 Silver 0.021 0.05 Thallium 0.175 0.04 Zinc 8.32 0.110 (USEPA, 1988 and Congress of the United States, 1989)

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71 metals. Landfill leachates contain many organic and inorganic ligands, particularly chloride (Cl') which help to determine the forms of metals in solution. High concentrations of inorganic ligands, primarily chloride (CI"), and high dissolved organic carbon (DOC) concentrations affect the speciation of metals in landfill leachates. Complexation reactions in solution lower the positive charge associated with the metal and therefore may affect the mobility of leachate metals in the soil and underlying sediments (Bolton and Evans, 1991) Therefore, methods have to be developed to decrease the amount of heavy metals in leachate. Leachate recycling is one method to decrease the amount of heavy metals present. One report showed that leachate recycling did not result in an increase in heavy metal concentrations, but actually removed some of the heavy metals in the solid waste columns. The explanation provided was that under anaerobic conditions that exist in a landfill, heavy metals precipitate as sulfides (Pohland et al. 1979) Anaerobic Digestion The main steps in anaerobic digestion of MSW are: (1) Pretreatment to remove undesirable material, upgrade and homogenize the feedstock for digestion and to protect

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72 downstream treatment processes; (2) Anaerobic digestion to produce biogas for energy and to deodorize, stabilize, and disinfect the digestate product; (3) Post -treatment to complete the stabilization and disinfection of the digestate, to remove residual inert undesirable material (glass and plastics) and produce a refined product of suitable moisture content, particle size and physical structure for the proposed end-use. The end uses are energy in the form of biogas, typically around 100-200 m^ biogas per ton of organic MSW digested, also solid and liquid by-products which can be used as compost products and have a value as a fertilizer or soil improver (lEA Bioenergy, 1994) The majority of the literature on anaerobic digestion deals with zinc (Zn^"") iron (Fe^*) cadmium (Cd^*) copper (Cu^*) chromium (Cr^* Cr^*) Nickel (Ni^*) and lead (Pb^M The focus of most of these articles was the inhibition of anaerobic digestion by these heavy metals. It was determined that concentrations and species of these metals were indicative of their toxicity and the ability of microorganisms (bacteria) to adjust to their effects. Temperature, pH, hydraulic retention time, and the ratio of the toxic substance concentration to the bacterial mass concentration are

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73 determinants in the inhibitory concentrations of these heavy metals. Some inhibitory concentrations given were Cd^* (180 ppm) Fe^" (1750 ppm) Cu^* (170) ppra, Cr+ (450 ppm) Cr^^ (530 ppm) and Ni^* (250 ppm) (Chynoweth and Pullammanappallil 1996; TEA Bioenergy, 1994; Peiffer, 1993; Lin, 1993; Lin 1992; Mueller and Steiner, 1992) Determination of the heavy metal content after the digestion process is the principal determining factor in the feasibility of using the liquid or solid digestate as compost. Post -treatment follows the actual anaerobic digestion process which allows for two to four weeks of maturation for the solid fraction whereas the liquid fraction may be directly applied onto farmland as slurry, if it is of good quality and meets all set regulations. Review of the literature provided no information about the presence of Hg. Hence, it is obvious that Hg has rarely been included in past studies. In an anaerobic environment, microorganisms that do not need oxygen are called anaerobes. There are two types of anaerobes, obligate and facultative. Obligate anaerobes cannot use oxygen at all, and tend to be poisoned by it. Facultative anaerobes are able to use oxygen when it is present, but can use other sources when oxygen is absent. The

PAGE 88

74 other sources that replace oxygen are nitrate (NO3") sulfate (SO42-) and carbon dioxide (COj) When these compounds are used in the energy-generating process they are reduced. Nitrate is reduced to nitrogen gas (Nj) sulfate is reduced to hydrogen sulfide (HjS) and carbon dioxide is reduced to methane (CH4) Bacteria that use nitrate are called denitrifying bacteria and those that use sulfate are called sulfate reducers, and those that use carbon dioxide are called methanogenic bacteria (methanogens) Denitrifying bacteria are facultative anaerobes and sulfate reducers and methanogenic bacteria are called obligate anaerobes (Brock and Madigan, 1988; Brock and Brock, 1978) In an MSW anaerobic digester, there is a consortium of microorganisms that play a part in methane fermentation. Large -molecular weight compounds such as polysaccharides, proteins, and fats are converted to methane by these microorganisms. For conversion of a typical polysaccharide, this consortium includes cellulolytic bacteria which cleave high molecular weight cellulose molecule into cellobiose and free glucose, fermentative anaerobes which ferment the glucose to a variety of products including acetate, propionate, butyrate, hydrogen (H2) and CO2 Methanogenic bacteria

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75 consume any H2 that is produced. Also, certain methanogens are capable of converting acetate to methane (Brock and Madigan, 1988) Other organisms that convert complex materials to methane are Hj-producing fatty acid-oxidizing bacteria which use fatty acids or alcohols as energy sources and grow very well in the presence of Hj-consumers such as methanogens and sulfate reducers. The association between hydrogen producers and hydrogen consumers is known as interspecies hydrogen transfer which is important in the anaerobic digestion process. Syntrophomonas and Syntrophohacter are Hj -producers which generate acetate, COj, and Hj Hydrogenconsuming acetogenic bacteria (Acetoiiacterium) consume hydrogen and produce acetate for methanogenesis During methanogenesis, methanogens reduce the acetate, COj and Hj to methane and carbon dioxide which are the main end products of anaerobic digestion (Chynoweth and Pullammanappallil, 1996; Brock and Madigan, 1988). This literature review has included all three phases of the environment, namely air, earth, and water, and the many Hg interactions that take place in the environment. This literature review showed that data of Hg in landfills and MSW

PAGE 90

76 is extremely limited. However, data on metal concentrations in landfill leachate is well documented. The fact that Hgcontaining devices are still finding their way into landfills in spite of more stringent recycling efforts, is the primary reason why more information is needed on the fate of Hg in landfills. The literature on anaerobic degradation processes strongly suggests that transformation of inorganic and elemental Hg to volatile organic forms is highly likely to occur in landfills.

PAGE 91

CHAPTER 3 MATERIALS AND METHODS Phase I-Hg in the Alachua County Landfill Samples that were analyzed for total Hg in this study had been obtained from the Alachua County Landfill during a previous study (Miller et al. 1996; Miller et al 1994). Five areas were sampled and were designated as LFSl, LFS2, LFS3, LFS4, and LFS5 (Figure 3-1) LFSl is a leachate recycle area (Infiltration Pond 1) LFS2 is a control area for leachate recycling, LFS3 is a 4.5 hectares closed landfill, LFS4 is a 12 hectare closed landfill, and LFS5 is a leachate recycle area with horizontal injection. The active lined landfill (11 hectares) at this site began receiving MSW in 1988. Leachate recirculation was performed on a portion of this lined landfill (LFSl) which is adjacent to the leachate recycle area. LFS2 is actually a portion of the lined landfill that was used as a control area where there was no 77

PAGE 92

78 ^^ c rH H n XJ c U OJ Ot C c M H rH 4J IB 10 XJ U C U iH N -H H •^ ^ G t-t ,— t f-i I ~ -H -H TO U-J iw nJ 41 T3 o 0) < C c < ID 01 a 0) u >. to tn >^ o to u 11 a) rH U u OJ a 01 ij iH E e iJ (0 X M (0 £ Vj X U iJ in :] J CPi H 0) h5 •H U-l T! C (tl a >^ jj c :3 u x: V Q.1 3 en

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79 leachate recirculation occurring. LFS3 received waste from 1985 to 1988 and LFS4 received waste from 1973 to 1985 and both were in a portion of the landfill that was unlined. Solid waste samples were collected from LFS 1 both prior to and following the filling of the infiltration pond with leachate. Three boreholes were normally augured at each site with a 10.16 cm truck mounted open flight power auger that penetrated through the landfill until it reached the desired depth. In general, four samples were collected per hole at 3.05 m (10 feet) intervals. The maximum depth of the boreholes ranged from 9.14 13.72 m. The top portions of some samples (0 1.52 m) were discarded because they were mostly cover soil. Table 4-1 shows the depths of each collected sample. Additional information on sampling procedures can be found in Miller et aJ. (1996) After collection, the samples were placed into polyethylene bags and brought back to the Bioprocess Engineering Research Laboratory of the Agricultural and Biological Engineering Department at The University of Florida and stored in freezers until processed. Processing of these samples involved separation and characterization of waste components, moisture content, and solids composition

PAGE 94

80 determinations and biochemical methane potential assays. The samples were dried at 105C, grossly picked, and then passed through a 6 mm screen. The material that remained on the 6 mm screen was finely picked and ground to 0.76 mm using an Urschell Mill (Comitrol Model 3600) The material that passed the 6 mm screen was then passed through a #40 screen (0.42 mm) The material that remained on the #40 screen was finely picked and then ground to 0.76 mm using the Urschell Mill. For the purpose of total Hg analyses, the samples consisted of three fractions. One fraction was that remaining on the 6 mm screen, a second fraction was that passing the #40 screen, and the third fraction was that remaining on the #40 screen. These are shown in Table 4-1. These three fractions represented the entire sample taken from the landfill after gross picking and fine picking. In the study by Miller et al. (1996) the SW from the landfill was also characterized as paper, cardboard, food waste, glass and stone, metal, wood, plastics and rubber, and fabrics. Gross and fine picking removed glass and stones, metal, wood, plastics, rubber, and fabrics, leaving essentially paper, cardboard, and food wastes

PAGE 95

81 The elaborate separation scheme used by Miller et al. (1996) was developed to obtain the fraction of SW that is directly responsible for methane production, namely the volatile organic fraction. This is shown in Figure 3-2. For Phase I of this study, the three fractions described above were analyzed to evaluate total Hg concentrations existing in the landfill and to give an idea of the Hg distribution that occurred in such an environment Mercury was being evaluated because there has been little research concerning Hg in biologically stabilized SW residues. The fact that these samples are volatile organic fractions are important concerning Hg because Hg binds strongly to organic material, therefore the majority of the Hg in the landfill should be bound to any existing organic matter present. Total Hg in these samples was measured on a ng/g dry weight basis to give an idea of the Hg levels existing in the Alachua County Landfill. These Hg concentrations could have been converted back to a wet weight basis, but it was not deemed necessary since it is not scientifically incorrect to report on a dry weight basis. Furthermore, if the landfill is ever reclaimed and the SW applied to land, application rates will be computed on a dry weight basis. It must also be noted

PAGE 96

82 t)i> inf^ iniptf 1 t;ro.i nAini STf(r)m^ (6 niia itrttn) Fine picking S Sicipdr cKj nf if riiilion 5^ food vm If C: CUiii JL Jlorw M: MrLjl W: UooJ P: Pbilics i rubber F: Fahrki Grinding Lo 0.03" lizc \ rionbwd< (; pin ;'• A .No. 40 LVd for j5 Figure 3-2. Sample Handling Procedure (Lee, 1996)

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83 that the Hg values reported in this study may be lower than the actual values because the samples had been dried at 105C in the study of Miller et al. (1996) Elemental Hg vaporizes at around 40-50C and any elemental Hg that may have been in the original landfill samples would have been lost in the drying process Total Hg was determined using standards ranging from to 55 ng Hg. A check sample was used in every run as a way to evaluate whether or not the system was performing adequately and also as a reference point to the validity of the other samples A duplicate and a spike were run with each sample batch (12 samples) and a continuous calibration blank and a continuous calibration variance (a 50 ng Hg spike of a blank sample) were used throughout every sample batch to ensure that the instrument absorbances were being maintained. The acceptance criteria for the calibration curves (6 standards, 1 blank) was r^ > 0.99. The blanks had to be below the Method Detection Limit, the spikes had to be +/20% of the true value, and for duplicate samples, the standard error had to be < 20%. This QA/QC was used in a previous Hg study (Delfino et al 1993) Results of the QA/QC are given in Appendix A.

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84 Total Hg in the SW fractions was determined using the digestion procedure described in EPA Method 7471 for the determination of Hg in soil and sediment. Following the degestion procedure, Hg was analyzed by cold vapor atomic absorption spectrophotometry (USEPA, 1986) Two grams of sample were weighed to the nearest 0.0001 g on an enclosed Mettler analytical balance. The sample was transferred quantitatively to an acid rinsed 300 mL BOD bottle with a 10 mL double distilled deionized (DDDI) water rinse. The digestion involved 2.5 mL of trace metal grade concentrated nitric acid and 5 mL of trace metal grade concentrated sulfuric acid. The sample was heated at 95C for two minutes, then 15 mL of potassium permanganate (50 g/L) and 8 mL of ammonium peroxydisulf ate (50 g/L) were added to the digestion mixture. The sample was heated at 95C for one to two hours. An additional 15 mL of potassium permanganate was added if the color disappeared within fifteen minutes of the initial addition. Upon completion of digestion, samples were cooled and decolorized by the addition of 6 mL of hydroxylamine hydrochloride solution (120 g hydroxylamine sulfate, and 120 g sodium chloride per liter of deionized water solution) Figure 3-3 shows a schematic of the procedure.

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85 2 g dry solid waste sample + 10 mL DDDI in BOD bottle Add 5 mL sulfuric acid and 2 5 mL nitric acid Add 15 mL potassium permanganate Add 8 mL ammonium peroxydi sulfate Heat at 95 C for 1 hour Cool and decolorize samples with 6 mL hydroxylamine hydrochloride solution Measure elemental mercury absorbance on CVAAS Figure 3-3. Total mercury analysis of solid waste fractions as described in EPA Method 7471.

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86 Trace metal grade nitric and sulfuric acid are oxidizers. Potassium permanganate is another strong oxidizer that was added to eliminate possible interference from sulfide. Ammonium peroxydisulf ate is another added oxidizer to ensure the complete oxidation of the sample. Heating the sample in a water bath, along with the strong oxidizers, converts all of the Hg to Hg2*. Each digested SW sample was transferred to a plastic reaction vessel fitted for a Perkin Elmer MHS-10 cold vapor unit The sample was purged with high purity nitrogen gas Stannous chloride was used to reduce Hg^* to the elemental state, Hg. Elemental Hg was then swept into an open ended quartz tube (1 cm diameter) with a 16 cm cell path length. The Hg was quantified by CVAAS using a Perkin Elmer Model 3 03 Atomic Absorption Spectrophotometer {X = 253.6 nm, SBW = 0.7 nm) with Hg hollow cathode lamp (1 = 6 MA) The standard calibration curve working range (0 to 550 ng) gave an absorbance range of 0.017 to 0.260. Total Hg (ng/g) was determined using the equation derived from the regression analysis of the Hg standard curve. From the equation:

PAGE 101

"! 87 y = mx + b Mercury (ng) was obtained using the noted absorbance from the atomic absorption minus the yintercept divided by the xcoefficient, i.e., (abs b) /m Since the SW samples from the Alachua County landfill had been dried prior to Hg analysis, then the determination of Hg(ng/g) involved Hg(ng) divided by dried weight (g) measured out prior to analysis. The Hg stock solution (1000 ppb) was made using 0.1 mL of 1000 ppm +1% Mercuric Nitrate, where 1 mL = 1 mg Hg (Fisher Scientific) and adjusting the volume to 100 mL in a 100 mL volumetric flask. The top of the volumetric flask was sealed with parafilm and the entire flask was enclosed in aluminum foil to slow any possible degradation that may occur due to the infiltration of light. Successive dilutions of the stock Hg solution were made to obtain a working standard containing 10 ppb or 100 ppb. This working standard was prepared fresh daily prior to sample preparations. Acidity of the working standard was maintained at 0.15% trace metal grade nitric acid.

PAGE 102

Aliquots were taken from the working standard to develop a range of standards that the sample data should fall within. The Hg standards were adjusted to a total of 10 mL with DDDI water. Next, the same procedure as stated above was repeated, i.e., the addition of the acids, potassium permanganate, ammonium peroxydisulf ate, the digestion in the water bath at 95C, and finally the addition of sodium chloridehydroxylamine sulfate. The stannous chloride on the MHS-10 unit was used as the reducing agent and the elemental Hg absorbances were then noted on the AAS Due to the toxicity of Hg vapor, it was imperative that there be a well working vent over the instrument during analyses Compost samples were obtained from Dr. D. Graetz in the Soil and Water Science department at the University of Florida. These were previously collected from the Palm Beach County compost facility and consisted of a 1:1 mixture of biosolids to yard waste. These samples were air-dried, ground, and total Hg analyses were performed using EPA Method 7471. The Hg concentrations in these aerobic compost samples were compared with those found in the SW from the Alachua County Landfill. The raw data for the Palm Beach County compost samples are provided in Appendix A.

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89 Phase II -Landfill Simulated Anaerobic Reactors Reactor design A schematic of the reactor design is shown in Figure 3-4. Each reactor was made from PVC pipe measuring 50.8 mm diameter and 635 mm long. PVC threaded end caps with inserted plastic threaded tees were used at either end of the reactor. There was a 38.1 mm diameter and 304.8 mm long schedule 40 PVC insert which contained the SW. The insert had holes drilled throughout it to enable the free movement of leachate into and out of the SW. At the bottom end of the insert there was a glass fiber mesh to hold the SW intact, but at the top it was left open. Two additional 38.1 mm diameter PVC inserts were placed at either end of the SW insert to decrease the amount of SW movement in the leachate circulation process. The additional inserts also had glass fiber mesh at the ends to keep the SW in place. The plastic tee at the bottom of the reactor had septa that were secured with copper wire and/or clamps to allow for the collection of leachate. The tee at the top of the reactor had a septum at one end for the collection of gas samples via syringe. The other end of the tee had a short piece of 6.35 mm diameter flexible tygon tubing (Norton Co.) to which a 3 -way valve was inserted.

