Assessment of Mercury Methylation in Aquatic Sediments

A thesis submitted to the

Division of Research and Advanced Studies

of the University of Cincinnati

in partial fulfillment of the requirements for the degree of

MASTER OF SCIENCE

in the Department of Civil and Environmental Engineering

of the College of Engineering

2003

by

YI ZHOU

B.E., Chongqing University, 1992

Committee Chair: Dr. George Sorial

ABSTRACT

Mercury is one of the most hazardous contaminants distributed in the environment.

Inorganic mercury may be transformed to more toxic organic methylated species.

Methylmercury (MeHg), the most toxic form of organic mercury, is readily

bioaccumulated in aquatic food chain. Anoxic aquatic sediment has been considered to

be one of the most important mercury methylation sites. Scientists and researchers

generally agree that mercury methylation is primarily a microbial process catalyzed by

enzymes. Sulfate reducing bacteria (SRB) have been identified as the principal

methylators in aquatic sediments. A corrinoid protein, cobalamin (Vitamin B12), was

found to be the methyl carrier in SRB. Mercury methylation rate (MMR) is tightly

correlated with the sulfate reduction rate (SRR) and community composition of SRB. The

bioavailability of mercury for methylation depends on mercury speciation. Dissolved

neutral mercury complexes, such as HgS0 rather than free Hg2+, have been considered

more likely to be taken by bacteria. Mercury methylation is also influenced by a variety

of ambient factors like temperature, pH, salinity, organic matter, concentrations of sulfide

and sulfate. It was reported that 35°C was an optimal temperature for microbial

methylation in river sediments. Acidification of lake sediments resulted in a significant

decrease in methylation rates. There is a strong inverse relationship between the salinity

of anaerobic sediments and their methylation ability. Increasing MeHg concentrations

have been observed in sediments with elevating levels of organic carbon. However,

dissolved organic matter was also found to decrease the mercury methylation rate. There

is an optimal sulfate concentration range (0.2-0.5mmol/L) for mercury methylation.

Methylation and demethylation occur in aquatic systems. Some substances, such as

propyl iodine and group VI anions (MoO42-, SeO4

2-, TeO42- and WO4

2-), have been

observed to inhibit mercury methylation.

Key words: Methylmercury, mercury methylation, sulfate reducing bacteria, speciation,

demethylation, inhibition.

ACKNOWLEDGEMENTS

I gratefully acknowledge my advisor, Dr George Sorial. Without his patient guidance,

constant support and valuable suggestions, this extensive and critical literature review

research would not be accomplished. I would like to thank Dr. Makram Suidan and Dr.

Dionysios Dionysiou for serving on my committee and for their advices. I also would

like to thank my colleagues, Daekeun Kim and Qianrui Wang, for their help and

collaboration. Moreover, I would like to thank all the members of my research group,

Dinesh Palaniswamy, Hao Zhang, Pablo Campo, Ramakrishnan Balaji, Subhashini

Chandrasekar and Zhangli Cai, for their encouragement, suggestions and friendship.

i

TABLE OF CONTENTS

TABLE OF CONTENTS………………………………………………………………….i

LIST OF TABLES.............................................................................................................iii

LIST OF FIGURES………………………….……............................................................v

ACRONYMS……..……………………………………………………………………....vi

CHAPTER 1. Introduction and Project Objectives………………………………......…1

1.1. Introduction .……………………………………………………….1

1.2. Objectives………………………………………………………….3

CHAPTER 2. Mercury Methylation Mechanism and Pathways………....…………….4

2.1. Principal methylators……………………………………………...4

2.2. Mechanism………………………………………………………...6

2.3. Pathways……………..…………………………………………..10

CHAPTER 3. Mercury Methylation Reaction Rates……………...………………….19

3.1. Model describing the relationship between MMR and SRR ……19

3.2. Model estimating MMR based on SRR and SRB species…….....22

3.3. Model measuring MMR by using stable isotope tracer…...….….24

CHAPTER 4. Mercury Speciation……...………………...………………………….28

CHAPTER 5. Factors Affecting Mercury Methylation………………………...……38

5.1. Organic matter…..…………………………………....…………38

5.2. Complexation……………………….…...………………………38

5.3. Temperature………………………………..…………………….41

5.4. pH…………..……………………..………………………….….41

ii

5.5. Sulfide and sulfate…………………...………...………………...42

5.6. Salinity…………………...………..…………………………….42

CHAPTER 6. Inhibition of Mercury Methylation…….….………………………… 46

CHAPTER 7. Specific Site Study…………..………………………………………..50

7.1. Seine River ( France) ………….……....……………………….50

7.2. Carson River ( Nevada )……….……………………………….51

7.3. Everglades sediments ( Florida )....……………….……………53

7.4. Pine Barrens area ( New Jersey ).………….……...……………54

7.5. Pantanal floodplain ( Brazil )……..…...………………………..55

CHAPTER 8. Methylmercury Demethylation………….…………………………...58

CHAPTER 9. Bioavailability of Methylmercury…………….…………..…………60

CHAPTER 10. Conclusions and Recommendation………………………….. ……...62

10.1. Conclusions.….………………………………………………...62

10.2. Recommendation for future work....…………………...………63

REFERENCES………………………………...………………………………………...66

iii

LIST OF TABLES

Table

2.1. Effect of substrates and electron acceptors on the synthesis of CH3Hg+ (MeHg) in

sediments……..……………….………………………………………………….14

2.2. Enzymes and various physiological reactions involved in D. desulfricans

extracts…………………………………………………………………………...15

4.1 Mercury-sulfide complexes and equilibrium constants (Kf) in the Speciation

models for dissolved Hg in the presence of excess cinnabar…………………….33

4.2. Mercury-sulfide complexes and equilibrium constants (Kf ) used in the speciation

models for dissolved Hg with sorption to the solid phase……………………….34

4.3. Reactions and constants for mercury-sulfide interactions……………………….35

4.4. Reactions used to explain the cinnabar solubility………………………………..36

4.5. Formation reactions and constants for Hg-Sx species…….……..……………....37

5.1. Reactions and constants for MeHg and ligands………………………………….44

iv

5.2. Equilibrium binding constants and binding capacities for formation of MeHg

complexes with humic acids……………………………………………………..45

6.1. Effect of various sulfate concentration on the methylation rate inhibited by group

VI anions…………………………………………………………………………49

7.1. Physiochemical characteristics of three Pine Barrens lakes……………………..57

v

LIST OF FIGURES

Figure 2.1. Synthesis of MeHg in anoxic estuarine sediments slurry spike with Hg 2+

and the effects on the process by BESA and Na2MoO4………...………16

Figure 2.2. Evolution of the methane from anoxic estuarine sediments slurry

and the effects on the process by spiking BESA and Na2MoO4……..….17

Figure 2.3. Proposed metabolic pathway involved in mercury methylation by

Desulfovibrio desulfuricans LS...…………………………………….…18

vi

ACRONYMS

AMP – Adenosine Monophosphate

ADP – Adenosine Diphosphate

ATP – Adenosine Triphosphate

BESA – 2-Bromoethane sulfonate

CODH – CO Dehydrogenase

DOM – Dissolved Organic Matter

Fd – Ferredoxin

FDH – Formate Dehydrogenase

Metr – Methyltransferase

MeHg – Methylmercury

MMR – Mercury Methylation Rate

NAD+ – Nicotinamide Ademine Dinucleotide

NADP+ – Nicotinamide Ademine Dinucleotide Phosphate

SRB – Sulfate Reducing Bacteria

SRR – Sulfate Reduction Rate

SHMT – Serine Hydroxymethyltransferase

THF – Tetrahydrofolate

dw – dry weight

1

CHAPTER 1

Introduction and Project Objectives

1.1. Introduction

Mercury is a very special heavy metal. Elemental mercury (Hg0) is the only liquid metal

under room temperature. It is a volatile silver white metal which melts at -38.7 °C.

Mercury is one of the hazardous contaminants. It is reported that the oral LD10 of Hg is

1429 mg/kg (man) (Langford and Ferner 1999). The most common natural oxidized

states of mercury, mercuric sulfide or cinnabar, has been mined and processed throughout

the industrial age for various uses, including medical applications. Because of its wide

use, the amount of mercury mobilized and released into the open environment has

steadily increased over years.