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90 i IS u u li a Cn 0) )-l a m e a" W Eh (8 > S H m > Tygon Tubing u a > nJ ft u /^ r en H 03 0) •rJ o u (0 (U o u H 4J n3 e 0) U I U tn -H fa
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91 followed by tygon tubing that led into a large test tube fitted with a 2 -hole rubber stopper that served as a water trap to prevent any leachate from entering the carbon traps Tygon tubing then connected the water trap to two sulfur impregnated carbon traps in series. The sulfur impregnated carbon (10-12% Sulfur) in the traps was obtained from Norit Americas and the grade was Norit RBHG 3 and the sample number was A9805. The encasing for the traps was predrying tubes with end caps (Bel -Art Products) The tubing from the carbon traps then went into an inverted 1000 mL plastic graduated cylinder fitted with a 2 -hole rubber stopper which served as a gas manometer. The gas manometers contained an aqueous 5% hydrogen chloride (HCl) solution for the purpose of reducing the amount of COj being dissolved. The acid solution in the gas manometers which was displaced by gas generated in the reactors was connected to a main acid solution reservoir consisting of a 1000 mL graduated cylinder, which was open to the atmosphere. This main reservoir served to refill the gas manometers with acid solution as needed. To measure the amount of gas produced by each reactor, the main reservoir was moved over to each inverted graduated cylinder and their liquid levels were

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.1 ] 92 matched and then noted, ensuring that all gas volumes were measured at atmospheric pressure. The reactors were incubated at 50 2C. Figure 3-5 shows the entire setup of the system. Figure 3-6 shows a view of the water traps that prevented leachate from entering the carbon traps. Figure 3-7 shows how the solid waste insert was placed into the encasing. Figure 3-8 shows the glass fiber mesh at the bottom of the SW inserts. Figure 3-9 shows the top part of the system including the carbon traps. Figure 3-10 shows a close-up view of the carbon traps. There were two reactor runs performed using this system. I Solid Waste The SW used in this study was artificially made to represent the acutal landfill SW left behind when metals, glass, plastics, and wood have been removed i.e., gross and fine picked. The artificial waste consisted of office paper, newspaper, and dog food, the last ingredient being added to represent the food portion in an MSW sample. The office paper and newspaper were finely shredded. The dog food consisted of dry pellets that were first mixed with the shredded paper, ground, and finally the entire mixture homogenized as well as possible

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93 Figure 3-5. Entire setup of the reactor system, Figure 3-6. Liquid collection traps behind the reactors

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94 Figure 3-7. Solid Waste insert into reactor Figure 3-8. Glass fiber mesh at bottom of solid waste insert

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95 OOOT Figure 3-9. Top view of the reactor system Figure 3-10. Close-up view of the activated carbon traps,

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96 The insert was packed with 60 g of the artificial SW at a packing density of 3 00 Ib/yd^ This density was chosen because it represents the density of SW in a packing truck, i.e. 300-400 Ib/yd^ The packing involved determining the volume of the insert and how much SW would be needed to yield a packing density of 300 lb/yd\ Volatile solids analyses on the SW were performed in triplicate according to Standard Methods for the Examination of Water and Wastewater, 1989, Section 2540E Fixed and Volatile Solids, pp. 2-77. Reactor operation The nine reactors were all operated in the same manner. During setup, the SW was spiked with Hg prior to being placed into the reactor inserts. Spiking was done by adding the appropiate volume of a 100 /xg Hg/L standard to the SW. These were then left to sit for about an hour at room temperature so that the Hg could bind to the SW. Next, these inserts were placed into the reactors and 600 mL of leachate from another digester was used as the innoculum in each reactor. Initial analyses of methane, pH, VFAs, volatile solids and Hg levels were perfomed on separate samples of the SW and the leachate. In the first run, the first three reactors were used as controls. They contained the SW and the leachate, but they

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97 were not spiked with Hg. Three replicates were used at two different Hg levels, namely 100 ng Hg and 2000 ng Hg. In the second run, the first three reactors were once again used as controls, but two replicates were used at three different Hg levels, namely 100, 1000, and 2000 ng Hg. In both runs, a total of nine reactors were used. At the same time on a daily basis, the liquid levels in the gas manometers were noted to determine the amount of gas produced by the reactors in a 24 hour period. The valves leading from the reactors were then closed off to the reactors to allow for the collection of gas samples without introducing air to the system. After gas sample collection, the reactors were inverted to allow for recirculation of the leachate and also to flush out the volatile fatty acids. The reactors remained inverted for a period of about 3 minutes and then returned to their original position after which leachate samples were collected. Liquid levels in the gas manometers were then reset for the next day's readings. The carbon traps were removed at the end of the reactor run for the analysis of Hg in the gas phase. The leachate was also analyzed for Hg at the completion of the reactor run. It was imperative that the solids not be sampled until the run

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98 was complete, otherwise the entire system would be disturbed as air would be introduced into the system. The septa at the top and the bottom tees of the reactors were changed every couple of days as the constant sampling would create holes large enough for the introduction of air into the system, thus causing destabilization of the redox conditions. General maintenance was performed on a regular basis, such as tightening the ties on the tygon tubing, tightening the screw caps on the reactors, and checking the gas manometers for liquid leaks. The temperature on the incubator was also checked regularly to ensure that it was being maintained at 5 0C. All readings and sampling were taken inside the incubator at this temperature. The only time that the incubator door was open was when a person was entering or leaving. It was never left open for any amount of time so that the temperature could be maintained. MSW Analysis Total solids and volatile solids analyses were conducted in triplicate. A portion of the sample was dried at 105C for 24 hours to determine the moisture content, after which it was placed in a 550C oven to ash. The difference between the dried solids and the ashed solids provided volatile solids

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99 present in the sample. A smaller percentage of volatile solids in a sample would indicate less organic matter and hence, a more highly stabilized sample. Volatile solids indicates the potential degree of waste decomposition. Gas Sampling and Analysis Gas samples were taken at the same time each day for the first seven days, and then every other day afterwards. The samples were taken prior to the recycling of leachate. The septum at the top tee of the reactor allowed for access with a 10 mL syringe. Samples were taken to the laboratory right after collection and analyzed immediately. These samples were run in triplicate for statistical purposes. Gas composition analysis for CH4, COj, O2 and Nj was measured on a Fisher Model 1200 Gas Partitioner chromatograph equipped with two stainless steel columns and a thermal conductivity detector. A 10 mL aliquot of gas was injected manually into a sampling loop which conducted sample into each of two columns. The first column was 2 m x 3.2 mm packed with 80/100 mesh Poropak Q (Supelco Inc.) used for the separation and detection of CO2, and the second column was 3.35 m X 4.8 mm packed with 60/80 mesh molecular sieve 13X (Supelco, Inc.) for the separation of O2, Nj, and CH4 The

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100 column temperature was 50C, the detector bridge current was 175 mA, and the attenuator was set at 4. Helium was the carrier gas at a flow rate of 3 cm^/min. Standard gas samples with known concentrations of the four compounds were initially injected to make sure that the instrument was performing properly. Volatile Fatty Acids Standards Preparation To a tared 10 mL volumetric flask, 1 g isobutyric acid was added and the weight recorded. The flask was filled to the line with DDDI water and mixed well. To a 1 L volumetric flask, 2 00 mL DDDI water was added. One hundred mL of Phosphoric acid was added slowly to the water in the 1 L flask. The isobutyric acid solution was then added to the 1 L flask and the 10 mL flask was rinsed six times with DDDI water and the rinsate was added to the 1 L flask. The 1 L flask was then filled to the line with DDDI water. This isobutyric acid solution was decanted into a 1 L amber bottle and sealed with a teflon cap. This bottle was then labeled "VFA Internal Standard" and dated. Next, another 10 mL volumetric flask was tared and 1 g each of acetic, propionic, butyric, isovaleric, and valeric

PAGE 115

101 acids were added to the flask. Each weight was recorded and rezeroed between acids. The 10 mL flask was then filled to the line with DDDI water and mixed. To a 1 L volumetric flask, 500 ml of DDDI was added. Once, again, the 10 mL flask was rinsed six times with DDDI and the rinsate added to the 1 L flask. The 1 L flask was then filled with DDDI water and mixed well. The acids solution was then decanted into a 1 L amber bottle and sealed with a teflon cap. It was labelled "VFA Calibration Standard" and dated. The VFA internal standard solution and the VFA calibration standard solution were kept refrigerated between uses These were made fresh prior to the start of the first run and were kept in a refrigerator for preservation purposes. Leachate Sampling Leachate samples were collected for the first seven days of the reactor run and then every other day after that Samples were collected through the septum at the bottom tee of the reactor after the recycling of leachate. This sampling was performed using a needle on a 10 mL syringe. Nine mL of the unfiltered leachate sample was stabilized with 1 mL of 85% H3PO4. This was then distributed into three 2 mL centrifuge tubes and spun down in a centrifuge at 3000 RPM for at least

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102 three minutes. To an 8 mm gas chromatograph glass sample vial, lll.l|aL of the internal standard, isobutyric acid, was added, followed by 1 mL of the centrifuged supernatant sample. The sample vials were then sealed and mixed by vortex. At this point, the sample was ready for analysis. These samples were run in triplicate. Leachate Analysis A Shimadzu gas chromatograph/f lame ionization detector (GC/FID) fitted with an auto sampler and equipped with a 1.8 m X 2 mm glass column of 10% SP-1000 on 100/120 Chromosorb W/AW (Supelco, Inc.) was used for VFA analysis (Chynoweth et al. 1990) The column was maintained at 180C, the injector at 180C, and the flame ionization detector at 200C. Nitrogen gas was used as the carrier gas at a flow rate of 3 cmVmin. Air was used to operate the autosampler and also the FID. A Perkin Elmer Model integrator was used to calculate the concentration in mg/L of the VFA's. The VFA's that were analyzed were acetate, propionate, isobutyrate, butyrate, isovalerate, and valerate. One microliter of sample was injected into this GC.

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103 pH AnalYsis Prior to VFA analysis, pH was determined on the leachate samples that were collected. A Fisher pH meter, model #805 MP was used after calibration with pH 4 and 9.0 buffers. Mercury Analyses MSW Mercury Analysis Total Hg analysis was performed on the SW prior to and after the reactor run, thus providing initial and final Hg concentrations in the SW. This involved using the digestion procedure described in EPA Method 7471. The solid samples were collected at the end of the reactor run. The SW was placed in a ziploc freezer bag and mixed as well as possible. If analysis was not going to be performed at that time, then the bags were placed in the freezer until then. The normal holding time for these samples was 2 8 days. Carbon Trap Sampling The carbon traps were removed by clamping both ends of the tygon tubing entering and leaving the traps so as not to allow for the introduction of air. When the traps were removed, the ends were sealed with parafilm and taken to the lab for the carbon to be analysed by CVAAS Activated carbon

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^ 104 (especially sulfur and iodine impregnated activated carbon) has been distinguished as an excellent sorbent for elemental Hg (Vidic and McLaughlin, 1996; Krishnan et al. 1994, Moffitt and Kupel, 1971) Carbon Trap Analysis and Method Development A method had to be developed for the determination of Hg in the carbon traps. The carbon in the traps was impregnated with sulfur which binds Hg tightly to produce one of the most stable forms of Hg, mercury sulfide, when the Hg in the gas phase was passed through it The usual EPA method for the determination of Hg in solid material had to be tested and modified as necessary to achieve full recovery of Hg from the carbon traps Prior to running the reactors, the traps were built and tested for their ability to trap elemental Hg. This was performed by reducing a known Hg standard in solution to the elemental form by using the reductant stannous chloride in the MHS-10 cold vapor unit. The standard samples were analysed to note the absorbance readings and then the sulfur impregnated carbon traps were put on the line leading from the cold vapor unit to the quartz cell and the absorbances were noted again. It was determined that the traps did in fact trap all of the

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105 elemental Hg being passed though the tube since trapping completely eliminated any absorbance as measured by the AA. It was then observed that the EPA digestion procedure described in Phase I could not recover any of the Hg caught on the carbon traps. The literature was exhausted for means to remove Hg from the Simpregnated carbon traps, but no reports were found. It apparently required a stronger digestion step than the EPA Method 74 71 used. In the original method, 10 mL of DDDI water, 2.5 mL of nitric acid, 5 mL of sulfuric acid, 15 mL of potassium permanganate and 8 mL of ammonium peroxydisulf ate were used in the Hg digestion of the carbon traps. The first modification of the method involved using 10 mL of distilled deionized water, 5 mL of nitric acid, 10 mL of sulfuric acid, 3 mL of potassium permanganate and 16 mL of ammonium peroxydisulf ate, but no Hg was again recovered from the carbon traps. The final modification of the method used 10 mL of DDDI water, 7.5 mL of nitric acid, 15 mL of sulfuric acid, 45 mL of potassium permanganate, and 24 mL of ammonium peroxydisulfate and the time in the water bath lengthened to over two hours. This time, the Hg recovered from the traps was >98%. The modified digestion procedure was repeated to confirm these recoveries and they were found to be reprocible.

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106 Figure 3-11 shows a schematic of the method for the analysis of Hg in sulfur impregnated carbon. The data for the method development for Hg analysis of the carbon traps are presented in the Results section. Mercury in Leachate All of the leachate at the end of the reactor runs was collected by removing one of the bottom septums and letting it flow into an acid-rinsed bottle which was then taken to the lab for analyses. If analyses were not going do be done right away, the samples were placed in a freezer until such time. The Hg in the leachate samples was determined using EPA Method 7470 which is a cold vapor atomic absorption procedure that is used for the determination of Hg in mobility-procedure extracts, aqueous wastes, and groundwaters. It can also be used for the analyses of solid sludge-type wastes, even though EPA Method 7471 is better suited for sludge analyses. EPA Method 7470 involved transferring 100 mL of the sample, in this case, leachate, to an acid rinsed 3 00 mL BOD bottle. Added to this leachate sample were 5 mL of concentrated sulfuric acid and 2.5 mL of concentrated nitric acid. The sample was heated at 95C for two minutes, then 15 mL of potassium permanganate (50 g/L) and 8 mL of ammonium

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107 4 g sulfur impregnated carbon + 10 mL DDDI in BOD bottle Add 15 mL sulfuric acid and 7 5 mL nitric acid Add 45 mL potassium permanganate Add 24 mL ammonium peroxydi sulfate Heat at 95 C for 2 hours Cool and decolorize samples with 6 mL hydroxylamine hydrochloride solution Measure elemental mercury absorbance on CVAAS Figure 3-11. Modification of EPA Method 7471 for determination of total mercury in sulfur-impregnated carbon.

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)' M 108 peroxydisulfate (50 g/D were added to the digestion mixture. The sample was heated at 95C for one to two hours. It was not necessary to add any extra potassium permanganate as the leachate samples were not prone to change color. Upon completion of digestion, samples were cooled and decolorized by the addition of 6 mL of hydroxylamine hydrochloride solution (12 g hydroxylamine sulfate, and 12 g sodium chloride per liter of deionized water solution) Figure 3-12 shows a schematic of the procedure. Tyqnn Tubinq Kxperim pint for Mercury Another experiment tested the ability of tygon tubing at room temperature (24C) to absorb elemental Hg since tygon tubing was used to connect the carbon traps to the anaerobic digesters. This experiment was performed to determine the role that tygon tubing may or may not play in the entrapment of elemental Hg. Tygon tubing was attached to the cold vapor unit leading to the quartz cell and the absorbances of known Hg standards were noted on the CVAAS The tubing did not appear to have any effect on the Hg absorbances, thus it did not appear to absorb elemental Hg. An additional experiment was performed to determine whether tygon tubing could trap elemental Hg and Hg^^ at 50C

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109 100 mL leachate sample in a BOD bottle Add 5 mL sulfuric acid and 2 5 mL nitric acid Add 15 mL potassium permanganate Add 8 mL ammonium peroxydisulfate Heat at 95 C for 1 hour Cool and decolorize samples with 6 mL hydroxylamine hydrochloride solution Measure elemental mercury absorbance on CVAAS Figure 3-12. Total mercury analysis in leachate samples as described in EPA Method 7470.

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110 and whether Hg^* could volatilize at 5CPC and be trapped on the S-impregnated activated carbon. This was a 60 hour experiment that was set up in an incubator room that operated at 50C. There were two pairs of flasks each connected in series to a pair of S-impregnated activated carbon traps. An aquarium pump with a total flow rate of 30 ml/min was attached to the two pairs of flasks. All four flasks contained DDDI water and were spiked with 100 ng Hg^* as Hg(N03)2. To one pair of flasks, SnClz was added to reduce the Hg^* to Hg No SnClz was added to the other pair, leaving the Hg in the +2 valence form. Although these experiments showed that tygon tubing did not trap either elemental or divalent Hg, the tygon tubing that ran between each reactor and the carbon traps was weighed, cut up into small pieces and digested as in EPA Methd 7471 following each run. The tygon tubing was analyzed as a safety precaution in case there were other forms of Hg, i.e., organic forms such as DMHg, trapped by the tubing. Thus, all forms of Hg being volatilized, including organic forms, would be quantitatively recovered during the anaerobic digestion experiments

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CHAPTER 4 RESULTS This chapter will be divided into three sections. The first section will present Hg data from Phase I The second section will present anaerobic reactor data obtained in Phase II and the last section will present Hg data obtained in Phase II. Hg concentrations in the Alachua County Landfill In Phase I, the SW composite samples from the Alachua County landfill, which had been split up into fractions as indicated in Chapter 3, were analyzed for total Hg. These fractions had been oven dried and were analyzed for Hg (Hg) using USEPA Method 7471. The Hg data that were obtained for each fraction from the Alachua County landfill are presented in Table 4-1. The fractions are called remaining on 6 mm, remaining on #40, and passing #40. These fractions add up to a total of 100% of dried sample minus what was included in "gross picked" and "fine picked" debris as described in Methods and Materials. Overall, there were 261 fractions for 111

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112 Table 4-1. Alachua County Landfill Total Mercury Concentrations in three fractions. SAMPLE DEPTH (m) Remaining on 6mm Passing #40 Remaining on #40 [Hg] ng/g [Hg] ng/g [Hg] ng/g LFS 1.030 1.52-4.57 69.1 45.1 162 LFS 1.030 4.57-7.62 162 75.6 204 LFS 1.030 7.62-10.67 145 410 368 LFS 1.030 1.52-4.57 407 18.0 76.2 LFS 1.030 4.57-7.62 82.0 79.0 LFS 1.030 7.62-10.67 131 304 LFS 1.030 10.67-13.72 183 109 447 LFS 1.030 1.52-4.57 68.4 45.1 123 LFS 1.030 4.57-7.62 144 70.6 272 LFS 1.030 7.62-9.14 40.1 66.8 102 LFS 1.030 7.62-10.67 188 LFS 1.040 1,52-4.57 101 48.1 194 LFS 1.040 7.62-10.67 49.8 84.6 315 LFS 1.040 10.67-13.72 116 96.8 208 LFS 1.040 3.05-6.10 96.5 LFS 1.05 162 LFS 1.060 0-4.57 491 855 LFS 1.060 4.57-9.14 91.0 239 LFS 1.060 0-4.57 62.9 64.2 129 LFS 1.060 4.57-9.14 293 337 LFS 1.080 0-3.05 267 328 307 LFS 1.080 3.05-6.10 2760 413. 2590 LFS 1.080 6.10-9.14 274 267 LFS 1.080 0-3.05 215 86.7 198 LFS 1.080 3.05-6.10 208 62.4 274 LFS 1,080 6.10-9.14 304 168 220

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Table 4-1 (continued) 113 SAMPLE LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 1.090 LFS 2.030 LFS 2.030 LFS 2.030 LFS 2.030 LFS 2.030 LFS 2.030 LFS 2.060 LFS 2.060 LFS 2.070 LFS 2.070 LFS 2.070 LFS 2.070 LFS 2.080 LFS 2.080 LFS 2.080 LFS 2.080 LFS 2.080 LFS 2.080 LFS 2.080 LFS 2.080 DEPTH (m) 1.52-4.57 4.57-7.62 7.62-10.67 10.67-13.72 1.52-4.57 4.57-7.62 7.62-10.67 10.67-13.72 1.52-4.57 4.57-7.62 7.62-10.67 0-3.05 3.05-6.10 6.10-9.14 0-3.05 3.05-6.10 6.10-9.14 0-4.57 4.57-9.14 0-3.05 3.05-6.10 6.10-9.14 7.62-8.38 0-3.05 3.05-6.10 6.10-9.14 9.14-12.19 12.19-13.72 0-3.05 3.05-6.10 6.10-9.14 Remaining on 6mm [Hg] ng/g 263 543 21600 245 206 209 313 217 170 268 12.5 461 181 12.5 153 110 35.3 148 131 80.2 586 350 312 67.0 89.3 481 Passing #40 [Hg] ng/g 713 416 35.1 30.4 97.8 120 13.1 20.3 511 66.9 60.2 205 205 68.7 14.8 23.9 196 Remaining on #40 [Hg] ng/g 252 252 > 70,000 1650 299 227 578 308 249 359 2290 539 71.2 150 271 185 50.5 113 373 149 139 369 205 159 78.3 103 544

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Table 4-1 (cont'd) 114 SAMPLE DEPTH (m) Remaining on 6mm Passing #40 Remaining on #40 [Hg] ng/g [Hg] ng/g [Hg] ng/g LFS 2.090 0-3.05 17.1 259 LFS 2.090 3.05-6.10 245 561 LFS 2.090 6.10-9.14 483 2110 LFS 2.090 0-3.05 46.4 137 LFS 2.090 3.05-6.10 358 179 LFS 2.090 9.14-12.19 367 860 LFS 2.090 12.19-13.72 99.0 1080 LFS 2.090 6.10-9.14 149 283 LFS 2.090 10.67-13.72 221 190 LFS 3.090 3.05-6.10 1050 426 LFS 3.090 6.10-9.14 623 1110 LFS 3.090 9.14-12.19 39.8 265 LFS 3.090 12.19-13.72 1020 797 LFS 3.090 3.05-6.10 209 659 LFS 3.090 6.10-9.14 1920 606 LFS 3.090 9.14-12.19 1010 3260 LFS 3.090 12.19-13.72 219 1560 LFS 4.080 3.05-6.10 131 46.7 104 LFS 4.080 6.10-9.14 363 60.2 284 LFS 4.080 3.05-6.10 28.5 42.0 48.8 LFS 4.080 6.10-9.14 224 112 214 LFS 5.040 0-3.05 33.9 97.1 144 LFS 5.040 3.05-6.10 115 61.7 124 LFS 5.040 6.10-9.14 781 177 1020 LFS 5.040 9.14-12.19 129 3300 425 LFS 5.040 0-3.05 1270 962 LFS 5.040 3.05-6.10 122 229 LFS 5.040 6.10-9.14 555 1160 LFS 5.040 9.14-12.19 129 3120 10400 LFS 5.040 0-3.05 33.6 41.4 2.4 LFS 5.040 3.05-6.10 357 328 884 LFS 5.040 6.10-9.14 94.1 162 287 LFS 5.040 9.14-12.19 64.0 102 98.3