Mercury can exist in three valence states (0, +1, +2). The chemistry of Hg is complex; it

can exist in various forms under different physiochemical conditions. Inorganic mercury

can be converted into organic mercury. Among them, methylmercury (MeHg) are the

most toxic species. According to the Mercury Study Report to Congress (USEPA 1997),

the reference dose (RfD) was 0.1µg/kg/day. MeHg usually exists as methylmercuric salt

such as methylmercuric chloride (CH3Hg+Cl-) and methylmercuric hydroxide

(CH3Hg+OH-) rather than a free ion.

2

MeHg was recognized as a serious health hazard and attracted the interest of researchers

after the outbreak of the neurological disease occurred in the 1950s at Minamata Bay,

Japan (Choi and Bartha 1993). A local chemical plant used mercury sulfate as a catalyst

to produce acetaldehyde and discharged wastewater into Minamata bay. The MeHg that

was produced as a byproduct was accumulated into fish. The incident caused 2,200

poisonings with 750 fatalities due to consumption of the MeHg contaminated fish

(Sakamoto 1991).

MeHg is readily accumulated in aquatic biota due to its lipophilic and protein binding

properties. It thus, poses a health threat to humans via food chain. Mercury methylation

is a very complex process. It can be an abiotic course (Nagase 1988; Nagase 1986;

Weber 1993) or a biotic process. More than 90% environmental mercury methylation is

associated with biological activities (Berman and Bartha 1986). (Jensen and Jernelov

1969) first reported the formation of MeHg by microorganisms. They demonstrated a

microbial methylation of mercuric chloride to MeHg by using mixed cultures from

aquatic sediment and from decaying fish. Aquatic sediment has been found the most

important site for mercury methylation and most methylation activities occur at the

interface between water column and anoxic sediments (Gilmour 1991; Korthals and

Winfrey 1987;Watras et al. 1995).

Considerable efforts regarding mercury methylation have been conducted so far. The

principal methylators have been identified, methylation pathways were proposed and

environmental factors that affect mercury speciation and bioavailability were studied.

3

Mercury methylation, however, is still not fully understood. Mercury cycling from the

sediment compartment to other media and ecological receptors is influenced by the

formation of MeHg. Mercury methylation is thus a very important step in the cycling of

mercury. In order to effectively manage and control the methylmercury pollution, better

understanding of mercury methylation is necessary.

1. 2. Project objectives

The purpose of this research is to conduct a critical and extensive literature review to

better understand mercury methylation. Specific objectives of this project include:

• identification of the role of microorganisms involved in mercury methylation.

• description of the chemical and biological pathways of MeHg formation in

aquatic sediments.

• description of the reaction equations for MeHg formation.

• determination of the factors that influence the rates of reaction.

• identification of the rate limiting steps under conditions likely to occur in aquatic

sediments.

4

CHAPTER 2

Mercury Methylation Mechanisms and Pathways

Mercury (Hg), one of the most hazardous contaminants, is widely distributed in aquatic

systems. Methylmercury(MeHg), the most toxic form of mercury, is readily accumulated

by aquatic biota and thus pose a threat to fish-eating animals and humans. In the 1950s,

a severe incident occurred at Minamata Bay, Japan (Choi and Bartha 1993). The incident

caused 2,200 poisonings with 750 fatalities due to consumption of the MeHg

contaminated fish(Sakamoto 1991). Since then, MeHg was recognized as a serious

health hazard. To better understand mercury methylation in aquatic sediments, an

extensive literature review is performed. In this chapter, the principal methylators are

identified, and the mechanisms of mercury methylation are described. In addition, the

reaction equations and pathways of MeHg formation are presented.

2.1. Principal methylators

The conversion of inorganic mercury into methylmercury is a critical step in the cycling

of mercury in aquatic systems. Aquatic sediment has been considered to be one of the

most important mercury methylation sites (Mauro et al. 1999). Mercury methylation is a

very complex process which can be abiotic or biotic process. (Jensen and Jernelov 1969)

were the first to report that MeHg is formed in aquatic sediments by sediment

microorganisms. (Berman and Bartha 1986) demonstrated that the MeHg levels resulting

from biochemical methylation in anoxic sediments were approximately one order of

magnitude higher than those formed by chemical mercury methylation.