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Table 4 -1 (cont 'd) SAMPLE DEPTH (m) Remaining on 6mm Passing #40 Remaining on #40 [Hg] ng/g [Hg] ng/g [Hg] ng/g LFS 5.070 1.52-2.29 298 16.3 LFS 5.070 4.57-5.33 LFS 5.070 7.62-8.38 247 296 LFS 5.070 10.67-11.43 175 133 LFS 5.070 0-3.05 36.5 LFS 5.070 0-3.05 115 36.5 LFS 5.070 3.05-6.10 342 211 LFS 5.070 6.10-9.14 73.5 33.7 LFS 5.070 9.14-12.19 326 LFS 5.070 0-3.05 438 925 LFS 5.070 3.05-6.10 258 657 LFS 5.070 6.10-9.14 108 162 LFS 5.070 9.14-12.19 465 460 LFS 5.070 1.52-2.29 65.2 181 LFS 5,070 4.57-5.33 99.6 418 LFS 5.070 7.62-8.38 340 241 LFS 5.070 10.67-11.00 63.8 404 LFS 5.090 1.52-4.57 60.3 188 LFS 5.090 4.57-7.62 196 225 LFS 5.090 7.62-10.67 348 646 LFS 5.090 10.67-13.72 299 977 LFS 5.090 1.52-4.57 97.5 247 LFS 5.090 4.57-7.62 75.2 139 LFS 5.090 7.62-10.67 674 2500 LFS 5.090 10.67-13.72 251 318 115 All Raw Data are presented in Appendix A. Spike recoveries were within limits (80 120%) R values for standard curves were > 0.990

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116 which total Hg was determined. The total Hg that was computed in the various fractions ranged from 12.5 to greater than 70,000 ng/g Hg. Raw data along with AQ/QC for Phase I are shown in Appendix A. It should be noted that in the digestion step, an excess of potassium permanganate (KMn04) one of the oxidizers, had to be added so that the purple color would persist in the sample solution. The KMrxO^ that was required was 3 0-45 mL versus the required 15 mL in the USEPA Method. After total Hg was determined in each of the three fractions, total Hg in the composite samples was calculated by multiplying the percent of the composite weight that each fraction represented by the total Hg concentration in that fraction and adding the three numbers together. This gives the Hg (ng/g) in the composite sample, i.e. Hg in Composite = £FiCi/lOO where, F^ is the weight percent of one of the three fractions, Ci is the Hg concentration corresponding to the weight fraction i, and i goes from 1 to 3 Table 4-2 shows the percentages that each fraction comprised of the total composite sample weight, F^ Values for C^ were taken from Table 4-1. Table 4-3 shows the Hg concentrations in the

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117 Table 4-2. 's Dry weight of Total Composite Sample SAMPLE LFS 1.0301 LFS 1.0301 LFS 1.0301 LFS 1.0302 LFS 1.0302 LFS 1.0302 LFS 1.0302 LFS 1.0303 LFS 1.0303 LFS 1.0303 LFS 1.0303 LFS 1.0401 LFS 1.0401 LFS 1.0401 LFS 1.0402 LFS 1.05 LFS 1.0602 LFS 1.0602 LFS 1.0603 LFS 1.0603 LFS 1.0801 LFS 1.0801 LFS 1.0801 LFS 1.0802 LFS 1.0802 LFS 1.0802 Remaining on 6 mm Passing #40 46.0 40.1 38.5 41.9 41.8 34.9 25.3 52.2 38.8 31.6 25.9 17.0 25.7 18.7 31.0 21.1 39.9 14.3 35.1 35.1 34.7 64.0 29.7 39.1 36.9 48.1 44.3 51.3 46.3 49.4 58.8 33.4 47.7 48.3 53.9 66.3 63.6 65.9 46.6 53.6 22.2 22.0 54.4 52.2 52.1 22.7 60.2 44.3 Remaining on #40 17.1 11.8 17.2 6.8 11.9 15.7 15.9 14.4 13.5 20.1 20.2 16.7 11.0 15.5 22.4 25.3 37.9 63.7 10.4 12.8 13.3 13.3 10.2 16.6

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118 Table 42 (Cont'd) SAIMPLE Remaining on 6 mm Passing #40 Remaining on #40 LFS 1.0901 35.1 54.4 10.4 LFS 1.0901 35.1 52.2 12.8 LFS 1.0901 34.7 52.1 13.3 LFS 1.0901 64.0 22.7 13.3 LFS 1.0902 29.7 60.2 10.2 LFS 1.0902 39.1 44.3 16.6 LFS 1.0902 43.6 44.8 11.6 LFS 1.0902 43.5 49.0 7.5 LFS 1.0903 35.3 52.5 12.3 LFS 1.0903 39.1 44.4 16.5 LFS 1.0903 44.7 39.0 16.3 LFS 2.0301 100.0 LFS 2.0301 38.1 51.1 10.8 LFS 2.0301 22.5 59.8 17.7 LFS 2.0302 29.7 57.9 12.4 LFS 2.0302 31.0 61.0 8.0 LFS 2.0302 42.6 36.8 20.6 LFS 2.0603 43.3 29.5 27.2 LFS 2.0603 21.1 60.4 18.5 LFS 2.0701 33.5 52.1 14.4 LFS 2.0701 25.8 62.8 11.4 LFS 2.0701 33.9 49.1 17.0 LFS 2.0702 LFS 2.0801 43.6 44.8 11.6 LFS 2.0801 43.5 49.0 7.5 LFS 2.0801 35.3 52.5 12.3 LFS 2.0801 39.1 44.4 16.5 LFS 2.0801 44.7 39.0 16.3 LFS 2.0802 49.8 37.7 12.5 LFS 2.0802 47.0 42.2 10.8 LFS 2.0802 25.6 64.7 9.8

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Table 4-2 (Cont'd) 119 SAMPLE Remaining on 6 mm Passing #40 Remaining on #40 LFS 2.0901 47.0 42.2 10.8 LFS 2.0901 25.6 64.7 9.8 LFS 2.0901 16.6 74.8 8.6 LFS 2.0902 26.0 62.3 11.7 LFS 2.0902 16.4 74.9 8.8 LFS 2.0902 54.5 28.7 16.8 LFS 2.0902 62.8 19.3 17.9 LFS 2.0903 46.1 37.4 16.5 LFS 2.0903 29.5 55.3 15.2 LFS 3.0901 LFS 3.0901 LFS 3.0901 LFS 3.0901 LFS 3.0902 LFS 3.0902 LFS 3.0902 LFS 3.0902 LFS 4.0801 16.6 74.8 8.6 LFS 4.0801 26.0 62.3 11.7 LFS 4.0802 16.4 8.8 74.9 LFS 4.0802 30.6 14.5 54.9 LFS 5.0401 42.1 40.2 17.7 LFS 5.0401 56.5 26.0 17.5 LFS 5.0401 56.3 30.0 13.7 LFS 5.0401 18.8 62.5 18.7 LFS 5.0402 66.4 20.1 13.5 LFS 5.0402 18.6 65.9 15.5 LFS 5.0402 41.5 39.0 19.5 LFS 5.0402 32.7 53.2 14.1 LFS 5.0403 28.6 59.7 11.7 LFS 5.0403 63.0 22.3 14.7 LFS 5.0403 63.4 22.3 14.3 LFS 5.0403 33.1 45.7 21.2

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Table 42 (Cont'd ) SAMPLE Remaining on 6 mm Passing #40 Remaining on #40 LFS 5.0701 24.1 70.4 5.6 LFS 5.0701 LFS 5.0701 71.3 12.2 16.5 LFS 5.0701 79.0 11.1 9.8 LFS 5.0702 LFS 5.0703 33.4 55.5 11.1 LFS 5.0703 LFS 5.0703 53.8 29.0 17.2 LFS 5.0703 41.9 32.7 25.4 LFS 5.0704 48.6 37.5 13.9 LFS 5.0704 LFS 5.0704 58.7 29.5 11.8 LFS 5.0704 56.5 18.5 25.1 LFS 5.0705 25.4 67.5 7.1 LFS 5.0705 22.9 61.5 15.7 LFS 5.0705 71.9 17.1 11.1 LFS 5.0705 16.1 73.7 10.3 LFS 5.0901 LFS 5.0901 LFS 5.0901 LFS 5.0901 LFS 5.0902 LFS 5.0902 LFS 5.0902 LFS 5.0902 120

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Table 4-3. Alachua County Landfill Mercury Concentrationsl21 in the composite (weighted sum of the 3 fractions) SAMPLE DEPTH (m) Remaining on 6 mm [Hg] ng/g Passing #40 [Hg] ng/g Remaining on #40 [Hg] ng/g Composite [Hg] ng/g LFS 1.0301 LFS 1.0301 LFS 1.0301 LFS 1.0302 LFS 1.0302 LFS 1.0302 LFS 1.0302 LFS 1.0303 LFS 1.0303 LFS 1.0303 LFS 1.0303 LFS 1.0401 LFS 1.0401 LFS 1.0401 LFS 1.0402 LFS 1.05 LFS 1.0602 LFS 1.0602 LFS 1.0603 LFS 1.0603 LFS 1.0801 LFS 1.0801 LFS 1.0801 LFS 1.0802 LFS 1.0802 LFS 1.0802 1.52-4.57 4.57-7.62 7.62-10.67 1.52-4.57 4.57-7.62 7.62-10.67 10.67-13.72 1.52-4.57 4.57-7.62 7.62-9.14 7.62-10.67 1.52-4.57 7.62-10.67 10.67-13.72 3.05-6.10 0-4.57 4.57-9.14 0-4.57 4.57-9.14 0-3.05 3.05-6.10 6.10-9.14 0-3.05 3.05-6.10 6.10-9.14 31.8 65.0 55.8 171 34.3 46.3 35.7 55.8 12.7 26.3 8.5 29.9 152 25.1 41.9 93.6 967 95.0 138 61.8 119 16.6 36.3 182 9.2 36.6 64.5 64.2 15.1 33.7 32.3 25.9 56.1 61.6 107 48.8 14.2 179 216 139 19.7 37.5 74.6 27.7 24.1 63.3 5.2 0.0 47.8 71.0 17.8 36.8 20.4 39.2 52.6 22.8 191 60.5 49.0 215 32.0 330 26.3 27.8 36.4 76.2 125 301 185 70.9 112 182 68.5 126 65.4 91.4 117 114 107 343 109 88.3 257 305 1510 234 184 127 230

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Table 43 (conf d) 12 2 SAMPLE DEPTH (m) Remaining on 6 mm Passing #40 Remaining on #40 Composite [Hg] ng/g [Hg] ng/g [Hg] ng/g [Hg] ng/g LFS 1.0901 1.52-4.57 92.6 26.3 119 LFS 1.0901 4.57-7.62 190 32.1 222 LFS 1.0901 7.62-10.67 7470 9280 16800 LFS 1.0901 10.67-13.72 219 219 LFS 1.0902 1.52-4.57 72.8 30.4 103 LFS 1.0902 4.57-7.62 80.7 i?.i 118 LFS 1.0902 7.62-10.67 91.2 66.8 158 LFS 1.0902 10.67-13.72 136 23.1 159 LFS 1.0903 1.52-4.57 76.4 30.6 107 LFS 1.0903 4.57-7.62 66.5 59.2 126 LFS 1.0903 7.62-10.67 120 120 LFS 2.0301 0-3.05 12.5 12.5 LFS 2.0301 3.05-6.10 176 366 247 789 LFS 2.0301 6.10-9.14 40.6 249 95.5 385 LFS 2.0302 0-3.05 3.7 20.3 8.8 32.8 LFS 2.0302 3.05-6.10 47.4 18.6 12.0 77.9 LFS 2.0302 6.10-9.14 46.7 36.0 55.8 139 LFS 2.0603 0-4.57 15.3 15.3 LFS 2.0603 4.57-9.14 31.2 72.7 34.2 138 LFS 2.0701 0-3.05 6.8 7.3 14.1 LFS 2.0701 3.05-6.10 12.8 12.9 25.7 LFS 2.0701 6.10-9.14 0,0 63.4 63.4 LFS 2.0702 7.62-8.38 LFS 2.0801 0-3.05 57.2 30.0 17.2 105 LFS 2.0801 3.05-6.10 34.9 29.5 10.4 74.8 LFS 2.0801 6.10-9.14 207 108 45.2 360 LFS 2.0801 9.14-12.19 137 90.9 33.8 262 LFS 2.0801 12.19-13.72 139 26.8 26.0 192 LFS 2.0802 0-3.05 33.4 5.6 9.8 48.7 LFS 2.0802 3.05-6.10 42.0 10.1 11.1 63.2 LFS 2.0802 6.10-9.14 123 127 53.1 303

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Table 4-3 (cont'd) 123 SAMPLE DEPTH (m) 6 mm Passing #40 Remaining on #40 Composite [Hg] ng/g [Hg] ng/g [Hg] ng/g [Hg] ng/g LFS 2.0901 0-3.05 8.1 27.9 36.0 LFS 2.0901 3.05-6.10 62.8 54.7 118 LFS 2.0901 6.10-9.14 80.0 182 262 LFS 2.0902 0-3.05 12.1 16.1 28.2 LFS 2.0902 3.05-6.10 58.5 15.7 74.2 LFS 2.0902 9.14-12.19 200 145 345 LFS 2.0902 12.19-13.72 62.2 193 255 LFS 2.0903 6.10-9.14 68.7 46,7 115 LFS 2.0903 10.67-13.72 65.2 28,9 94.2 LFS 3.0901 3.05-6.10 LFS 3.0901 6.10-9.14 LFS 3.0901 9.14-12.19 LFS 3.0901 12.19-13.72 LFS 3.0902 3.05-6.10 LFS 3.0902 6.10-9.14 LFS 3.0902 9.14-12.19 LFS 3.0902 12.19-13.72 LFS 4.0801 3.05-6.10 21.7 34,9 8,9 65.6 LFS 4.0801 6.10-9.14 94.3 37.5 33.2 165 LFS 4.0802 3.05-6.10 4.7 3.7 36.6 44,9 LFS 4.0802 6.10-9.14 68.7 16.2 118 203 LFS 5.0401 0-3.05 14.3 39.0 25.4 78,7 LFS 5.0401 3.05-6.10 64.7 16.0 21.6 102 LFS 5.0401 6.10-9.14 440 53.1 139 632 LFS 5.0401 9.14-12.19 24.3 2060 79.4 2160 LFS 5.0402 0-3.05 845 193 1038 LFS 5.0402 3.05-6,10 80.6 35.5 116 LFS 5.0402 6.10-9.14 217 226 443 LFS 5.0402 9.14-12.19 42,3 1660 1460 3160 LFS 5.0403 0-3.05 9.6 24.7 0.3 34,6 LFS 5.0403 3.05-6.10 225 73.2 130 428 LFS 5.0403 6.10-9.14 59.7 36.1 41.1 137 LFS 5.0403 9.14-12.19 21,2 46.5 20.8 88,5

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Table 4-3 (cont'd) 124 SAMPLE DEPTH (m) 6 mm Passing #40 Remaining on #40 Composite [Hg] ng/g [Hg] ng/g [Hg] ng/g [Hg] ng/g LFS 5.0701 1.52-2.29 71.6 11.5 83.1 LFS 5.0701 4.57-5.33 LFS 5.0701 7.62-8.38 176 36.0 212 LFS 5.0701 10.67-11.43 139 14.8 154 LFS 5.0702 0-3.05 LFS 5.0703 0-3.05 38.3 20.3 58.6 LFS 5.0703 3.05-6.10 LFS 5.0703 6.10-9.14 39.5 9.8 49.3 LFS 5.0703 9.14-12.19 107 107 LFS 5.0704 0-3.05 164 129 293 LFS 5.0704 3.05-6.10 258 657 915 LFS 5.0704 6.10-9.14 108 162 270 LFS 5.0704 9.14-12.19 465 460 925 LFS 5.0705 1.52-2.29 44.0 12.8 56.8 LFS 5.0705 4.57-5.33 61.3 65.5 127 LFS 5.0705 7.62-8.38 58.1 26.7 84.8 LFS 5.0705 10.67-11.00 47.0 41.4 88.4 LFS 5.0901 1.52-4.57 LFS 5.0901 4.57-7.62 LFS 5.0901 7.62-10.67 LFS 5.0901 10.67-13.72 LFS 5.0902 1.52-4.57 LFS 5.0902 4.57-7.62 LFS 5.0902 7.62-10.67 LFS 5.0902 10.67-13.72 Composite = Z FiC^ where, Fi = weight % of one of the 3 size fractions Ci = Hg concentration corresponding to F^

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125 composite samples. The total Hg in the composite samples ranged from 32.8 to 16800 (ng/g Hg) Total Hg determinations were performed to quantify the amount of Hg in solid waste samples from the Alachua County Landfill which had received MSW from 1973 to the present. Mercury was detected in every sample analyzed. "Hot spots" exist in the Alachua County Landfill which probably result from the decomposition of discrete Hg-containing devices such as batteries and light bulbs that are still finding their way into the landfill today in spite of more stringent recycling efforts Mercury concentrations in Palm Beach County compost samples are shown in Table 4-4. The values ranged from 368 ng/g Hg to 5320 ng/g Hg with a mean of 1100 ng/g Hg. The mean Hg concentration for the compost samples was higher than that for the MSW samples from the Alachua County landfill and the distribution of concentrations exhibited a much narrower range Phase II -Anaerobic Reactor Data Results from Phase I established that Hg did in fact exist in the Alachua County landfill. The next step was to evaluate what happens to Hg during anaerobic digestion like

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126 Table 4 -4. Palm Beach County Mercury Data (1:1 Biosolids to Yard Waste) Sample Hg (ng/g) 1 733 2 914 3 930 4 635 5 405 6 420 7 950 8 894 9 931 10 372 11 399 12 5320 13 1260 14 1680 15 604 16 368 17 450 18 591 19 1580 20 2770 21 1700 22 1610 23 433 24 416 average 1100

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127 that occurring in the Alachua County landfill. Data for the simulated landfill Run 1 are presented in Figures 4-1 to 4-12. Run 2 data are presented in Figures 4-13 to 4-28. These figures show that the pH's, volatile fatty acids, methane, and methane yield acted as expected under anaerobic conditions. The acid-producing microbes were active around day 5 in both runs for all of the reactors as indicated in the pH graphs. The pH was lowered to less than 7 at day 5, but increased to alkaline levels after day 5. The methane content initially started around zero in all cases because at the start of the runs, mostly nitrogen and oxygen were present until anaerobic conditions were established. The methane level steadily increased from day 1 until it leveled off at around 55%. The volatile fatty acids (VFAs) were indicative of the action of the microbes in the system and at day 5 it was evident that there was an increase in the VFA concentrations which then decreased after that time. The methane yield figures are the most important determinants in establishing that the reactors were operating under anaerobic conditions. A document by Chynoweth et al. (1991) provided data on ultimate methane yield values for newspaper and office paper. The value given for newspaper was

PAGE 142

128 9 8.5 e ^ — A 7.51 A J^^ .-^^•^-f-'* — 11 -I s 0< 7 '-^-r-''*^^^ — •— Ri 6 — — R2 5.5 — *— R3 =; 5 10 IS 20 25 3Q 35 Days Figure 4-1. Run 1 (Control) pH vs. Days 8.5 8 7 5l J • 1 7 6.5 6 "-H^-R4 5.5 — 4— R6 5 10 15 20 Days 25 30 35 Figure 4-2. Run 1 (lOOng Hg) pH vs. Days Figure 4-3. Run 1 (2000ng Hg) pH vs. Days

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129 70.00 SO. 00 50.00 SSrst=-^ ''^^ It 40.00 s 30.00 20.00 — •— Rl 10.00 — *— R3 0^ i 5 10 15 20 25 30 35 40 Days Figure 4-4. Run 1 (Control) %Methane vs. Days 60 00 50.00 ^ 5^ Ix^. --*-^ >-* ^ -*^ j'Ta a* \Jf aX \v..*-t f A -^"^^ •^..^^^^ 40.00 >rv V^ ^A\/ 'V^^"^ 30.00 1 ;6^ ^ # 20.00 / ^ ^— R4 J --R5 10.00 ( — *— R6 0.00 1 l^ 5 10 15 20 Days 25 30 35 40 Figure 4-5. Run l (lOOng Hg) %Methane vs. Days 7" r
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130 H 1 to > 2000 1500 1000 500 o' ( jV^ :^ ^^^ -*— R3 ^ ) 5 10 15 Day 20 25 30 35 Figure 4-7. Run 1 (Control) Volatile Fatty Acids vs. Days 1200 1000 A -•-R4 itr\ --R5 1 800 600 /I'Aw. -4— R6 a ii rW^N^ f^ i > 400 200 Jy v-< ^r ^ ^ 1 ^=^ ^-^ 1 c r4* ) 5 10 15 20 25 30 35 Day Figure 4-8. Run 1 (lOOng Hg) Volatile Fatty Acids vs. Day ^ < > 10 15 20 25 35 Day Figure 4-9. Run l(2000ng Hg) Volatile Fatty Acids vs. Day