5

Microorganisms capable of forming MeHg have been found among aerobes, anaerobes

and facultative anaerobes, but the higher potential of microbial methylation generally

appears under anaerobic conditions(Robinson 1984). Anoxic microbial metabolism

includes fermentative, nitrate reducing, sulfate reducing and methanogenic. Fermentation

products serve as substrates for organisms using three other types of metabolism(nitrate

reducing, sulfate reducing and methanogenic). The products of fermentative metabolism

are oxidized while the electrons generated reduce nitrate, sulfate, or carbonate at

progressively decreasing redox potentials.

In order to identify the metabolic groups that are actually responsible for MeHg synthesis

in anoxic sediments, (Coppeau and Bartha 1985) performed a series of experiments on

anoxic, low salinity (0.4%) and neutral pH (pH=6.8) estuarine sediment samples. The

sediment samples were collected at a depth of 10~ 30 cm in Cheesequake State Park,

N.J. Mercury was added as mercuric chloride (HgCl2 ) at 75µg/g of dry sediment and the

samples were kept under strictly anaerobic conditions. At first, various substrates with or

without electron acceptors were spiked into samples, the sediment samples were

anaerobically incubated at 25 ºC for 2 days. The sediments were then extracted and

MeHg was determined by gas chromatography. The effect of substrates are shown in

Table 2.1. MeHg synthesized by pyruvate was found more than three times higher than

those by other substrates. All other substrates suppressed MeHg synthesis. In a H2-CO2

gas atmosphere (20% H2 and 80% CO2) for enriching methanogenic activity, no MeHg

was detected in sediment samples. Pyruvate can be utilized by sulfate reducers

6

fermentatively. Therefore, sulfate reducers were suggested to be the principal

methylators.

A different approach was conducted to determine the contribution of various metabolic

groups to MeHg synthesis. In this stimulation-inhibition experiment, the incubated Hg2+

sediment samples were spiked with 30mM 2-Bromoethane sulfonate(BESA) or 20mM

sodium molybdate (Na2MoO4). The samples were then incubated at 25 ºC for 72 hours.

The results indicated that BESA, a specific inhibitor of methanogens, stimulated MeHg

formation; whereas Na2MoO4, a specific inhibitor of sulfate reducers, suppressed

mercury methylation by more than 95% (see Figure 2.1). To ensure the effectiveness and

specificity of the inhibitors, methane evolution was measured in the same experiment (see

Figure 2.2). Subsequent enrichment and isolation procedures clearly identified the sulfate

reducer as a strain of Desulfovibrio desulfuricans LS, where LS refers to low salinity.

The complete inhibition of mercury methylation by Na2MoO4 and the stimulation of

mercury methylation by BESA in anoxic sediment pointed sulfate reducers as the primary

methylators. A substantial amount of field information also supports this conclusion. For

example, (Macalady et al. 2000) observed mercury methylation activity in sediments is

often significantly correlated with distribution of sulfate reducing bacteria (SRB)

populations.

2.2. Mechanism

Although SRB were identified as the principal environmental methylators and strains of

Desulfovibrio desulfuricans have already been isolated and described, the mechanisms of

7

mercury methylation by these microorganisms need to be further studied. (Landner

1971) and (Wood et al. 1968) suggested that methylcobalamin, a vitamin B12 derivative

produced by many organisms, is involved in the microbial mercury methylation. The

methyl transfer to Hg2+ is a carbanion (CH3-) process (DeSimone et al. 1973). Although

there are a lot of methyl donor molecules in the aquatic system, the prevalence of

methylcobalamin in anaerobic ecosystems and living organisms makes it the most likely

source for environmental mercury methylation, and it is thought to be the only natural

methylating agent capable of transferring methyl groups as carbanions by (Ridley et al.

1977).

(Choi and Bartha 1993) conducted a series of experiments to study the methylation

activity and the methyl carrier of Desulfovibrio desulfuricans. At first, they investigated

MeHg formation under different culture conditions, which is fermentative or sulfate-

reducing condition. Desulfovibrio desulfuricans strains were isolated from the anoxic,

low salinity salt marsh sediment in Cheesequake State Park, N. J. The media used for

bacteria growth were Postgate’s lactate-sulfate medium and pyruvate medium,

respectively. Appropriate concentration of HgCl2 were spiked into the media. After

incubation at 27 ºC for 1 ~ 2 days under fermentative or sulfate-reducing condition, the

cell suspensions were extracted and MeHg produced were measured with a gas

chromatography. Under fermentative condition, in which there are no terminal electron

acceptors, up to 37% of 0.1µg/ml HgCl2 was methylated, but only1.5% of 10.0µg/ml

HgCl2 was methylated; cell growth and methylation activity were significantly inhibited

when the concentration of HgCl2 was higher than 10.0µg/ml. While under sulfate

8

reducing condition, less than 1% of the added HgCl2 was methylated, concentration of

methylmercury produced gradually declined when levels of HgCl2 added were higher

than 25µg/ml. Comparing with sulfate reducing condition, there was a high degree of

mercury methylation and low degree of mercury tolerance under fermentative condition.

A second set of experiments were then conducted aiming to determine the cobalamin in

Desulfovibrio desulfuricans LS. 57Co label (CoCl2 ) was incorporated into cultures of

Desulfovibrio desulfuricans LS and the cultures were incubated under strictly anaerobic

conditions at 27 ºC for 4 days. After extraction and purification procedures, a corrinod

extract was analyzed by high performance liquid chromatography (HPLC). The analysis

of the corrinoid yielded a single peak with the retention time matching cobalamin, and

97% of the 57Co radioactivity was associated with the peak. Fast atom bombardment

and UV- visible spectra of the isolated corrinoid also matched those of cobalamin. These

results suggested that cobalamin was the only corrinoid in Desulfovibrio desulfuricans.

In order to identify the role of cobalamin in the mercury methylation, a third set of

experiments were performed. The isolated cobalamin was methylated by 14CH3I, then

the 14CH3 -cobalamin formed was reacted with mercuric ions at pH 4.5 with acetate

buffer. The corresponding specific activity ratio (specific activity of the methylmercury

produced / specific activity of the added 14CH3I ) was 93.9% , suggesting methyl

groups were transferred spontaneously to Hg2+ by the isolated cobalamin from

Desulfovibrio desulfuricans LS. Methylation rates, however, were observed three times

lower at pH 7. Various anions such as Cl - and HCO3 - were found to significantly

9

interfere with the spontaneous mercury methylation process. Therefore, mercury

methylation under physiological conditions may be an enzymatically catalyzed process

rather than a spontaneous chemical reaction .

Aiming to determine whether mercury methylation is a spontaneous or an enzymatically

catalyzed process, (Choi et al. 1994a) investigated both enzymatic mercury methylation

and nonenzymatic mercury methylation by using cell extract of Desulfovibrio

desulfuricans. The D. desulfuricans strain used in this study was still isolated from the

anoxic, low salinity salt marsh sediment in Cheesequake State Park, N.J. 14C radiolabel

was incorporated into 5-14CH3-tetrahydrofolate (5-14CH3 -THF) to determine the

enzymatic production of methylmercury. Label 57Co (57CoCl2) was added into the

medium to monitor corrinoid protein. The experimental results indicated that just 5-

14CH3 - THF alone or in combination with cobalamin can not methylate HgCl2.

However, methylmercury production was observed with the addition of cell extracts. In

cell extracts of D. desulfuricans, over 95% of the 57 Co label was associated with

macromolecules rather than free cobalamin. A single corrinoid protein of 40- kDa size

was identified by gel filtration and electrophoresis of cell extracts. Under reducing

condition, cell extracts containing the corrinoid protein formed MeHg from 5- 14CH3 -

THF and Hg 2+. When cells of D. desulfuricans were preincubated with propyl iodide,

their ability to form MeHg from Hg 2+ was blocked. According to these findings, the 40-

kDa corrinoid protein is proposed to be the in vivo methyl carrier in Desulfovibrio

desulfuricans. The synthesis of methymercury were divided into the following steps.