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131 10 15 20 Days 25 30 35 40 Figure 4-10. Run 1 (Control) Methane Yield vs. Days -^ 0.12 D U a •a to (A > < E 0.1 0.08 0.06 2 .£ 0.04 >o c O S 0.02 10 15 20 Days 25 30 35 40 Figure 4-11. Run 1 (100 ng Hg) Methane Yield vs. Days

PAGE 146

132 ^^ 0.14 D 0) a •a 0.12 n CO > U.I O) m CO 0.08 < E 06 a o > 0.04 w c n r 0.02 0) S 10 15 20 Days 25 30 35 40 Figure 4-12. Run 1 (2000 ng Hg) Methane Yield vs. Days

PAGE 147

133 8 5 8 ^ dti^ ^^=^ 1 7.5I '^^^^v z*"::^ ^ r-^* ^ --^ 5 Di 7 6.5 e ^ ^ — •— Rl 5.5 -•-R2 5 -*-R3 3 2 4 6 8 10 12 14 Days Figure 4-13. Run 2 (Control) pH vs. Days 8 5 8 ^ ^ 1 -^ 7 5 '^^'•^ r^ j^ 7 ^^*^ .^ 6.5 6 5.5 5 N / -— R5 ) 2 4 6 Days 8 10 12 1 4 Figure 4-14. Run 2 (lOOng Hg) pH vs. Days

PAGE 148

134 Figure 4-15. Run 2 (lOOOng Hg) pH vs. Days 8.5 -, 8 1 _,^ -7-* 1 7 5" --/ /^ --\ >
PAGE 149

135 0) c n) X V a S 60 50 40 30 20 10 0| c .-5=*=^*=:^^!^^^^ ^ Ht^-^ //^ — ^~~~~4^ ^ ni --Ri — *— R3 ) 2 4 6 Days 8 10 Figure 4-17. Run 2 (Control) % Methane vs. Days Figure 4-18. Run 2 (lOOng Hg) % Methane vs. Days

PAGE 150

136 Figure 4-19. Run 2 (lOOOng Hg) % Methane vs. Days Figure 4-20. Run 2 (2000ng Hg) % Methane vs. Days

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137 10000 1 /\ — •— Rl 9000 / \ --R2 8000 7000 ^ \ -A-R3 l-l ''**-,,,,,^^ 1 6000 ^^ 5000 \ 01 \ '< 4000 ~"~~~--0 v^ \ > 3000 2000 1000 J ^ k 1 ( ^ ) 2 4 6 8 10 12 14 Day Figure 4-21. Run 2 (Control) Volatile Fatty Acids vs. Days 9000 8000 7000 -•-R4 — — R5 0) 5 6000 5000 a 4000 > 3000 2000 1000 1 -*-..4^ ^ 1 t ) 2 4 6 8 10 12 1 4 Day Figure 4-22. Run 2 (lOOng Hg) Volatile Fatty Acids vs. Days

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138 7000 6000 5000 ^^ iH ^^ 01 e 4000 -#-R6 3000 \, \ — — R7 \ \ > 2000 1000 1 \^ J ( • ^l 3 2 4 6 8 10 12 14 Days Figure 4-23. Run 2 (lOOOhg Hg) Volatile Fatty Acids vs. Days 6000 5000 ...... r-l 4000 1 3000 ta '< tn > 2000 \ — • — R9 1000, x"^ ^v. C ^ ^^. 1 ) 2 4 6 8 10 12 1 4 Day Figure 4-24. Run 2 (2000ng Hg) Volatile Fatty Acids vs. Days

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139 £ 0-25 o 1 0.2> G) ^ 0.15CO < E ^ 0.1 > 2 0.05 re ^ E oJ ( ^^<:==^^^ ^^^.-^ — .,1^^^ Ri V^^^ R2 R3 ^y^ ) 5 10 Days 15 20 Figure 4-25. Run 2 (Controls) Methane Yield vs. Days ^^ 0.2 •a 0) •a 0.18 •a n 0.16 W > 0.14 D) 1^ 0.1? ro < E 0.1 o 0.08 0) >U.06 0) c 0.04 re 0.02 s 10 Days 15 20 Figure 4-26. Run 2(100 ng Hg) Methane Yield vs. Days

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140 ^ 0.25 § o 0.2 at ^ 0.15 n < E 2 0.1 = 0.05 n *-< s oJ c _,„>-^""^ /-^^ ^^^^ — y(^^ X^ ^ R6 R7 Poly, (R6) y ) 5 10 Days 15 20 Figure 4-27. Run 2(1000 ng Hg) Methane Yield vs. Days Figure 4-28. Run 2(2000 ng Hg) Methane Yield vs. Days

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141 0.100 mVKg "VS added" and 0.369 mVKg "VS added" for office paper. Since 90% of the SW in the reactors consisted of newspaper and office paper, these are the values that were used for the determination of the expected yields of the reactors. The methane yield for all of the reactors fell within the range 0.100 0.369 mVKg "VS added". The methane yield results for Run 2 were higher than that for Run 1 and this is attributed to silicone being used as a sealant along with the teflon tape. In Run 1 only teflon tape was used. This section includes all of the reactor data excluding the Hg. The raw data are provided in Appendix B. Phase II-Mercurv Data The raw data from the total Hg analyses performed on the I solid, liquid, and gas phases from the reactors are presented in Appendix C. The solid phase was the SW, the liquid phase was the leachate, and the gas phase included the contributions from the activated carbon traps and the tygon tubing leading from the reactors to the carbon traps. Table 4-5 shows a compilation of the Hg mass balances. The data from the 1 individual runs are presented in Figures 4-29 to 4-35. I In doing the Hg mass balances, it must be noted that most of the levels recovered did not add up to 100% of the Hg that 1

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142 0) U 0) u a:; >1 u 0) ID >* X! Eh Percent Recovery o O H 00 o H o H ID O CN CTi o CO o H o CN CD 00 CO (J) Gas Tygon (ng) O O o H 00 CN U5 H U3 co H 00 o O O CO H H CO CN 00 00 rH ^ Gas Act. C. (ng) O O o 00 H 00 00 H CN O o O O in CN CN o LD o CN m Leachate (ng) O o o CN H o O O O o o O O o o CO CN o o o Solids* (ng) 00 00 '^ in in O H O 00 H o CTl H CO (T\ O rH in >* m 00 o rH 00 o cr. 00 H o H CO H Total Hg Added (ng) o o o o o H o o H O O O o o CN O O O CN o o o CN o O O o o o o H o o o o o o rH o o o CN o o o CN u -U u dj Pi H CN m ^ if) KD r00 m rH CN n >:}< in U3 rCO a\ CI C! H CN T3 0) > E a) Sh CD m m K 5 tB iJ c nj x: 00 jj m -U rri xi S CD -H CO 0) QJ S m >. >. rH M m t3 c (d 4H rH rH m m e m in Sh o tn 4H LD JJ rH -H E 1 -rl J IJD C 0) T) -H d 4J rH U U (D c 4J H (1) Q -U -rt c C) jz: Cfi JJ 0) a; s Q X K

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Ol c X 100 90 -80' 70 60 50 40 30 20 10 ./ |R1 R2 DR3 Solids Leachate Gas 143 Figure 4-29. Run 1 Hg content in 3 phases (control! 70 60 50 o) 40 ^ 30 20 10 j^m ^H R4 BRS D Pfi 1 ^BB ^^1 1 1 1 I 1 1 1 1 M_ / 1 F j^ta^Solids Leachate Gas Figure 4-30. Run 1 Hg content in 3 phases (100 ng Hg) 2000 n (^ 1800 1600 1400 X 1200 1000 800 600 400 ,-' 200 / / IR7 BRS nR9 Solids Leachate Gas Figure 4-31. Run 1 Hg content in 3 phases (2000 ng Hg)

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144 Ol X 100/ 90^ ,,-' 80R1 MKl DR3 706050403020104H^^ 0! w .^ma^m^' .^mm^^' Solids Leachate Gas Figure 4-32. Run 2 Hg content in 3 phases (control) I R4 R5 Solids Leachate Gas Figure 4-33. Run 2 Hg content in 3 phases (10 ng Hg)

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145 |R6 BR? Solids Leachate Gas Figure 4-34. Run 2 Hg content in 3 phases (1000 ng Hg) Solids Leachate Gas Figure 4-35. Run 2 Hg content in 3 phases (2000 ng Hg)

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146 was originally spiked into the reactors. Part of this can be attributed to experimental error. Another part can be attributed to the SW that was removed from the reactors after the completion of the run for the purpose of volatile solids analysis. Run 1 In Run 1, reactors 1, 2, and 3 served as the control reactors and did not receive any Hg. No Hg was detected in either the leachate or gas phase of the control reactors. However, low levels of Hg (8, 8, and 4 ng) were observed in the SW itself, but none of these values exceeded the MDL of 8 ng. Reactors 4, 5, and G were spiked with 100 ng of Hg and 30, 36, and 34 ng Hg, respectively, were found in the carbon traps and tygon tubing leading to the traps. This accounted for 30%, 36%, and 34% of the Hg added being volatilized. Twelve ng Hg were detected in the leachate in reactor 4, but none was found in reactors 5 and 6 The rest of the Hg was detected in the solid phase, i.e., 55%, 53% and 67% respectively. The recovery of Hg was 97, 81, and 101% for reactors 4,5, and 6, respectively, not counting that in the 6 g of SW that had been removed for volatile solids analysis.

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147 Reactors 7, 8, and 9 were spiked with 2000 ng of Hg and 50, 48, and 42 ng Hg, respectively, were found in the carbon traps and tygon tubing. This accounted for 2.5%, 2.4%, and 2.1% of the added Hg being volatilized. There was no Hg detected in the leachate in any of these reactors. The Hg detected in the SW was 1760, 1480, and 1910 ng Hg for reactors 7, 8, and 9, respectively. This accounted for 88%, 74% and 95% of the total Hg added, again not including the 6 g of SW that were removed for volatile solids analysis. Run 2 Reactors 1, 2, and 3 again served as controls and received no Hg. In these reactors, there was no Hg detected in the gas or leachate phases, although low levels were again detected in the SW, but the values were at or below the MDL, being 8, 9, and 10 ng for reactors 1,2, and 3, respectively. Reactors 4 and 5 were spiked with 100 ng Hg and 37 and 48 ng Hg, respectively, were found in the carbon traps and tygon tubing. This accounted for 37% and 48% of the Hg added being volatilized. No Hg was detected in either of the leachates. The rest of the Hg was detected in the solid phase, i.e., 61% and 47% respectively. The recovery of Hg was 108 and 95% for reactors 4 and 5, respectively, not including 10 g of SW that was removed for volatile solids analysis. The distribution of

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148 Hg between solid and gaseous phases was similar to that observed in Run 1 Reactors 6 and 7 each received 1000 ng of Hg and 126 and 83 ng Hg, respectively, were found in the carbon traps and tygon tubing. This accounted for 12.6%, and 8.3% of the added Hg being volatilized. Reactor 6 contained 28 ng Hg in the leachate, but reactor 7 did not show any Hg in the leachate. The Hg detected in the SW was 836 and 810 ng Hg for reactors 6 and 7, respectively. This accounted for 84% and 81% of the total Hg added. The recovery of Hg was 98 and 88%, again not including the 10 g of SW removed for volatile solids analysis. Reactors 8 and 9 each received 2000 ng of Hg and 48 and 136 ng Hg, respectively, were found in the carbon traps and tygon tubing. These levels represented 2.4% and 6.6% of the total Hg being volatilized. There was no Hg detected in the leachate of either reactor. Reactor 8 contained 1890 ng Hg in the SW which was 94% of the total Hg added and reactor 9 had 1810 ng Hg which was 90% of the total. The mass balance was 97% for both reactors, not including that in the 10 g of SW that was taken for volatile solids analysis.

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CHAPTER 5 DISCUSSION Phase I Hg in the Alachua County Landfill In the Phase I experiments, the data showed that there was Hg in the Alachua County Landfill. The average concentration for composite samples was 176 ng/g with over half of the values being below 150 ng/g. Figure 5-1 shows the frequency distribution of Hg obtained in composite samples from the Alachua County Landfill. From the figure, it is evident that most of the samples lie within 100-150 ng/g with fewer samples appearing at the other concentrations. However, five values exceeded 1000 ng/g with one exceeding 16,000 ng/g. These values are important in that they are considerably higher than background Hg concentrations in Florida soils which range from 0.5 ng/g to 43 ng/g (Ma et al, 1997) However, they are considerably lower than current target clean-up levels set by Florida DEP for soil. The target value for Hg in residential soil is 23,000 ng/g and in industrial 149

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o o o CO o Q_ o O o o o IT) CO o CO I 150 ij CQ I o OJ CJ 3 O u m H •H: 4-1 IX! >i a, d o I o CD o •H 01 en 3 CQ C (C CO O CM CD CNJ LO Td '^ CD O Aouenbejj in o u o o H I in :(U M 01 Cm

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151 soil is 480,000 ng/g (Florida Department of Environmental Protection, 1995) The maximum concentration for Hg in sewage sludge as set forth in CFR 40 Part 503 is 57,000 ng/g (Code of Federal Regulations, 1997) Thus, if the Alachua County Landfill were to be reclaimed and the SW residue taken out and land applied, it would meet req[uirements for Hg that are much more stringent than is currently required for sewage sludge. Palm Beach County compost data is also included in Figure 5-1. This compost consisted of a 50:50 mixture of biosolids to yard waste. These are consistently higher than the MSW Hg data, but are still well below the regulatory limits. LFS 1, LFS 2 and LFS 5 are sites on a fairly new portion of the landfill and therefore might be expected to contain lower levels of Hg due to modern recycling efforts. A comparison of Hg concentrations found in LFS 2 and LFS 4, which differed only in age and the presence of a liner, showed that Hg concentrations were actually lower at LFS 4 than at LFS 2 but the differences were not significant at the 5% probability level. LFS 4 received MSW beginning in 1973 while LFS 2 began receiving MSW in 1988, the year that disposal of batteries in MSW landfills was prohibited. Although there have been recycling efforts for household batteries that

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152 contain Hg since 1988, Hg is still making its way into landfills via other sources such as Hg light switches, electric lighting, paint residues, fever thermostats, pigments, dental uses, and special paper coating (Bureau of Solid and Hazardous Waste, 1995) Since Hg is present in MSW from these sources there is a need to better understand the fate of this Hg The Hg levels determined in the SW fractions in Table 4-1 ranged from 12.5 ng/g to greater than 70,000 ng/g Hg The Hg levels in the composite samples ranged from 32.8 ng/g to 16,800 ng/g Hg Table 4-2 is missing Hg concentration data for some composite samples. These missing data resulted from the fact that dry weight data for some size fractions were not available. In addition, for some core samples, one or more fractions were missing and could not be analyzed for Hg However, in spite of the missing fractions and weight data, results for composite samples were calculated for 93 core samples. This number should be sufficient to say something about Hg concentrations in the Alachua County Landfill. It is assumed that these concentrations would also be representative of the Hg concentrations that exist in landfills for other municipalities around the country. Unfortunately, a search of

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153 the literature did not yield Hg data from other landfills to allow validation of this assumption. Samples were taken from all five areas of the landfill i.e. LFS 1 to LFS 5. LFS 1 and LPS 2 were among the more recent additions to the landfill, with LFS 1 having leachate recycling and LFS 2 acting as a control without leachate recycling. Thus a comparison of Hg concentrations in LFS 1 and LFS 2 would show the effect of leachate recycling. As a result of leachate recycling, LFS 1 would have enhanced microbial activity due to the higher moisture content than would LFS 2 The enhanced microbial activity would accelerate the degradation rate of the SW. Differences between LFS 1 and LFS 2 were evaluated at each depth and the results are shown in Table 5-1. It must be noted that the 3 05 m (0 10 ft) depth in LFS 2 was equivalent to the 1.52 4.57 m (5 15 ft) depth in LFS 1 with the other depths corresponding in like fashion. The landfill workers had added fresh material to LFS 1 after the sampling had started and markers were placed, so the first 1.52 m were discarded in order to be consistent with the depths in LFS 2 The overall means for LFS 1 (155 ng/g) and LFS 2 (153 ng/g) were almost identical and not significantly different.

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154 Table 5-1. Mean Hg concentrations in composite samples for LFS 1 and LFS 2 Alachua County Landfill Depth (m) LFS 1 (ng/g Hg) LFS 2 (ng/g Hg) 3.05 152 (11)^ 36 (8)* 3 05 6 10 142 (9) 174 (7) 6.10 9.14 167 (8) 221 (8) 9.14 12.19 168 (4) 234 (3) Overall (w/in a row) 155 (32) 153 (26) Means within a row are significantly different at the 1% probability level based on student's two-tailed t-test. ^ Number in parentheses are the number of observations for that treatment

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155 Thus leachate recycling did not affect the total amount of Hg remaining in the solid residues. Since LFS 1 and LFS2 were both from lined areas, the only way that they could differ in Hg concentration is if Hg volatilization was different. Thus, leachate recycling did not appear to affect Hg volatilization However, leachate recycling did affect the distribution of Hg with depth. At LFS 1, the Hg concentrations were fairly consistent with depth, ranging from 152 ng/g in the top layer to 168 ng/g Hg in the bottom layer. These differences were not significant (Fig. 5-2) Thus, leachate recycling resulted in a more even distribution of Hg from the top layer (1.52 4.57 m) to the bottom layer (10.67 13.72 m) Confidence intervals (95%) were calculated using the formula, C.I. (95%) = to. 05 s.e. where to.os represents the two-tailed test at a probability value of 0.05 for a given value of degrees of freedom. The to. 05 values were obtained from Snedecor and Cochran (1968) The s.e. represents the standard error and this is determined using the standard deviation divided by the degrees of freedom. The error bar on the 9.14 12.19 m depth for LFS 2 (Fig. 5-1) is very large for two reasons. First, the mean

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TCM CO in _l _l Q 156 D. Q IM J2 a X) X! OJ H w c -H J-1 m >-i jj c (D U C Dl w c I to Dl -H (6/6u) suo!;ej)usouoo 6h ue9|/\|

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157 consisted of only three observations, yielding a small number of degrees of freedom and a relatively high to 05 value. Second, the three observations ranged from 94 to over 3 00 ng/g, resulting in a large standard deviation. At LFS 2, there was a trend of increasing Hg concentration with increasing depth. The averages that were computed for the depths also exhibited a considerable range, i.e. 3 6.5 ng/g at the top to 234 ng/g Hg at the bottom. Confidence intervals (95%) for each depth (Figure 5-2) showed that the top layer (0 3.05 m) at LFS 2 was significantly lower in Hg than the lower three depths The lower three depths did not differ significantly in their Hg concentrations. The trend of increasing Hg concentrations with depth was also evident at LFS 3 and LFS 4 where Hg concentrations two to three times higher than that in the top layer could be found in lower layers (Table 4-2) This suggests that, in the absence of leachate recycling, Hg tends to leach and settle in the deepest layers. Leachate recycling may be beneficial in evenly distributing Hg, and maybe other heavy metals within the residue, so that the highest concentrations are not at the lowest layers. These data also suggest that Hg does tend to leach downward through a landfill

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158 and that soil and groundwater contamination with Hg could pose a threat in the case of unlined landfills. Statistical comparisons were also perfo2rmed for the 3 05 9.14 m depth at LFS 2 and LFS 4 (Table 5-2) This was the only depth for which Hg data were available at LFS 4. These two sites (LFS 2 and LFS 4) differed with respect to age and neither was involved with leachate recycling. LFS 4 is the oldest site at the landfill and had been in operation since 1973. It was unlined, whereas LFS 2, a site that had been receiving MSW since 1988, was lined. Thus, in addition to volatile losses that could occur at both sites, LFS 4 could also have lost Hg by leaching over the years it had been in use. LFS 4 was capped in 1985 to prevent leaching losses, but was open from its initiation in 1973 until then. A comparison of LFS 2 and LFS 4 would provide some indication of the effect of the age of the landfill with the Hg concentrations. The Hg concentrations in the composite samples at LFS 4 were lower than those at LFS 2. However, a t-test showed that the difference between the two means in this table was not significant. In a recent review of leachate attenuation at landfills, Christensen et al.(1994), concluded that heavy metals, including Hg, do not constitute a significant