10

14CH3 - THF + Co-protein methyltransferase I 14CH3 - Co-protein

14CH3 - Co-protein + Hg 2+ methyltransferase II Co-protein + 14CH3 Hg+

Where Co-protein represents corrionoid protein and methyltransferase I & II are

enzymes.

2.3. Pathways

(Choi et al. 1994b) performed experiments to investigate formation pathways of

methylmercury (MeHg) by Desulfovibrio desulfuricans. Desulfovibrio desulfuricans

was isolated and incubated on lactate-formate medium. During the exponential growth of

cells, [14C ] formate and HgCl2 were added simultaneously. After 2 days incubation at

37ºC, methylmercury formed was extracted and quantified by gas chromatography.

Incorporation of 14C was measured by liquid scintillation counting of samples of the

MeHg extracts. High rates of 14C incorporation into MeHg from [ 14C ] formate and

serine prompted the assay of enzymes of the acetyl coenzyme A (CoA) synthase

pathways. The enzymes involved in this pathway include methyltransferase, CO

dehydrogenase, acetyl-CoA synthase and the THF pathway enzymes as well as formate

dehydrogenase and corrinoid protein. The enzymes and the physiological reactions for

various enzyme activities in Desulfovibrio desulfuricans LS extracts are summarized in

Table 2.2.

11

Propyl iodine, an inhibitor of methylmercury formation, inhibited acetyl-CoA synthase in

the experiments. Hg2+ was found as an inhibitor of acetyl-CoA synthase due to its

competition with methyl groups. On the basis of above experimental results and the

presence of the enzymes of acetyl-CoA synthase, (Choi et al. 1994b) proposed pathways

for methylmercury formation (see Figure 2.3). They suggested that in MeHg synthesis

by Desulfovibrio desulfuricans, the methyl group is transferred from CH3-

tetrahydrofolate via methylcobalamin. The CH3 group may originate from formate, C-3

of serine, or from C-3 of pyruvate via serine and be transferred to cobalt in a corrinoid

protein by a methyltransferase. In addition, the CH3 group could arise from C-3 of

pyruvate by its oxidation to acetyl-CoA, cleavage of acetyl-CoA by CO dehydrogenase,

and transfer to the corrinoid protein. Many sulfate-reducing bacteria, i.e.,

Desulfotomaculum acetoxidans, Desulfobacterium autotrophicum, Desulfococcus

multivorans and Desulfosarcina variabilis, utilize the acetyl-CoA pathways in their

metobalism. They use the acetyl-CoA pathway in reverse, for oxidation of acetate. More

incorporation of 14C from [3-14C] serine and H14COO - in CH3 Hg+ than from [3-14C]

pyruvate suggests that in Desulfovibrio desulfuricans LS the pathway is proceeding in

the direction of acetyl-CoA synthesis. If the CH3- Corrinoid were formed from pyruvate

via reversal of the acetyl-CoA synthase reaction, then [3-14C] pyruvate would yield

14CH3 Hg + with higher efficiency.

Sulfidogens are capable of methylating Hg2+ in anoxic aquatic sediments. The rate of

Hg2+ methylation is far lower in high-sulfate estuarine sediments than the rate of Hg2+

methylation in low-sulfate freshwater sediments (Compeau and Bartha 1987). One

12

explanation for this phenomenon appears to be the generation of H2S by the Hg2+

methylating sulfidogens. (Pak and Bartha 1998b) measured high rates of Hg2+

methylation in oligotrophic lake sediments that are free of detectable H2S and evolved

methane vigorously. A series of inhibition studies excluded methanogens and implicated

sulfidogens in Hg2+ methylation. This led to question how sulfidogens stay active and

methylate Hg2+ in a sulfate limiting sediments. In order to explore the mechanism of

such transfers on the methylation of Hg2+, (Pak and Bartha 1998a) modeled the

conditions by using pure cultures of sulfidogens and a methanogen. A specially

formulated sulfate-free lactate medium was provided as coculture medium to incubate

Desulfovibrio desulfuricans and Methanococcus maripaludis. Neither D. desulfuricans

nor Methanococcus maripaludis was observed to grow individually in the coculture

medium. However, both of them grew in the coculture when they were incubated

simultaneously in the medium. Methanococcus maripaludis alone failed to form any

methylmercury, and Desulfovibrio desulfuricans alone formed only a trace amount, but

the coculture produced 22 ng of methylmercury per µg of initial protein in 8 days,

methylating 2.6% of the available Hg2+. The coculture experiment was repeated by

replacing Desulfovibrio desulfuricans LS with Desulfovibrio desulfuricans ND 132 ,

and the results were only slightly different. Desulfovibrio desulfuricans strains oxidize

lactate to pyruvate, and pyruvate yields CO2, acetate, but they are unable to use acetate.