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159 Table 5-2. Mean Hg concentrations in composite samples for LFS 2 and LFS 4 Alachua County Landfill Depth (m) LFS 2 (ng/g Hg) LFS 4 (ng/g Hg) 3.05 9.14 200 (4)1 120 (4)* Means were not significantly different at the 5% probability level based on student's twotailed t-test. 1 Number in parentheses are the number of observations for that treatment

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160 pollution problem at landfills due to low concentrations in the leachate and strong attenuation by sorption and precipitation. If this is true, then leaching would not be a factor in this comparison. Thus, any differences in Hg concentration in the solid residue could be attributed to differences in Hg volatilization. Assuming similar volatilization losses between LFS 2 and 4, the data suggest that the implementation of recycling laws has had little impact on Hg concentrations in the Alachua County Landfill. This may result from the fact that many Hg-containing devices are still making their way into landfills. Hg concentrations were computed for the fractions remaining on 6 mm sieve, passing #4 sieve, and remaining on #40 sieve and are shown in Table 5-3. Overall means for these three fractions were 296 ng/g, 253 ng/g and 486 ng/g for that remaining on a 6 mm, passing the #40 sieve, and remaining on the #4 seive, respectively. In computing the means, extremely high numbers, i.e., > 10,000 ng/g, were excluded. This amounted to three numbers being excluded from the entire data set. These were 21,600 ng/g for the fraction remaining on 6 mm sieve, 70,000 ng/g and 10,400 ng/g for the fraction remaining on #40 sieve. It was assumed that these extremely

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161 high values resulted from sampling directly into a battery or some other Hg containing device and were not representative of the concentrations typically found in SW (see Fig. 5-1) Table 5-3 shows that the largest concentrations of Hg were found at LFS 3 No data are shown for the fraction passing a #40 sieve at this site because these samples were missing from storage. Mercury concentrations for the other two fractions are about 4 and 7 times those found at LFS 4, the most appropriate site of the five for comparison with LFS 3 (i.e., both being approximately the same age and both being unlined sites) No explanation can be given for this finding. However, it has been suggested (Dr. J.F.K. Earle, Personal Communication) that perhaps this site was the recipient of medical and biological wastes that the landfill routinely accepted at the time LFS 3 was being filled. These wastes may have contained higher concentrations of Hg than conventional MSW. In the discussion that follows, one should keep in mind that something was definitely different about the MSW that was loaded into LFS 3 than that for the other four sites. In the fraction remaining on the 6 mm sieve, Hg concentrations ranged from 187 ng/g at LFS 4 to 761 ng/g at LFS 3. In the passing #40 fraction, the values ranged from 65

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162 Table 5-3. Mercury Means in different fractions at different sites. Fractions LFS 1 LFS 2 LFS 3 LFS 4 LFS 5 Rem. on 6 mm 289 208 761 187 217 Passing #40 133 164 N/A 65 427 Rem. on #4 396 440 1085 163 502 All units are in ng/g

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163 ng/g at LFS 4 to 427 ng/g at LFS 5. In the remaining on #40 fraction, the values ranged from 163 ng/g at LFS 4 to 1085 ng/g at LFS 3. In general, lower Hg concentrations were detected at each site in the fraction passing #40 sieve which was expected since that fraction contained mostly sand and had the lowest content of volatile solids (Table 5-4) The lower organic matter contents in these fractions probably accounted for the lower amounts of Hg that were bound to these residues At all sites, except LFS 4, the highest Hg concentrations were in the fraction remaining on the #40 sieve. Since the fraction remaining on the 6 mm sieve generally had a higher volatile solids content than that remaining on a #40 sieve (Table 5-4) this suggests that the organic matter in the latter fraction was more reactive toward Hg. This could have been due to a larger surface area and/or a more decomposed state for organic matter in the fraction passing the 6 mm sieve compared to that remaining on the 6 mm sieve. From Table 4-1, the usual trend noted in the fractions at the sites was increasing Hg with increasing depth, except at LFS 1 which had leachate recycling. Leachate recycling is believed responsible for distributing Hg throughout the depths, so that there was no concentration effect at the

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164 Table 5-4. MSW Sample Volatile Solids Content Average by Round and Area % VS % VS %VS Above 6 mm Above #4 Pass #40 Round 3 LFSl 72.4% 53.7% 5.5% LFS2 81.3% 50.9% 4.1% LFSl* 69.9% 48.4% 6.2% Round 4 LFSl 80.7% 41.0% 7.0% LFSl* 77.4% 38.8% 6.6% LFS5 87.2% 64.4% 5.6% Round 6 LFSl 89.2% 58.1% 6.2% LFS2 90.2% 59.6% 9.3% LFSl* 89.5% 49.4% 5.8% Round 7 LFS2 45.9% 2.5% LFS5 57.6% 4.1% Round 8 LFSl 41.3% 3.3% LFS2 54.6% 3.7% LFSl* 34.9% 3.4% LFS4 41.6% 10.3 Round 9 LFSl 64.7% 36.7% 3.5% LFS2 78.5% 54.4% 4.6% LFSl* 64.7% 35.3% 4.0% LFS3 77.0% 31.1% 4.8% LFS5 77.7% 61.4% 4.4% *Combined fraction of material above 6 mm screen and material above the #4 sieve. (Miller et al 1996)

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165 lowest depth. The sites without leachate recycling indicated that there was possibly some leaching of Hg taking place. If that is the case, then for unlined landfills, there is a possibility that Hg could be getting into the groundwater system. Christensen et al. (1994) reported Hg concentrations in landfill leachate for landfills of less than 20 years of age. These concentrations ranged from 0.05 to 160 jUg/L. Thus, the potential for Hg to leach through landfills is quite variable but could certainly occur with Hg concentrations at the upper end of this range. The volatilization of Hg from the Alachua County Landfill to the atmosphere was not evaluated during Phase I of this study. However, the potential for Hg volatilization during anaerobic digestion of MSW was evaluated during Phase II. Phase II An important part of the Phase II experiments was to determine Hg levels in the gaseous phase. Quite a few authors had indicated that impregnated activated carbon was capable of absorbing elemental Hg (Vidic and McLaughlin, 1996; Krishnan et al. 1994; and Moffitt and Kupel, 1971). It was stated that sulfurimpregnated activated carbon enhanced Hg removal over those nonimpregnated types, due to HgS formation on the

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166 carbon surfaces. This chemisorption process was enhanced by temperatures between 25C and 90C, resulting in increased removal of elemental Hg. It was stated that at a temperature of 140C, there was a decrease in adsorptive capacity, thus a decrease in HgS formation (Vidic and McLaughlin, 1996; Krishnan et al 1994) The problem with these reports was that they did not indicate how to desorb the trapped Hg from the impregnated activated carbons. Desorbing trapped Hg was imperative in this study in order to quantify the amount of Hg volatilized during anaerobic digestion. Krishnan et al. (1994) used high temperatures to desorb the trapped Hg from the carbon while Moffitt and Kupel (1971) stated that the impregnated charcoal was analyzed by AAS The Simpregnated activated carbon used in the method development and experiments in Phase II was obtained from Norit Americas and they state that it is capable of trapping 99.9% of the elemental Hg in a gas stream. This proved to be true in the initial experiments of this study that were performed to determine trapping efficiency. It did trap all of the elemental Hg that was passed through it, but the problem that arose was how to desorb the Hg from it. The USEPA Method 7471 for determination of total Hg in solid material was not

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167 sufficient to enable desorption of the trapped Hg. Therefore, a modification of this method was developed to recover the trapped Hg from the carbon. This modification is shown in Figure 3-10. The development of this method was very important for the Phase II experiments because it was important to quantify the Hg that existed in the gaseous phase. The modified method yielded >98% recovery of the Hg that was sorbed to the Simpregnated activated carbon. Additional experiments that were performed involved the setup of a system to test two hypotheses: (1) that tygon tubing may be able to trap elemental Hg and Hg^* at 50C, and (2) that Hg^* in the nitrate form may be volatilized at 50C and trapped on the Simpregnated activated carbon. The Merck Index (1989) states that HgClj volatilizes unchanged at 300C and volatilizes appreciably at 100C. Thus it was important to know if Hg^"" in the nitrate form could volatilize at 5(yc. This was a 60 hour experiment that was set up in an incubator room that operated at 50C. There were four flasks altogether, hooked up to S-impregnated activated carbon traps and also attached to a pump that had a total flow rate of 3 mL/min. All four flasks contained DDDI water and were spiked with 100

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168 ng Hg^* as HgCNOjjj. However, two flasks also received SnClj to reduce the Hg^"" to Hg It was found that essentially all of the Hg in the two flasks that received SnClj was trapped on the carbon and there was none in the tygon tubing, thus indicating that even at 50C, tygon tubing is not able to trap elemental Hg. In the case of the two flasks that did not contain SnClj, no Hg was found in the tygon tubing, but 3 0-35% of the Hg was found in the carbon traps. This proved that Hg^* has the ability to volatilize directly at 50C, and that it is not trapped in the tubing. Thus, any Hg found in the tygon tubing cannot be attributed to volatilization of elemental Hg or Hg^*. The results of this experiment show that if Hg is found in the tygon tubing following the reactor experiments, then other forms of Hg were probably being volatilized. Tygon tubing has the ability to trap organic vapors (Dr. R.D. Rhue, Personal Communication) therefore in the case of Hg, the forms that would be expected to be trapped in the tubing would be the organic forms, possibly including dimethylmercury Also, the death in 1997 of Dr. Karen Wetterhahn resulted from dimethylmercury penetrating her latex gloves indicating that polymertype materials can absorb dimethylmercury.

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169 Phase II was conducted in order to determine the fate of Hg in landfills and involved the use of anaerobic digesters to simulate the processes occurring in a landfill. These digesters allowed precise control over experimental conditions as well as a mass balance in determining the fate of Hg. The mixture of the SW used in the simulated landfill systems or reactors was assumed to be representative of the normal waste stream after recycling (Owens and Chynoweth, 1993) Paper contributes 3 to 50% of MSW and is normally considered the largest component (US Congress, 1989) while food waste represents roughly 10% of MSW (Tchobanoglous, 1993). The biodegradable portions of MSW would include paper and food waste. In the report by Owens and Chynoweth (1993), biochemical methane potentials (bmp) were determined for yard waste, office paper, printed and unprinted newspaper and food packaging. The ultimate methane yields that were determined from these components provided a basis on which to evaluate the digester results in this study. For office paper, unprinted newspaper, and printed newspaper, the ultimate methane yields were 0.369, 0.084, and 0.100 mVkg, respectively. Using these numbers as a basis for ideal behavior in anaerobic digesters, it was important for the

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170 methane yields to fall within these parameters. Figures 4-10, 4-11, 4-12, 4-25, 4-26, 4-27, and 4-28 show this to be true. There are four major microbial steps involved in the anaerobic digestion process, namely 1) Hydrolysis, 2) Fermentation, 3) Acidogenesis, and 4) Methanogenesis Hydrolysis involves hydrolytic bacteria breaking down complex organic compounds (carbohydrates, lipids, and proteins) into simpler units known as monomers. Fermentation involves fermentative bacteria acting on these monomers to produce organic acids, alcohols, neutral compounds, hydrogen, and carbon dioxide. During acidogenesis, some of the fermentation products are converted to hydrogen, carbon dioxide, and acetate by acid-producing bacteria. Specifically, hydrogen producing acetogenic bacteria are responsible for converting hydrolysis and fermentation products to acetate, hydrogen, and carbon dioxide, while homoacetogenic bacteria ferment hydrogen, carbon dioxide, and formate to acetate. In the fourth and final stage called methanogenesis, two types of methane producing bacteria are active. One converts acetate to methane and carbon dioxide and the other reduces carbon dioxide to methane (Earle et al, 1991; Chynoweth et al 1984)

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171 Therefore, to determine whether or not a digester is operating as it should under anaerobic conditions, it must exhibit these steps. The two most notable steps would be the acidforming step which would be represented by a drop in pH and the methanogenic step which would be represented by a steady increase in methane content until there is a stabilization at around 55%. The data in Figures 4-1 to 4-28 indicated that the reactors in both runs were behaving as expected under anaerobic conditions. In Run 1, the pH data showed a decrease from originally 7.5 to 6.5 at roughly day 5 indicating that the acid forming bacteria were active at that time. The control, 100 ng, and 2000 ng Hg reactors showed consistent data with a decrease in pH from above neutral to about 6.6. The methane content in the gas phase appeared to peak at 55% at around day 7 and then was maintained for the following days. At 100 ng and 2000 ng Hg, there was some fluctuation in the methane content at the later days I In Run 1, the VFA peaked at day 5 to about 18 mg/L for the control reactors and 1100 mg/L for the 100 ng Hg reactors and tapered off for the following days. The 2000 ng Hg i reactors peaked to 24 mg/L at day 5 for Reactor 7 and 10 i .1 1

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172 and 1400 for Reactors 8 and 9, respectively. Additional peaks were noticeable at other days for unknown reasons. Methane yield graphs (Figures 4-10 to 4-12) were all quite consistent and indicated that the reactors were operating properly. They yielded greater than 0.100 mVKg "VS added" which indicated that they were producing adequate methane yields In Run 2, the pH's in all of the reactors appeared to decrease from 7.6 to 6.6 at day 3 to day 4. This implied that the acidforming bacteria came into play one day earlier compared to Run 1. The percent methane (Figure 4-17 to 4-20) also peaked at about 53% at an earlier day, i.e. day 4. The VFAs in Figure 4-21 increased to 2 500 mg/L at day 3 and then to greater than 700 mg/L at day 7. In Figure 4-22, the VFAs increased to 2 000 mg/L at day 3 and then to greater than 6 000 mg/L at day 8. In Figure 4-23, there was a marked increase in the VFAs at day 3 to about 2 000 mg/L and then increased between days 6 and 8 to greater than 500 mg/L. In Figure 424, the VFAs showed an increase to 1800 mg/L at day 3 and then an increase to greater than 4000 mg/L at day 6. The methane yields for the reactors in Run 2 were consistently higher than those for Run 1. In Run 2, silicone, in conjunction with teflon tape, was used to seal all fittings and connectors.

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173 while in Run 1 only teflon tape was used. The silicone, along with the teflon tape seemed to make the reactor system more air-tight. The gas production in Run 2 was higher than Run 1 even though both methane yields fell within acceptable limits, i.e. > 0.100 mVKg "VS added". In Figure 4-27, the methane yield for reactor 7 (R7) was lower than that for reactor 6 (R6) R7 was nearly 0.25 mVKg "VS added" whereas R6 was 0.20 mVKg "VS added". Lab notes showed that air was accidentally let in. Figure 4-28 showed a lower methane yield in R9 than R8 RB showed a value of 0.23 mVKg "VS added" whereas R9 showed a value of 0.17 mVKg "VS added". Although R9 had a lower value than R8 it was still within the acceptable limits stated previously. R9 had several problems including leaky septum. These problems were not "fatal" to the system, but were considered nuisances. To help explain the data, it is necessary to understand the role of microorganisms in these systems. During anaerobic digestion, the same microorganisms responsible for producing methane may influence the transformations of Hg in this system. In this system, the main transformations which occur are reduction and methylation. Mercury must be in the divalent form for methylation to occur. Mercury can be

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174 methylated by anaerobic microorganisms such as those listed above, to form either monoor dimethylmercury The transformations are shown below: i Hg^^ + RCH3 > CHjHg* + R(1) 2CH3Hg* + HjS > (CH3)2Hg + HgS + 2K* (2) The formation of mono and dimethylmercury depends on the Hg concentration and the pH of the system. Monomethylmercury (1) is formed under acidic pH (low pH) conditions and in the event that Hg concentrations are high. Therefore, the bacteria responsible for producing acetate during methane fermentation would be creating favorable conditions for monomethylmercury production. Dimethylmercury (2) is normally formed under alkaline or neutral conditions when strong complexing agents such as hydrogen sulfide (HjS) are present or when Hg concentrations are low. As a result, dimethylmercury may be observed in an anaerobic digestion system. The reaction rate constant for synthesis of monomethylmercury is 6000 times larger than that for dimethylmercury. Thus, in most environmental systems only about 3% of methylmercury is dimethylmercury (Regnell, 1990) In the case of anaerobic

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175 digesters (reactors) this percentage may be greater since the system exists under mostly alkaline conditions which are more conducive to formation of dimethylraercury Lindqvist and Schroeder (1989) implied that transformations of divalent Hg to monomethylmercury by bacteria are mechanisms of detoxification and excretion. Gilmour and Henry (1991) suggested that sulfur-reducing bacteria may also produce methylmercury, but due to the fact that a lot of sulfur is not present in an anaerobic digestion system, it could be stated that mostly methanogens and a small amount of sulfur reducers are responsible for methylmercury production. The sulfur that is present, especially in the form of hydrogen sulfide would transform some of the divalent Hg to mercuric sulfide (HgS) which is extremely insoluble and would precipitate (Gilmour and Henry, 1991) The last transformation that can occur is the reduction of monovalent and divalent Hg to elemental Hg which can be lost to the atmosphere. Also, some microorganisms contain a detoxification mechanisms for methylmercury that reduces it to elemental Hg plus methane (Brock and Madigan, 1988) Some of the transformations are shown below:

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176 bacterial bacterial CH3-Hg-CH3 < CHjHg* > Hg + CH4 chemical bacterial *Hg-Hg* > Hg^* + Hg bacterial Hg2+ > Hg Therefore, all of the above transformations of Hg may result from the action of some of the organisms involved in methane fermentation, particularly, methanogens, sulfur-reducing bacteria, and hydrogen producing bacteria. From this microbial information one can develop an understanding of the Hg data from the anaerobic reactors that was presented in Figures 4-31 to 4-35. It was clear that in all of the cases, i.e., control reactors and the others spiked with 100, 1000, and 2000 ng Hg that the majority of the Hg was held tightly in the SW. With the 100 ng Hg spike, the range of Hg in the gas phase ranged from 30% to 48% of the total. It was previously stated that dimethylmercury can be formed microbially under alkaline conditions and with lower levels of Hg. Therefore the gaseous phase Hg could have been a mixture of dimethylmercury along with elemental Hg. Evidence for this was the fact that approximately 45% of the Hg that was

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177 volatilized in this treatment was found in the tygon tubing, which was shown not to react with either Hg^* or Hg. There were generally only trace levels of Hg in the leachate indicating that the Hg that had not volatilized was tied up in the SW itself. In the case of the anaerobic reactor with the 1000 ng spike, Hg in the gas phase accounted for 8.3% to 12.6% of the total added. This was lower than that with the 100 ng spike and this could possibly be attributed to some inhibition of the microbes that were responsible for converting the Hg to a volatile form, although methane, VFA, and total gas production gave no evidence that Hg inhibited the other anaerobic digestion processes. The higher concentration of Hg in the SW on which the microbes fed may have contributed to this reduction of Hg in the gas phase. The low Hg concentration in the leachate and the fact that 37% of the Hg volatilized was found in the tygon tubing also implicate dimethylmercury as one form of Hg that was volatilized since dimethylmercury is formed at low levels of Hg. Another possibility could be that at least some of the Hg found in the gaseous phase was elemental Hg which had been converted directly by certain microbes from methylmercury and divalent Hg

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178 In the case of the anaerobic reactor with the 2 000 ng Hg spike, the Hg in the gas phase ranged from 2.1% to 6.6%, but most of the reactors were in the 2.1% to 2.5%, with only one indicating 6.6%. Figure 5-3 shows the different Hg spiking levels in ng and the Hg volatilized (ng) for both runs. This figure is not based on the percent volatilized, but on the actual mass of Hg (ng) that was volatilized. The data appear to be quite variable, especially for the 1000 and 2000 ng Hg treatments. For the 100 ng Hg spike, the Hg volatilized ranged from 3 to 48 ng Hg. For the 1000 ng Hg spike, the Hg volatilized ranged from about 85 to 120 ng Hg, whereas in the 2000 ng Hg spike, the Hg volatilized ranged from 42 to 136 ng Hg. This figure shows the variability in the amounts of Hg volatilized in both runs. It is possible that this same variability will be observed when sampling the gas phase of landfills to determine Hg volatilization. Key microbes in the conversion of divalent Hg to elemental Hg and dimethylmercury or methylmercury to elemental Hg and methane should exist in these reactors and may have been inhibited or out-competed by other microbes in the system as Hg concentrations increased. Escherichia coli is one

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173 o o o o o o o ^3" CNI O 00 CD "^ CM (6u) p9Z!i!iB|0A 6h en 0) N •H i-i -H ra rH {> tn a U-) JJ si a CQ ^1 u nJ (U H H H i+-t T) C T3 (C H i -H W •H s 0) •H 00 I in u en. H:

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180 microbe that can absorb divalent Hg into its system where it combines with the cytoplasm and is converted into a different compound or the volatile elemental form. Other microbes adsorb divalent Hg on the outside of their cell walls where it is converted directly into a volatile form (D'itri and D'itri, 1977) It has also been stated that the enzymes involved in Hg volatilization are chromosomally encoded in specific bacterial genera, namely Staphylococcus aureus and Bacillus spp. (Nakamura et al. 1990) The chemistry involved in the Hg binding and precipitation taking place in these anaerobic reactors is complex. In anaerobic environments there is a tendency for Hg to be precipitated as the sulfide (HgS) The insolubility of HgS makes it resistant to methylation, but under aerobic conditions, HgS may be oxidized to the sulfate form which can then undergo methylation. Under low Eh and high pH conditions, HgS can be transformed to HgSj^*, or to the free metal (Barkay, 1992; Wilken, 1992; Schuster, 1991). The binding of Hg in SW may occur in a similar fashion as it would in soil. Some of these binding mechanisms would include: (1) specific adsorption strong binding due to covalent or

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181 coordinated forces; (2) chelation bound to organic substances; and (3) precipitation as sulfide, carbonate, hydroxide, phosphate, etc. The two most probable mechanisms would probably be chelation due to the high organic matter in SW and precipitation, especially in a reduced environment with sulfide present. An interesting finding from this study is that there is evidence that landfills could be responsible for Hg emissions into the atmosphere. This has not been shown before. Other sources such as municipal waste incinerators and medical waste incinerators are thought to be more important sources of Hg going into the atmosphere as result of waste disposal. Using the data obtained from this study, an estimate of the amount of Hg emitted from landfills in the state of Florida was made. This calculation was made using the following assumptions: 1) the amount of MSW going to landfills per person per day is 8 pounds, 2) the population of Florida is 15 million, 3) the average Hg concentration in MSW going into the landfills is 200 ng/g, and 4) the percentage of the Hg that volatilizes from MSW is 3 0%. Using these assumptions, the annual amount of Hg volatilizing from landfills in the state of Florida is about 2600 pounds. To put this number into perspective, Chu and Porcella (1995) reported that annual Hg

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182 emissions from total electric utilities in the United States was 40 tons. If these estimates are correct, then Hg emissions from landfills to the atmosphere could account for a considerable fraction of anthropogenic Hg sources. This finding may be particularly important with regard to South Florida where the problem of atmospheric Hg pollution has been well documented (Rood et al. 1995)

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CHAPTER 6 SUMMARY AND CONCLUSIONS This study has provided important information about the fate of mercury in municipal landfills. Phase I of the landfill experiments indicated that mercury does exist in the Alachua County landfill, and by implications, other landfills as well. The investigations in Phase II followed the Phase I work and showed that landfills could be a potential source of mercury to the atmosphere. The conclusions drawn from the two phases of this study are summarized below: Phase I 1. Over half of the samples analyzed for Hg had concentrations below 150 ng/g. However, the distribution of Hg was highly skewed with some samples containing thousands of ng Hg/g on a dry weight basis. The highest Hg concentration exceeded 16,000 ng/g. The distribution of Hg in the Alachua County Landfill was extremely heterogeneous. These "hot spots" probably resulted from 183

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184 samples being taken directly from areas where Hgcontaining devices, such as batteries and switches, had decomposed. 2. Mercury was found in all samples taken from the Alachua County Landfill. While Hg concentrations were all well above background levels for surface soils in Florida, they were two to three orders of magnitude lower than either the current soil clean-up goals used by FDEP or maximum allowable concentrations in sewage sludge as mandated in 40 CFR Part 503. Compost samples from Palm Beach County consisting of a 1:1 mixture of yard wastes and sewage sludge were higher in Hg than the landfill samples but still far below the regulatory levels used by FDEP and EPA. Therefore, Hg should not be a problem is landfills are reclaimed and the SW residue applied to land. 3. Leachate recycling at the Alachua County Landfill resulted in a more uniform distribution of Hg with depth compared to three other locations within the landfill in which leachate was not recycled. Phase II 1. The bulk of the Hg added to the anaerobic digesters was

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185 found in the solid waste at the end of the experiments. In 16 out of 18 leachate samples, no Hg was detected. Very low levels of Hg in leachate is consistent with landfill leachate data reported in the literature. 2. Mercury volatilized during anaerobic digestion of artificial MSW. The percentage volatilized ranged from over 3 0% at the lowest Hg concentration to about 3% at the highest Hg concentration used. The presence of Hg in the tygon tubing connecting the reactor to the carbon traps was taken as evidence that forms of Hg other than Hg and Hg^* were volatilized. A review of the literature provided strong evidence to support the hypothesis that DMHg can be volatilized under conditions that exist during anaerobic digestion of MSW. DMHg and possibly other volatile organic forms of Hg are believed responsible for the Hg associated with the tygon tubing. 3. Sulfurimpregnated carbon was used to successfully trap Hg that was volatilized during the anaerobic digestion experiments. EPA Method 7471 was unable to recover any Hg from these carbon traps. However, a modification of the method was successful in recovering >98% of the trapped

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186 Hg, allowing quantitative recovery for the mass-balance calculations 4. The distribution of Hg in the solid waste residue was extremely heterogeneous. An accurate determination of the Hg associated with the solid fraction required that the entire sample be analyzed. This condition may have resulted from the manner in which the Hg was added to the artificial MSW at the start of the experiment and to the fact that the solid waste at the end of the experiment was not dried and homogenized prior to analysis in order to avoid potential Hg losses during the drying step. 5. The amounts of Hg volatilized were highly variable, particularly at the two higher Hg concentrations. At the 2000 ng Hg treatment, the amounts volatilized ranged from 42 to 136 ng. 6. Run 1 lasted 37 days; Run 2 lasted 16 days. However, the volatilization of Hg was essentially the same for both runs, indicating that volatilization may occur only during the early stages of digestion. From these conclusions, certain recommendations can be made. These are as follows:

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187 (1) The recycling of mercury containing devices and lamps must increase if the amount of mercury that is making its way into landfills is to decrease further. (2) Landfill gases should be monitored for mercury and, if Hg is being volatilized, special traps should be placed on landfills to capture mercury that may be escaping into the atmosphere (3) The effect of leachate recycling on the distribution of Hg in landfills should be evaluated further. If the results observed in Phase I of this study hold true for other landfills, then this effect should be factored in when determining the value of leachate recycling at new installations (4) Further studies should be performed on landfills, especially in the South Florida region, to determine the contribution of mercury from landfills to the atmosphere. This would allow for a better understanding of the mercury problem that exists in the Florida Everglades and surrounding systems (5) Additional research should be designed to determine at what stages during anaerobic digestion of MSW Hg is being volatilized. It should also determine whether there are

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188 Stages during anaerobic digestion when Hg can be found in the leachate. Hg concentrations from the leachate recycling area of the Alachua County Landfill suggest that there is a stage during anaerobic digestion when Hg is mobile and redistribution with depth in a landfill can occur,

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APPENDIX A RAW MERCURY DATA FOR ALACHUA COUNTY LANDFILL RESIDUES AND PALM BEACH COUNTY COMPOST

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2 "^ i "" 1 n 5 in 00 o CN 00 ffi tji H m H o m CO H CT\ o H H H 00 H H CN H CN ro CO 00 H CO cn H in CO in H ro U) ro H o in JJ CD (N o H CN m rH o H 00 H O H n o\ o o H CN in o o H o CN O O H H CN O CN H H O H in CN o H m m o H 00 o ro o H H CN O O rn I-I o H H O O H cn CN CN o H cn o o H in CN o H CN CN O H in rn H o H ra % -H m o -* in H Ol o IN IN 00 o Ol m m •* CN o in H m H CO o U3 cn H o H H H H 00 O CN in H rH rro ro CO CO H o •aC7> in H CO 00 H ro U) fO H H in rH rH -H C ID ja U to 6 c m in 00 H O O CO m o o Ol o o 00 o o m o H o CO CM H o o in H o in H o U3 oo H o to o Cvl o n o o o O O o in H O CO cn o o o o o o CTi O CN O o o H CO O O o in o o 00 o o CN VD H o o^ cr^ O o t^ H O o CO m o o o o CN O H O in o o in H C3 CN (N O O in o o H o o JJ 8 re 3 U (0 H < w CO JJ (0 Q >. P O u (U 2 S (0 a: Q H <-l A! c to m C o H JJ to o o o in H n T) JJ CO c o o CM JJ CO oi c o in CN l/l 0 JJ CO C o o T3 jJ CO w c o m cn r~ JJ to c o o CO JJ CO o to C o o un o H T3 U 0) U m CO fO m in (N CN o m o H CO o C E u in ro in CN fSI o O H CO in in CN CN o cn o H CO o & in CN o m o r-t CO D Q o ^ cn & in in CN o n o H CO fa O § e J-l in in m CN o ro o H CO fa u in H 1 in CN o o H CO fa u in CM m H o o H CO fa •J o r 01 m U in CN in H ro o ro o H CO fa J CO o Dl CO (d a in CN in H ro o ro o H CO fa o d E 0) u u m (N in H ro O ro O H CO fa .J o u in in (N ro O n o H CO [K k5 PQ U U 9i w 01 c o in B o CQ CO tC Ol in H in H o •* o H CO fa O # e (U i-l in H in H o O H CO fa o to CO fd Ck in ro in
PAGE 205

192 "?, C B S u u U c#> O n m H (N H t^ H H m in Dl 0) 01 £ m OD rin 4J r-ro o t^ o H H cs o o o o 01 9i H H H H a a d B o a H m to 3* H •* c 3 s G O O H CN O o XJ C O O O O O o ^ LP a ""^ J2 u ns H ^ ^ 4*: : til m k ^ V o "^ ^^ .(Jt (IS O H H C u 3 u Q CO B e [13 H 01 QJ (U a V^ K -^ k r^ l> !) o o o O o in s (N r^ CN CN — CO fe W CO CQ B fc W b b U a a J _h£ U u H < OJ '^ 11) 4J (0 p H CN cC H H

PAGE 206

193 o 4J >. m G U u iH CO lie TJ -d TJ -d — 4) (U (U ID (N ^ ID tn CO oi 00 H H o H -^f CN f^ cn ro H CN CO m m in 01 ^ 2 tJi £ ^ CN m "co" ~ CN > u) "o" U3 in r^ O ^ H "h" CN 00 CO [VD 4J [^ CN o in H (N m n CO m a\ H 'I' n 00 in H c^ CJ^ H r^ s o H m ro ro m m m H H OJ o n m (N m ro (N T-t CM CO o O o o o o o o o O o o o o O o O o o O o H H H H 1-t H H H H H iH H H H H H H H fH H H a E (8 W (N m a\ a\ 00 [CS UJ U) H n m m CN CJl CD CO "i^ 77 (N CO 'J' CTt m ~ CO ID ^ m O m in m m m ^ KD CO ^ VD CTl LO 0) H O) m '^ in rH a\ H H H CN Vi) (N CJ m cn H CN CO 3 — % m ro in 13 Cn CTl •H K e: to ^^ 1) 0! ~ -H r^ m m o m o o H a\ m rji in o o rrcn H > H (N ^ m rf 00 H cn i/i in 'J' r~ CO '^ ^ VD ja o o O H H CN CN o ^ o o m ^ ^ i-l H H O o H O o o o o o o o O H H 13 VX) o o O O o O O O o o o H iH (N O O o O o o O o o o o o o o o O O (S m m J ^ in N ^^ 4J "~™ ^~ ^~ c ^~" ^^ ^^ ^^ 3 U o o ft B 9 0 X H ^ -* -H ^ ^t o o a _E O .^^ Ixl TJ o o iH •^ I-) 04 W ca tiJ rH ^ =tt at o H o 13 M o CU = = at c ^ c c 13 ro c — o CO ^ TT C ro ^1 •* — ^•^v. -^v u 01 o o # H H 01 to to tJl 4 cu u rn u U U H ro U < < <; < ro }^ CO 0) o o o o o o o o o o o o o o ~ o O m ^ ^ ^ ^ r>) CN CN m ro in n ro ro in m m m ro CO 1 1 1 1 1 1 1 1 1 1 1 1 H t H H H H 1 1 o o o o o o o O o o o 1 m in in in in <>. tn CJi cn tn Cn (N m m m cn m H H H (N
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194 01 p. c i 1 u HI u •a ^ • — H ^ CN 00 00 o H c^ rH ro H m VD c^ o CO m ro ^ in n Ul CO U> in rT< O H t^ lO 'J' CO ^ -a* oi H [N 01 H H in CN CN 04 ro H CO 'cr CN CD L/l M CN O] H CN r-i ^ H 01 -^ s tn _c y3 <^ O H U) CO r> H ro r^ Ln 00 U1 Ol H in ~ ^ ro 00 u> o •y* CN i.) H o U) CN ro ^ in H LO CN m H \i> in CN ro ^ O) ro ID in CN c^ CN CO ^ o CO 01 ro ^ Ol m m CTi ro CN 00 VD ro I~ CO Ul cri in H <*• CN r^ t^ in O 'a* in ''J* 0) H (N Ol i-i in CN CN CN ro H m >!< CN Ol IN CN CN rH CN CN in H 3 H LO V Cn 01 •H a c 01 lU ~ B! H o PO \£t VD Ol O ^ -sf 00 77 O ro ^ t^ t~H in in Ol r^ LD a^ > tt^ CN d H CN o H ro rH Ol ro CN ro ro CO H ro V4 O o O O H CO o CN H iH H H O H CN H Ul H H o o H O H H CN O O O o T3 C l^ O o O O O O o O o O o O O O O O CN O O o o O O o O O O o O o a m m J 3 in >i ~-4J ^^ ^^^ ^^ c 3 (H rH C c c d C C C 9* B' c 01 c 1) C 0) (U cu cu C 0) u (d § 0) CD 0) u ID 0) HI ^4 0) CU IH 0) 01 03 0) 1 IS TJ tn 01 01 91 >H U U V4 u CJ V U u -— TJ C c C U 01 U m CJ 01 m m 01 cn cu G 3 t3 •a •H •rt 0) 0] 01 0] n w 3 C (^ 4-4 14-4 rH s c = •a ^^ 1-1 3 3 ft z 'd* = ^ s ^ ^ ^ Tj< = V t-i u cn X—. T) T) E ^ ^--, -S" ^ 'fl* ^^ o — tn D4 ^ ^ c CO H H H rH H H "^ H ^ a — 0) O Ol 01 3 01 H r^ H rH M c # rH < QJ C! Q c C C C c C c — d -H "— iH rH V4 M H C C c c ^ Dl o -H 4-1 iw o o cd 01 01 CO c fO D e m O -i _0 E ] in rH s in H r-4 H CN o CN X rn ro tj) rH i-l H H H H H H H rH CN rH H CN CN 01 ro ro ro H H CN ro cn .— o o C O O O O O O o o O O O O o o O C o o o C3 O o O c 13 >jj VD CO CO CO CO CD 00 00 CO 00 CO CO cn a\ C\ m Ol CJl Ol Ol Ol m Ol o o o O o o O O o o o O o O o o o o o o o a O o o o o i4 m cn in CN CN "^ H I-) CN CN CN CN CN CN
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^ in (A i-S r**! c > 00 o "o E s E o g. O ^_ CO s CO CD CO cn CO "^ o CO in m o CO o O CD CO CO o CO o CM K CO CO CO 5 in t^ 5) CD CO •* ^5 CO o ^ O) ^— S I o> c w c c c c c c c c ,g c c c c c c o g g o o g o c >. o o g o g g j= = a> 3 ^ 5 5 3 _3 3 ^ E O "3 _3 ^ 3 _D _3 5 B ^ ^ u ^ 6 ^ p ss o o ^ •u •u ^ B o o o o o o o o O o o M o^ CD o in o o O o "n "O "U •o CD CD m CD to S CD 3 in CD in m in OO ID ^ ,$ o O o O o o O CM ^^— ^— ^— ^— ^— EC t£. V^ q: CM CM CM CM CM CM CM in (D o 00 CO 00 o rcn CO ^ CM to O) CO to CD .,_ m u O) CD ID CM S CM CO CO o r-^ 00 CM CM CO CM CM in o o c o! ^ '" X 3 u s: CO ^* CM CM ! ^ Q u 3 j5 ig llT a ^ c ^ o lU 0) _I tr IH ST ." Id Q o 5 o Q o a: b o o co o 3 a. E o 3 s a CL CO S c cn (0 V) ff o u o 01 CO CO o o 1 in Q. a. Q. in CO a. E g g g E 2 >11 d d d d d d ST 3 o o CO t t CM CM CM CO ci o o ci o cb in X _„^ "3 "3 "3 ) C3) CM CO CO CM X Ol c c c c c o c CM s CD 13 o o o o o 2 o o o o o o o ST c o m in in in o 55 o TjTf Tf CO 00 00 o> 1i c CQ in CM CO CO 3, in (D q in o in o iri q q q q m 5 3 Q Q. CO o in > -a u u D 73 o 0) W CO CO CO CO (0 CO _i < .^ CM CO CO in to r-rCO CD o CQ o u LL Ll LL LL U. LL IL o o o o o o o o o o o CD 55 CO S5 55 W w O _J _] _I -J -1 -1 —1 Q. CO CO CO CO CO CO CO CO to CO CO CO O O CD H 05 ; B CN 0) rH CD ^— Q i3 re H 1195

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210 1 — — — O en en CO OO o> .^ CM en m m en •Kiin ho m to t CM eo CM ev CO m CO CM m •t— o CM CM "3 O) I S" _c B c >> a) "o E O o o T3 a. o CM O D o in TO D in m CM CO CO CO CO CO m CM to to CO CO CO CO CO O 1 O O o (O CO o> fm O CO CM CM Nco en t^ t^ m en en tT en en r(O eo m m s to O) CM ^ >^ ^ ^ 11 W ^_ t^ C3) CJl 00 m CO CO 00 oo ~ CM ,_ to en r00 in o o m O CO o CM T. 3 (O m ••aen en rto (O in m m CO en CM rr CM 00 o> in in CO CO CM T— T3 CO CM eo CM CM in CM CM •^ 'CM ^ ^ CM •* Tt I-t ) ra Hi I _c_ nH H 4-1 T( C to O) yf in in •^ in en o "eo" } m r^ T— r^ rto h00 in o £ rt CJ) in CO in > < m 4-1 Ci u 3 18 H a: ^,^ ^ a. 1 D >u 1 Q ra c c c c c c c o o o o o o o o o o o o IB Q d i (U 0] 0) c i a; o tn o c/) (0 c o o o c o o o c o o o c o o o 3 o V) : S s (/) = = = = o o o o "5. £ k g; 5 5; g; 5 5; ^ 5; E E E (U E E i E E u. ^ u. u. S i_ i_ i_ £ E 0) £ E E E E E E E b b b b b b b b b b b b s S £ S S 0) 0} £ b !q J3 'n b XI -Q .a b .Q XI 5 b B 3 B a ^— V ^-, — ^.^ >_v ^_v 2 XI B .a --V ib in in --V in in in In ^-^ in io in ^-> u? in 57 ^-v ST in in --^ b b b in 1 CN 1X5 CO in in Ed 1 CN in CO in CO in m "oi in CM in CO in in in 1 CM in CO in m in 1 CM in CO in in D) o 1 CM ei CO CO ei lO ^CN (O in CN_ eg. CO X in Ci CO in CN to in CM CO X o CM CM ^^ T— ^rT— Y en en en en o> O) en o m en en (J) en en en en en en en en en g ej) en en en ij 1 o o o o o o o o o o o o o o o o o o o o o o o o o If) in in in in in iri iri iri T-^ c-^ r^ T-^ T-^ T-^ T-^ T^ T-^ T-^ T-^ T^ CM Cvi CM CM c u H 1 CO CO CO CO CO CO CO CO (0 CD O ? CO CO CO CO CO CO CO CO CO CO CO CO CD O > O CO CO CO CO 1 u. LL li. u. u. LL u. LL LL u. LL LL LL LL LL LL LL LL LL LL LL IL LL LL LL 1 _l _l _i _i _j _I _l _1 _I O O _l _I _l _l _l _l _l _l _l _l _i _l O o _l _l _l _l 1
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APPENDIX B RAW DATA FOR ANAEROBIC REACTORS

PAGE 225

(T> CO ^t CM rro t--. CM CO on m CD in ^ m CO 00 o 00 o CD CO 00 00 O) o> t^ t^ t-CD hCD r^ 1^ t^ hhfrhCO CI) (N tN t^ flCO m CO CO r~CO ro CD CD CD CM ^ in CO CO 00 o ro m 00 hr-CO CD CD CD CM r^ t^ r~(D rCD hr-hhr-^ hr^ r-rCO o> CO •* rO •< in on in CM CO ro en in r-^ CJ) CD CO CD t^ CO 00 CD CD CD CO O) S O) hCO in Ol co to (M co rCO Tf in CT> CM CD ^ in in CO O en co O in CO ho o o O T— hhrh O) CO c m 00 TT O (-in m en O m o> Ql t^ NK K CO CD r-. 1^ rt-00 hhfr^ i~1m CD ro
PAGE 226

213 s CO 5 CO in CM in CO O) CO CO 00 CO o 00 § 5 oo CM CM CJ) CM CO CO o ai CD CO in o 00 cci a. r^ CN CM (T> CO CD en CM CM CO CO m o CO oo CO CM 00 CO o 00 (O in CO CM CO CD CO co CO CM in CC} o oc> £ in CO CM r^ in CJ) si CD 00 in CO 00 m o 00 S CO in CD £ CO CO CM O) 00 in o oo q: 5 2 CO s CM CO CO CO f2 CO o o6 t--i CJ> I--' co CO o CO CM i CM CO CO § £ CD o in CO CO CO CO 1^ CO in en Ni S ra Q 1 X u. £ CO S r^ CO CD CO CO 00 CO CO in 1^; CO CO CM 1-^ CO oo O 00 csi m s n CO 15" Q o CM CO ^ ID CO t~00 O CM

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214 a: DC o o d o d csi CD g en CM s c\i m (D in o o in CO (6 in CO (D m CO in in m CM oo CM in O CO CM m 3 g CM iri m co CO g CD CM CM d CO CO g s cri m CD O) in en CO in CO a: o o d o o d CM 00 CO o csi CO en CM CD CO o o o to m m CD 5 CO CO in CO iri in 00 o iri in CO OCJ m en m CM r-. CO m o d oo q in to o C^i o m iri in en CM iri m m OC) CM 00 in CO to in § d o o d q 00 CM a> CO C3> CO CO CO m d t 00 m m CO CM csi m oo CO 00 oo CO m CO iri m CD o •4 f5 CO g CO tn CO CO o CN oi CM g CO q In (0 a. o o d o o d CD CO CO c\i CM m CO in < CO CD in CM d "if CO T— CO >! CJJ in 00 CM o m to CO -"J Si to to CO CD to m en o g CO CM CO m q CM iri d m S m o o d o o d CO CD d CO CO CN in CM m CO g m in CO d m CD (b m en in s a CM CM iri in o CO g 00 CM is CD CM m 00 m iri m 05 m CO CD CD in in CO in g s o o d o o d 00 d CO CO at d CO CO 5s CM m en CO CM m CO m o q o CO m CO CM CO in m CD d CO m CO "Ien d in 00 o d 00 '< iri m o iri o oo CM in g CO in 5) 5 s o o d o o d o o d CO CO CD CM Oi CO CM CM CM CO CM 9S5 "JCD CO cci CM T— CM m CO CO in csi m o in to in en oo CM in CO 5 CO r^ CO in CO g CN CO •) iri in oo CM cei m g CM a. g d (D CM d CM in d CD CM
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215 H rV j) rrCN CN V t00 in H 'J* 01 o in o xji 'J' CO m ** O! ^ ro H ^i) o m m o U3 U3 GO U3 00 o CD rU) (J\ t~ yj in ^ m VD a 00 rin E^ 0! t^ o (N n c\ •* [^ H 5 rtn rin CN •* ^ -^ -* -5" ^ •^ ^ '^ r[-• pn \o m H O c^ CN r^ m m m PS a* U! ^ O o m C^ in m m m VD m Oi >* (N o •*< (N IT> ro o o t£i in in \r\ fM ^ in Ln n in in in in in in m r^ m m "^ rN (T CN rC7\ o o o cs r^ H [^ CN CN OJ CN rcs in o U) ^ rt m in m ^0 KT) H [^ KO "S* ^ in in ^ (N > ^ ^ ^ in ^ in in tn in ITi O) C 3 l/l a> ^0 IN CO CN O ys) o ro o tn Bi (N 1^ m CTl \X) H 00 ^ C\ rCN o o H U m H H r^ rm H in CN H t£) ^ in u m IN ^ in in ^ in in in in m 4-1 D H a\ o m > [^ c^ U) CD o o tn t^ H [^ o 00 H H UD Ch tj< m r-l CN u) # B! H in m [^ m CO o H ^
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216 i o \i> rH r-UD m m t CO j H CN > Ol rt CN vo CN rt 1 C31 H in ro VO CN 1 r^ o 'I* CO I H m O 1 H 1 rvo H r^ vo CO 1* in 1* in in VD H n o (N 1 o CO in O CN o CO 1 m M l/l m CN ^ 1 H CM 1 in o u> vo m •a* H 1* > CN VO CN tO o CO o o in t^ 1 O Ol i CO CO o vo o in 11 t ro m CN in ro ro o\ vo o M o o o o H o 1 o o o o o o o O o o o o O o O 1 • in o o o O : O o o o o o O o o o o O o o a tn o 1 o o &i Tf o o o o 1 o o o o o o o 1 o o o o O o o o o 't CO m t> m m : ^ ^ 10 CO 1* 00 f^l 11" vo > vo in (T, (71 i vo -^ o in i vo en GO in fS r^ VD 1* 1* r^ VD CO CN o en ro VO rH fO ro CN V H H VD If) Ol VD o H H m in CN 01 m [^ o r4 H CM H H H CO H O H CO o CN ro VD r* 1 c\ 00 00 VO O in 1ro 1 (^ m rfO ro CO VD -"l* o o o O o o o o o O o o O O 1 O o o o O O o o o o O o o o o o o o o O 1 O o o o O o O 1 o CO o • 1 • O • a o o O o 1 o o o o o o o 1 O o o o O o C3 j o o o fO en CN cn m ] CO m r-i (N VO (7\ CO OI 1* CN 1< in vo CN ro voj fO u^ t^ CO [^ [^ H H ^ l/l o\ (N CO C\ H (*1 IN in H D H H in o in rO n 1 m in m VO o VD a\ en in 1/1 VD (N CT\ m CT\ o CO O ro vo o CN o m H 1 i-l CN in VD H CT\ o^ 11 in H O o f^ in Ol vo o o vo H o o m CO r(N H ^ o VD CTi m H 1* in CO rin H o H m o CO CO r^ vo VD VD 1/1 in n f^] CN m VD ro CN r~ vo tn o o o O i o o o o O o o o o o o O o o o o 1 o o o o O ; O o o o o o o o o o o o o o o O 1 o vo o • o o O I o o o o o o o o o o o o o o o o o o o \D H H rCN tn rLn r~ tf CO in VD 1< CN H in CO H vo ; T p* ^ in -^J* ^jt CO CN r^ a\ i/i m o^ O H VD o vo o > H : VD CN c^ o n CO VD in o o\ (N T 11 m 1> in O in o 11 ro en ro H U) CO 00 -* CO rH CD o\ CO Ol 00 rC3 en 1" ro Ol H t^ O o m in ^ C3 CO 00 VO VD 1/1 11 l< m CN 1* VD ro o O 00 rp O o o o o o o o o O C3 o o o o O o o o O o o O o o o o o o a o o o o o o O o o o o l/l • o o O oi o o o o o o o o o o o o o o o o o o o o o rrH CN ^ n CO fO m m vo >r m 1" VD 00 tn vo in CO en 1 00 H (T H -* CN CN 1 rH [^ Ol o o H 1" CO in in ro rin tn vo CN i n n in in in ; \D n en in (S o [^ 1* CN CN CTl CN 00 ro 00 vo in H o oa in > j vo (N r to H vo m CO H Ol VD in ^-^ o o ^ CN ro CN 1 m o in O 11 o •* o in m CO o in 011 m CO 13 1 CTl V£) CO CN 1 m M vo m m vo H 1" [^ H o o fO co r^ O i l> a u H 'a* o o o ^ •^ o o o O 1 o o o o C3 o o o o o o O o o H o o m m U> CO o o o o j o o o o o o o o o o o O o o O o o > CO m o o o o i o o o o o o o o o o o o o O o o in tn rH r00 1 ro in vo rvo vo 01 vo 1* rOl H m in VO in ; r^i 01 o o en m rH CN 1 in CD CO H r^ VD in H CO 1* CN in H CN in vo 1 ro S>j 1 o m rH 1 vo (N ^ M 1* H CO O o [^ in tn in o m H i CN ^ W W H o H t^ VD -* m H r^ O in 11 1> 1* in t^ Ol r00 o 1 o n (N rCN 'T in CO i 1* CO vo fO in en en 11 1-t^ CN 1 00 H m H CO IN r-i (N CO 1 in ^ tn OJ H VD r^ VD ta\ 1* CD 11 CO en CO 1 m S" : o un ^D CJ^ VD ^ in in CM o m CM m o H vo vo Ol CO ro 1 00 ro CN 1 r1 H lii O m C^ CN IN i vo o\ en VD rvo o ra\ H o o O m 1^ T^ o 01 1 o o H m "t ro
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217 >^ Ot H t^ un m in 1 m 10 O" H CO rO 'J' m 10 m Ol 00 H > m in Ol > M ru> M o H m VD O CO tm r^ VD •* (^ n r-' o 1 in 1 O rn CN CN H CN O \£} in o 1 O o O • 1 ce: 1 O o o i O O 1 o o I o o in H H un ^ Iin ; in r^ 00 1 CO VD r^ o '<:l' 00 m m ^ rCM tH H 1 r-( Ol in Ln m ol m rin CN f*l m o CN o VD 00 •* VD m o ^ lO o H ^ ^ o ^ a\ m cn c^ o 00 O m CN VD CN Ol CO o ra H (N (N H CN 1 H CN !-i H o o o o O O O o C3 1 o O O O o o 00 o o o o O o o o O o O o o H m r(N 00 in ^ 1 m in CN rVD VD u> m 00 H H CO in 1 'S* m I r~ CO en m n m n M H CJl ID 1 r* en 00 j Ol VD CN O m H H in in in m m CO m en m [^ m ro CN cin CN t ^ in o 00 o H H H (N H H H CN CN H H o o o O O O o o o O ; O o C3 o o o o O O o 1 o i o o 13 o O o H 00 in fl00 r^ 00 in ^ CO 00 ^ ^ r^ o (N 00 c^ o CJl o C\ m rIl/l VD H 1~ 1/1 f^ in 1 10 i H in in H OI H fN OJ \D CO o 10 ID 1 in in ^ CN CO a\ ro H t> ID ii> i VD ID C71 00 H O o o M M (N H H H m CN H H H o o o O O O O o o O O o O Pi o o o O O O O i o o O o o o CN m m Ol H -* in : 1 1 H CN H H H o o o o O O o a 1 o o o o o o c Pi o o o O O o 1 o o o o o o o VD VD (N m rH VXl ^ CN ^ COl I^ CO s m m Ol t-(Tl ^ rn CN in CO ^ m a\ o> m 10 rf H m H -^ Ol H ID u ro O) H VD H ra CO in ID m m Ol in H ^ U) 1/1 UJ o 00 C^ 00 ^ o CTi o o CN o H H tH H H H H H O IH •^ o o O o O O 1 o O o a o o O Pi o o o O O C3 o o o o o O i <^ rU) VD 00 r^ CN ; ~H~ m H ID en en T) i m H -^J* o in H \£> H tn H f^ H H ^ CO in CN o £? O O H O (N CN 1 O H CN CN o H H > n o O O o o o 1 o O o o O o 1 Oi o o o o O o, O o c^ ro OJ m U1 in fli rin 1 m in f r-'^.n m CD m H lO 00 00 CO in m vo o £ o 00 <* 1/1 1 ^ ID H CN VD H Ol — m cs ro H t~ CO m o ; en to VD o Ul H in I~ ^ H 1 c*l OJ ID i r~ 00 IN H 0 i o H H H CN O H H CM CN H rH O H o o o O O • O O O o o O 0) M O O H >< m o o o O O o O o o o O 1 (U o rH CX} r-t H CO (N VD H m ^ o c 1 C-m U3 l/l CTl VD i in CN H m o in 1' Id o 00 rH CO rn CN m O 'Jt tin 01 X! 00 a^ ro tH CO f^ ^ 1 in C^ o Ol 4J o o H H i-t H : H H 1 CN CN H H 1 o (U I o o o O O o j o O o O o O o s H i • \c _^ o o o O O o o o o a o O o H M ro 'J* in lO r^ CD o a ^ VD 00 H H rH H H CQ i 10 (0 i H Q 1

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APPENDIX C RAW MERCURY DATA FOR ANAEROBIC REACTORS

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Table C-1 Run 1 Reactor 1 Solids Data Sample Absorbance I.D. (253.6 nm) i Blank 0.017 Std 1 (20 ng) 0.027 Std 2 (40 ng) 0.036 Std 3 (80 ng) 0.055 Std 4 (100 ng) 0.065 Std 5 (200 ng) 0,110 Std 6 (300 ng) 0.160 ICB 0.016 MB 0.017 SP(40 ng) 0.036 R1:1 0.017 R1;2 0.019 R1:3 0.016 R1:4 0.016 R1:5 0.017 R1;6 0.017 R1:7 0.017 R1:8 0,017 R1:9 0.017 R1:10 0.016 R1:11 0,019 R1:12 0.017 R1:13 0.017 R1:14 0,017 R1:15 0,017 R1;16 0.015 R1:17 0.015 R1:18 0.017 R1:19 0.015 R1:20 0,017 R1;21 0,016 R1:22 0.017 R1:23 0,016 R1:24 0,016 R1:25 0,017 R1:26 0,015 R1:27 0,015 R1:28 0,014 R1:29 0.017 R1:30 0.016 R1:31 0.014 R1:32 0.017 R1;33 0.016 R1:34 0.017 Hg Sample Wt [Hg] (ng) (g) (ng/g) 20 38 76 96 186 286 38 2.0397 4 2.1330 10 2.2612 2.2974 2,1105 2,1049 2.0774 2,2601 2,2988 2,1192 4 2.3171 9 2,0248 2,3400 2,0583 2,0072 2,1623 2,0680 2,2206 2,0230 2,0842 2,1972 2,2233 2,1937 2,1667 2,3480 2,1263 2,1135 2,2838 2,1548 2,1548 2,1322 2,1597 2,1025 2,0686 Constant: 0.0171 X Coef.: 0.0005 Solids (g/g): 0.1861 219

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Table C-1 Run 1 Reactor 1 Solids Data 220 Sample Absorbance Hg Sample Wt [Hg] l.D. (253.6 nm) (ng) grams (ng/g) R1:35 0.017 2.0506 R1:36 0.016 2.2465 R1;37 0,016 2.1327 R1:38 0.016 2.2500 R1;39 0.017 2.1495 R1:40 0.015 2.1588 R1:41 0.017 2.2640 R1:42 0.017 2.0596 R1:43 lost in batli R1:44 0.016 2.2229 R1:45 0.016 D 2.3092 R1:46 0.017 2.1860 Sum 8 Sum 18

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Table C-2. Run 1 Reactor 4 (Solids) 221 Sample Absorbance Hg Sample Wt [Hg] I.D. (253.6 nm) (ng) (g) (ng/g) Blank 0.020 Std 1 (50 ng) 0.036 40 Std2(100ng) 0.055 88 Std 3 (200 ng) 0.090 175 Std 4 (300 ng) 0.136 290 Std 5 (400 ng) 0.178 395 Std 6 (500 ng) 0.210 475 Std 7 (600ng) 0.259 598 Std 8 (800 ng) 0.333 783 ICB 0.021 MB 0.021 SP(50 ng) 0.038 45 R4:1 0.021 2.2520 R4:2 0.022 2.3044 R4:3 0.020 0. 2.1007 1 R4:4 0.020 2.1259 1 R4;5 0.018 2.2484 R4:6 0.021 2.1313 R4:7 0.019 2.2756 R4:8 0.021 2.3444 R4:9 0.020 2.2259 R4:10 0.015 2.3364 R4:11 0.019 2.1908 R4:12 0.023 8 2.2734 15 R4:13 0.022 2.0508 R4:14 0.018 2.3432 R4:15 0.021 2.1236 R4:16 0.020 2.1227 1 R4:17 0.020 2.2820 R4:18 0.017 2.1323 R4:19 0.022 2.2966 R4:20 0.020 2.2742 R4;21 0.021 2.4045 R4:22 0.022 2.0026 R4:23 0.020 2.3805 R4:24 0.021 2.0927 R4:25 0.020 2.2851 R4;26 0.020 2.1467 1 R4:27 0.023 8 2.3498 14 R4:28 0.022 2.1857 R4:29 0.021 2.1243 R4:30 0.021 2.0814 R4:31 0.018 2.3296 R4:32 0.020 2.3366 Constant: 0.0167 X Coef.: 0.0004 Solids (g/g): 0.2312

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Table C-2 (cont'd). Run 1 Reactor 4 (Solids) 222 Sample Absorbance Hg Sample Wt [Hg] I.D. (253.6 nm) (ng) (g) (ng/g) R4:33 0.023 8 2.1322 16 R4:34 0.019 2.2312 R4:35 0.020 2.3566 R4:36 0.021 2.1812 R4:37 0.014 2.2184 R4;38 0.017 2.3134 R4:39 0.022 2.2359 R4:40 0.023 8 2.2083 15 R4:41 0.019 2.2501 R4:42 0.017 2.1221 R4:43 0.021 2.2996 R4:44 0.022 2.3528 R4:45 0.020 2.0839 1 R4:46 0.021 2.3864 R4:47 0.028 20 2.3681 37 R4:48 0.017 2.1516 R4:49 0.021 2.4390 R4:50 0.021 1.6379 R4:51 0.020 1.5218 1 55 103

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Table C-3. Run 1 Reactor 7 (Solids) Sample Absorbance I.D. (253.6 nm) 223 Blank 0.021 Std 1 (50 ng) 0.043 Std2(100ng) 0.067 Std 3 (200 ng) 0.122 Std 4 (400 ng) 0.210 Std 5 (600 ng) 0.324 Std 6 (800 ng) 0.410 Std7(1000ng) 0.470 ICB 0.021 MB 0.021 SP(50 ng) 0.045 R7;1 0.033 R7:2 0.041 R7:3 0.028 R7:4 0.021 R7:5 0.027 R7:6 0.021 R7:7 0.021 R7:8 0.015 R7:9 0.031 R7:10 0.021 R7:11 0.023 R7:12 0.021 R7:13 0.055 R7:14 0.023 R7:15 0.225 R7:16 0.020 R7:17 0.164 R7:18 0.020 R7:19 0.020 R7:20 0.021 R7;21 0.020 R7:22 0.021 R7:23 0.026 R7;24 0.142 R7:25 0.023 R7:26 0.026 R7:27 0.033 R7:28 0.024 R7;29 0.049 R7:30 0.035 R7:31 0.070 R7:32 0.025 R7:33 0.036 Hg Sample Wt [Hg] (ng) (9) (ng/g) 44 92 202 378 606 778 898 48 24 2.1642 55 40 2.1145 93 14 2.1744 32 2.0030 1 12 2.1892 27 2.1628 1 2.1053 1 2.3565 20 2.1364 46 2.1357 1 2.5004 2.0631 1 68 2.2410 148 2.4749 408 2.3168 854 2.0110 286 2.0581 674 2.0917 2.1566 2.3208 1 2.1609 2.1584 1 10 2.2037 23 242 2.3713 495 2.0301 10 2.1488 23 24 2.3520 50 2.0944 56 2.2484 122 28 2.0330 68 98 2.3653 202 8 2.0863 20 30 2.2845 64 Constant; 0.0248 X Coef.; 0.0005 Solids (g/g): 0.2064

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Table C-3 (cont'd). Run 1 Reactor 7 (Solids) 224 Sample Absorbance Hg Sample Wt [Hg] I.D. (253.6 nm) (ng) (g) (ng/g) R7:34 0.034 26 2.4125 53 R7:35 0.024 2.4265 R7:36 0.023 2.3693 R7;37 0.020 2.1092 R7:38 0.020 2.2330 R7:39 0.185 328 2.2244 715 R7:40 0.028 14 2.3256 30 R7:41 0.021 2.4785 1 R7:42 0.019 2.2832 R7:43 0.020 1.9144 R7:44 0.020 1.6099 s = 1757 3 = 3802

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Table C-4. Run 1 Leachate Raw Data 225 Sample Absorbance Hg Sample Wt [Hg] Constant: 0.0332 I.D. (253.6 nm) (ng) ml (ng/g) X Coef.: Solids (g/g): 0.0004 0.0034 Blank 0.033 Std1 (10 ng) 0.037 9 Std 2 (30 ng) 0.043 25 Std 3 (50 ng) 0.054 52 Std 4 (80 ng) 0.065 80 Std 5 (100 ng) 0.073 100 Std 6 (200 ng) 0.115 205 Std 7 (300ng) 0.150 292 R1 0.033 100 R2 0.033 100 R3 0.025 95 R4 0.038 12 100 R5 0.024 100 R6 0.029 100 R7 0.023 100 R8 0.023 100 R8 (dup) 0.022 100 R9 0.023 100 s = 12

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Table C-5. Run 1 Activated Carbon Data 226 Sample Absorbance Hg Sample Wt [Hg] Constant: 0.0182 I.D. Blank (253.6 nm) 0.018 (ng) (g) (ng/g) X Coef.: 0.0005 Std 1 (5 ng) 0.021 6 Stdl (10 ng) 0.023 10 Std 2 (20 ng) 0.030 24 Std 3 (30 ng) 0.033 30 Std 4 (50 ng) 0.044 52 Std 5 (80 ng) 0.053 70 Std 6 (100 ng) 0.069 102 Std 7 (200 ng) 0.102 168 Std 8 (300ng) 0.164 292 Std 9 (500 ng) 0.257 478 ACa:1 0.016 ACa;2 0.018 ACa;3 0.017 ACa:4 0.027 18 ACa:5 0.022 8 ACa:6 0.027 18 ACa:7 0.030 24 ACa:8 0.033 30 ACa:9 0.035 34 129

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Table C-6. Run 1 Tygon Tubing Data 227 Constant: 0.0214 X Coef.; 0.0005 Sample Absoitance Hg Sample Wt [Hg] I.D. (253.6 nm) (ng) (g) (ng/g) Blank 0.022 Std1 (10 ng) 0.026 8 Std 3 (30 ng) 0.033 22 Std 4 (50 ng) 0.044 44 Std 5 (80 ng) 0.054 64 Std 6 (100 ng) 0.067 90 Std 7 (200 ng) 0.112 180 R1 0.020 R2 0.023 R3 0.022 R4 0.028 12 R5 0.036 28 R6 0.030 16 R7 0.035 26 R8 0.031 18 R9 0.026 8 109

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Table C-7. Run 2 Reactor 1 Solids Data Sample Absorbance I.D. (253.6) I 228 Blank 0.034 Std1 (10 ng) 0,038 Std 2 (30 ng) 0.049 Std 3 (50ng) 0.057 Std 4 (80 ng) 0.067 Std 5 (100 ng) 0.076 Std 6 (200 ng) 0.132 Std 7 (400 ng) 0.219 R1:1 0.021 R1:2 0.033 R1:3 0.033 R1:4 0.032 R1:5 0.290 R1:6 0.020 R1:7 0.031 R1:8 0.026 R1:9 0.016 R1:10 0.032 R1:11 0.028 R1:12 0.030 R1:13 0.029 R1:14 0.026 R1:15 0.038 R1:16 0.030 R1:17 0.020 R1:18 0.030 R1:19 0.034 R1:20 0.026 R1:21 0.018 R1:22 0.020 R1:23 0.022 R1:24 0.019 R1:25 0.034 R1:26 0.018 R1:27 0.033 R1:28 0.035 R1;29 0.027 R1:30 0.031 R1;31 0.034 R1;32 0.034 R1:33 0.030 Hg Sample Wt. [Hg] (ng) (g) (ng/g) 8 30 46 66 84 196 370 2.2425 2.2619 2.0481 2.0592 2.0522 2.2757 2.1452 2.0302 2.1952 2.1445 2.0255 2.2599 2.0762 2.1444 8 2.0727 19 2.0568 2.0714 2.2542 2.0416 2.1244 2.0180 2.0215 2.0591 2.1308 2.0502 2.1781 2.2825 2.1243 2.1981 2.3415 2.2074 2.1787 2.2705 Constant: X Coef.: Solids (gig): 0.0332 0.0005 0.1911

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Table C-7 (cont'd). Run 2 Reactor 1 Solids Data Sample Absorbance Hg I.D. (253.6) (ng) 229 R1:34 R1:35 R1:36 R1;37 R1:38 R1:39 R1;40 R1:41 R1:42 R1:43 R1:44 R1:45 R1:46 R1:47 R1;48 R1:49 R1:50 R1:51 R1;52 R1:53 R1:54 R1;55 R1:56 R1:57 R1:58 R1:59 R1:60 R1;61 R1:62 0.023 0.020 0.024 0.023 0.025 0.032 0.037 0.019 0.035 0.029 0.036 0.024 0.024 0.033 0.022 0.020 0.025 0.030 0.034 0.032 0.022 0.027 0.023 0.026 0.034 0.026 0.032 0.032 0.033 Q ample Wt. [Hg] (g) (ng/g) 2.0555 2.1996 2.1790 2.0875 2.0690 2.1030 2.1175 2.2493 2.0530 2.0693 2.0998 2.1430 2.1707 2.1834 2.2002 2.2263 2.2540 2.0597 2.2778 2.2385 2.0860 2.1995 2.1515 2.0872 2.2480 2.0274 2.0675 2.2791 2.2619 s = 19

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Table C-8. Run 2 Reactor 4 Solids Data 230 Sample Absorbance Hg Sample Wt. [Hg] Constant: 0.0234 I.D. (253.6) (ng) (9) (ng/g) X Coef.: Solids (g/g): 0.0005 0.1946 Blank 0.024 Std 1 (30 ng) 0.037 26 Std 2(50ng) 0.047 46 Std3(100ng) 0.072 96 Std 4 (200 ng) 0.124 200 Std 5 (300 ng) 0.168 288 R4:1 0.025 2.1475 R4;2 0.027 2.1775 R4:3 0.024 2.1594 R4:4 0.030 12 2.0454 31 R4:5 0.023 2.2670 R4:6 0.024 2.2837 R4:7 0.023 2.1719 R4;8 0.023 2.2611 R4:9 0.027 2.1226 R4:10 0.022 2.3264 R4:11 0.022 2.2300 R4:12 0.027 2.2163 R4:13 0.025 2.1784 R4:14 0.027 2.2038 R4:15 0.023 2.1583 R4:16 0.024 2.2084 R4;17 0.024 2.1756 R4;18 0.021 2.1203 R4:19 0.022 2.1953 R4:20 0.023 2.1711 R4:21 0.023 2.1183 R4:22 0.023 2.2010 R4:23 0.022 2.1576 R4:24 0.026 2.3460 R4:25 0.021 2.1380 R4:26 0.023 2.1362 R4;27 0.024 2.3002 R4:28 0.029 10 2.2480 23 R4:29 0.026 2.1216 R4:30 0.036 24 2.3122 54 R4:31 0.024 2.1351 R4:32 0.021 2.3050 R4;33 0.021 2.3044 R4:34 0.025 2.3006 R4:35 0.022 2.1848

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Table C-8 (cont'd). Run 2 Reactor 4 Solids Data 231 iample Absorbance Hg Sample Wt. [Hg] I.D. (253.6) (ng) (g) (ng/g) R4:36 0.024 2.1627 R4:37 0.022 2.1472 R4;38 0.023 2.1720 R4:39 0.023 2.3194 R4:40 0.025 Q 2.1805 R4:41 0.023 2.1360 R4:42 0.022 2.2090 R4:43 0.024 2.3822 R4:44 0.028 8 2.3453 18 R4:45 0.024 2.3813 R4:46 0.025 2.1761 R4:47 0.026 2.1417 R4:48 0.025 2.1311 R4:49 0.027 2.2131 R4;50 0.024 2.1390 R4:51 0.028 8 2.2818 18 R4:52 0.022 2.4150 R4:53 0.025 2.3895 R4:54 0.026 2.3360 R4:55 0.024 2.2062 R4:56 0.024 2.3615 R4:57 0.023 2.0280 S= 65 S= 149

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Table C-9. Run 2 Reactor 6 Solids Data 232 Sample Absorbance Hg Sample Wt. [Hg] I.D. (253.6) (ng) (g) (ng/g) Blank 0.025 Std 1 (50 ng) 0.048 46 Std 2(100ng) 0.072 94 Std 3 (200ng) 0.122 194 Std 4 (400 ng) 0.215 380 Std 5 (600 ng) 0.318 586 Std 6 (800 ng) 0,400 750 R5:1 0.022 2.3608 R5;2 0.093 136 2.2450 319 R5:3 0.028 6 2.2845 13 R5:4 0.027 2.3485 R5:5 0.025 2.1863 R5:6 0.031 12 2.3042 27 R5;7 0.025 2.3055 R5:8 0.034 18 2.2039 43 R5:9 0.027 2.2691 R5:10 0.023 2.3413 R5:11 0.028 6 2.2565 14 R5:12 0.025 2.1522 R5;13 0.030 io 2.3636 22 R5:14 0.039 28 2.2378 66 R5:15 0.026 2.2502 R5:16 0.025 2.3163 R5:17 0.031 12 2.2271 28 R5:18 0.026 2.3315 R5:19 0.022 2,3368 R5;20 0.106 162 2.1828 391 R5;21 0.028 6 2,3310 13 R5;22 0.025 2,1930 R5:23 0.025 2,3371 R5:24 0.038 26 2,3455 58 R5:25 0.034 18 2.3666 40 R5:26 0.028 6 2.2518 14 R5:27 0.023 2.2771 R5:28 0.025 2.2286 R5:29 0.019 2.3506 R5:30 0.035 20 2.2202 47 R5:31 0.026 2 2.2081 4 R5:32 0.024 2.2433 R5:33 0.030 10 2.2805 23 R5:34 0.025 2.3310 Constant: X Coef.: Solids (g/g): 0.0256 0.0005 0.1894

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Table C-9 (cont'd). Run 2 Reactor 6 Solids Data Sample Absorbance Hg I.D. (253.6) (ng) 233 R5:35 R5;36 R5:37 R5:38 R5;39 R5:40 R5:41 R5:42 R5:43 R5:44 R5:45 R5;46 R5:47 R5:48 R5:49 R5:50 R5:51 R5:52 R5:53 R5:54 R5:55 R5:56 R5:57 R5:58 R5:59 R5:60 R5:61 0.024 0.041 0.031 0.024 0.026 0.028 0.029 0.025 0.031 0.034 0.024 0.042 0.025 0.030 0.032 0.026 0.099 0.025 0.025 0.027 0.023 0.039 0.027 0.035 0.023 0.037 0.025 32 12 Q 6 8 12 18 Q 34 10 14 148 28 20 24 )mple Wt. [Hg] (g) (ng/g) 2.1866 2.2244 75 2.3476 27 2.3787 2.3279 2.1518 14 2.3727 17 2.3778 2.2977 27 2.3397 40 2.3154 2.2587 79 2.3248 2.2124 23 2.2572 32 2.3511 2.1904 356 2.3177 2.2376 2.3832 2.2391 2.2586 65 2.3763 2.3403 45 2.2929 2.2615 56 2.3906 s = 836 1973

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Table C-10. Run 2 Reactor 8 Solids Data 234 Sample Absorbance Hg Sample Wt. [Hg] Constant: 0.0319 I.D. (253.6) (ng) (g) (ng/g) X Coef.: Solids (g/g): 0.0005 0.1899 Blank 0.031 Std 1 (50 ng) 0.056 50 Std2(100ng) 0.080 98 Std 3 (200 ng) 0.128 194 Std 4 (400 ng) 0.232 402 Std 5 (600 ng) 0.320 578 Std 6 (800 ng) 0.420 778 R8:1 0.073 84 2.1500 206 R8:2 0.029 2.2181 R8:3 0.034 2.2604 R8:4 0.039 16 2.3271 37 R8:5 0.103 144 2.3542 323 R8:6 0.035 8 2.3535 18 R8:7 0.032 2.2280 R8:8 0.049 36 2.3033 83 R8:9 0.042 22 2.2947 51 R8:10 0.036 10 2.1613 25 R8:11 0.095 128 2.2594 299 R8:12 0.050 38 2.2228 90 R8;13 0.038 14 2.2787 33 R8:14 0.037 12 2.2453 29 R8:15 0.133 204 2.2328 482 R8:16 0.034 2.2104 R8:17 0.030 2.3255 R8:18 0.035 8 2.1697 20 R8:19 0.048 34 2.3735 76 R8:20 0.037 12 2.2705 28 R8:21 0.035 8 2.2066 20 R8:22 0.043 24 2.2648 56 R8:23 0.047 32 2.2121 77 R8:24 0.032 2.3024 R8:25 0.032 2.3088 R8:26 0.033 2.2367 R8;27 0.030 2.2064 R8:28 0.125 188 2.3203 427 R8:29 0.050 38 2.2082 91 R8:30 0.030 2.2426 R8:31 0.040 18 2.3155 41 R8:32 0.042 22 2.1952 53 R8:33 0.045 28 2.3446 63 R8:34 0.033 2.2000

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Table C-1 (cont'd). Run 2 Reactor 8 Solids Data 235 Jample Absorbance Hg Sample Wt. [Hg] I.D. (253.6) (ng) (g) (ng/g) R8;35 0.027 2.3482 R8;36 0.028 2.2170 R8:37 0.032 2.2822 R8;38 0.082 102 2.1272 253 R8;39 0.038 14 2.2235 34 R8:40 0.037 12 2.3386 27 R8:41 0.072 82 2.2065 196 R8:42 0.030 2.3453 R8:43 0.035 8 2.2123 20 R8:44 0.084 106 2.2884 244 R8:45 0.035 a 2.1901 20 R8:46 0.038 14 2.2575 33 R8:47 0.041 20 2.2717 47 R8:48 0.037 12 2.3780 27 R8:49 0.030 2.2684 R8:50 0.031 2.3021 R8:51 0.076 90 2.3770 200 R8:52 0.034 2.3290 R8:53 0.031 2.3272 R8:54 0.035 8 2,3782 18 R8;55 0.037 12 2.2447 29 R8:56 0.049 36 2.2947 83 R8:57 0.047 32 2.3041 74 R8:58 0.031 2.2687 R8;59 0.097 132 2.3300 299 R8;60 0.044 26 2.3272 59 R8:61 0.030 2.1633 R8:62 0.050 38 2.1581 93 S = 1889 S = 4 4384

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Table C-1 1 Run 2 Leachate data 236 Sample Absorbance Hg Sample Wt. [Hg] Constant: 0.0322 I.D. (253.6) (ng) (ml) (ng/g) X Coef.: 0.0004 Solids (g/g): 0.0052 Blank 0.033 Stdl(lOng) 0.037 10 Std2(30ng) 0.044 28 Std 3 (50ng) 0.052 48 Std 4 (80 ng) 0.062 73 Std 5 (100 ng) 0.073 100 Std 6 (200 ng) 0.112 198 Std 7 (300 ng) 0.153 300 Rl 0.029 100 R2 0.025 100 R3 0.026 100 R4 0.029 100 R4(dup) 0.030 100 R5 0.020 100 R6 0.044 28 100 53 R7 0.027 100 R8 0.032 100 R8(spike)* 0.050 43 100 82 89 % recovery R9 0.028 100

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Table C-12. Run 2 Activated Carbon Data Sample Absorbance Hg Sample Wt. [Hg] I.D. (253.6) (ng) (g) (ng/g) 237 Constant: X Coef.: 0.0315 0.0004 Blank Stdl(lOng) Std 2 (30 ng) Std 3 (50ng) Std 4 (80 ng) Std 5 (100 ng) Std 6 (200 ng) Std 7 (300 ng) ACa:l ACa:2 ACa:3 ACa:4 ACa:5 ACa:6 ACa:7 ACa:8 ACa:9 ACb:l ACb:2 ACb:3 ACb:4 ACb:5 ACb:6 ACb:7 ACb:8 ACb:9 0.033 0.037 10 0.041 20 0.050 42 0.063 75 0.071 95 0.112 197 0.150 292 0.026 4.0000 0.030 4.0000 0.033 4.0000 0.043 25 4.0000 0.046 32 4.0000 0.057 60 4.0000 0.063 75 4.0000 0.045 30 4.0000 0.071 95 4.0000 0.030 4.0000 0.030 4.0000 0.031 4.0000 0.033 4.0000 0.030 4.0000 0.045 30 4.9096 0.029 4.0000 0.035 4.0000 0.048 37 4.0000 = 383 Total: 6 8 15 19 7 24 6 9 94

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Table C-13. Run 2 Tygon Tubing Data Sample Absorbance Hg I.D. (253.6 nm) (ng) 238 Sample Wt [Hg] (g) (ng/g) Constant: 0.0214 X Coef.: 0.0005 Blank 0.022 Std1 (10 ng) 0.025 8 Std 3 (30 ng) 0.033 22 Std 4 (50 ng) 0.044 44 Std 5 (80 ng) 0.054 64 Std 6 (100 ng) 0.067 90 Std 7 (200 ng) 0.112 180 R1 0.023 R2 0.022 R3 0.022 R4 0.031 18 R5 0.030 16 R6 0.036 28 R7 0.026 8 R8 0.031 18 R9 0.024 4 s = 94

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APPENDIX D ABBREVIATIONS

PAGE 253

Abbreviations Hg MHg MMHg DMHg Hg2* Hg HgS S c Fe Cr Cd Pb Ni ZN CH4 CO2 NH3 CO H2S O2 N2 H VS MSW SW MWC DDDI GC CVAAS CVAFS GC-HS-MIP QFAAS ICPMS AED BCD SBW TCLP EPA MercuryMet hylmercury Monome t hy 1 me r c u r y Dime t hylmercury Mercury (II) Elemental mercury Mercuric sulfide Sulfur Carbon Iron Chromium Cadmium Lead Nickel Zinc Methane Carbon dioxide Ammonia Carbon monoxide Hydrogen sulfide Oxygen Nitrogen Henry's constant Volatile solids Municipal solid waste Solid Waste Municipal Waste Combustion Double distilled deionized water Gas chromatography Cold vapor atomic absorption spectrometry Cold vapor atomic fluorescence spectrometry Headspace gas chromatography with microwave induced plasma detection Quartz furnace atomic absorption spectrometry Inductively coupled plasma mass spectrometry Atomic emission detector Electron capture detector Slit band width Toxicity characteristic leaching procedure Environmental Protection Agency 240

PAGE 254

241 FDEP = Florida Department of Environmental Protection CFR = Code of Federal Regulations QA/QC = Quality assurance/quality CI' = Chloride OH" = Hydroxide

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BIOGRAPHICAL SKETCH Celia D.A. Earle was born on October 8, 1967, in Surrey England. She moved to Jamaica in 1970 and remained there until 1983. She moved to Gainesville, Florida, in 1983 and graduated from Gainesville High School in 1985. Celia immediately went to the University of Florida and received her Bachelor of Science degree in Microbiology and Cell Science in 1989. She worked as laboratory technician in a microbiology teaching laboratory from 1988 to 1989. She worked as an assistant chemist in a water quality laboratory at the University of Florida from 1989 to 1991, and entered the graduate program in the Department of Environmental Engineering Sciences in Summer 1991. She specialized in the water chemistry area and received her Master of Science degree in August 1993 She continued on with her Ph.D. in the same department up until May 1995 when she switched to the Soil and Water Science department to complete her Ph.D. Also, in May 1995, she decided to pursue a Bachelor of Science degree in environmental engineering. 262

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263 She received this second Bachelor of Science degree in August 1997. She will be receiving her Ph.D. in December 1997. She is currently a member of the Society of Women Engineers, Air and Waste Management Association, American Water Works Association and numerous other societies. She intends to work in the environmental consulting field as both an environmental engineer and a scientist.

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I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Plriiosophy. R. Dean Rhue Chair Professor of Soil and Water Science I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholaz'ly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy. -^^ic David P. Chyno^Veth, Professor of Agricultural and Bioloaical Enaineerina I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of D^;::^r of ^_Phil9soph]) ^seph C^Delfin^', Professor of Environmental Engineering Sciences I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of, Doctor of Philosophy. -lA Konda R. Reddy Graduate Research Professor of Soil and Water Science

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I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, a dissertation for the degree of Doctor, of Philosophy Lena Q. Ma, Assistant Professor of Soil and Water Science as I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality_ ^ as a dissertation for the degree of Docp^^^sof P)iilosoph:i TimotKyy^cT' Townsend, Assistant Professor of Environmental Engineering Sciences This dissertation was submitted to the Graduate Faculty of the College of Agriculture and to the Graduate School and was accepted as partial fulfillment of the requirements for the degree of Doctor of Philosophy. December, 1997 Dean, College of Agriculture Dean, Graduate School