Methanococcus maripaludis, like many other methogens, can use acetate as a substrate

for methanogensis and growth . Therefore, Desulfovibrio desulfuricans strains not only

can transfer lactate hydrogens to Methanococcus maripaludis for CO2 reduction, but also

can produce acetate as a methanogenic substrate. The removal of acetate and H2 by

13

Methanococcus maripaludis allowed Desulfovibrio desulfuricans strains to use lactate in

the absence of sulfate. The interspecies hydrogen and acetate transfer between

sulfidogens and methanogens provided a pathway for mercury methylation in low sulfate

anoxic freshwater sediments. The experiments were performed under growth and

nongrowth conditions, respectively. Almost no mercury methylation occured under

nongrowth conditions, whereas mercury methylation was observed during bacteria

growth. It suggests that growth conditions are necessary for mercury methylation.

14

Table 2.1. Effect of substrates and electron acceptors on the synthesis of CH3Hg+

( MeHg ) in sediments

Supplement CH3Hg+ produced

( ng/g of sediment ) a

None 62 (10)

Glucose ND

Acetate 40 (6)

Pyruvate 203 (12)

Lactate 40 (7)

Glucose-nitrate 13 (5)

Lactate-sulfate 30 (8)

a Average of duplicate determinations. The number in parentheses represent one- half of the range

between duplicate determinations. ND, none detected.

Source from (Coppeau and Bartha 1985).

15

Table 2. 2. Enzymes and various physiological reactions involved in D.desulfuricans

extracts.

Source from (Choi et al. 1994b).

Enzyme Reaction catalyzed

Hydrogenase H2 + 2 cytochrome c3 = 2 cytochrome c3 reduced + 2 H +

Carbon monoxide

dehydrogenase

CO + 2 Fd + H2O = HCO3 - + 2 Fd reduced + H +

Serine

hydroxymethyltransferase

Serine + THF = glycine + methylene- THF

Formate dehydrogenase

HCOO - + 2 cytochrome c3 = CO2 + 2 cytochrome c3 reduced + H +

N10- formyl-THF synthetase

HCOO - + THF + ATP = 10-formyl- THF + ADP + Pi

N5, N10- methenyl –THF

cyclohydrolase

10-formyl- THF + H + = methenyl- THF + H2O

N5, N10- methylene-THF

dehydrogenase

Methylene - THF + NADP + = methenyl-THF + NADPH

N5, N10- methylene -THF

reductase

Methylene-THF + 2Fd red + 2H + = methyl-THF + 2Fd

Acetyl-CoA synthase

CH3-THF + CO + CoA-SH = CH3COSCoA + THF

16

0

100

200

300

0 2 4 6 8 10 12

Time ( Days )

MeH

g (

ng

)/g

sed

imen

t

Na2MoO4

BESA

Spiked Sedim ent

Figure 2.1. Synthesis of MeHg in anoxic estuarine sediments slurry spiked with Hg2+ at 75µg/g (dry weight) and the effects on the process by 30mM BESA and 20mM Na2MoO4, specific inhibitors of methanogensis and sulfate reduction, respectively. Error bars represent the range of measurements on duplicate samples and are omitted when smaller than the symbols. Source from (Coppeau and Bartha 1985).

17

0

0.5

1

1.5

0 2 4 6 8 10 12Time ( Days )

CH

4 ( m

icro

mol

es )

/ g s

edim

ent

Na2MoO4

Sedim ent

Hg2+

BESA

Figure 2.2. Evolution of the methane from anoxic estuarine sediments slurry and the effects on this process by spiking Hg2+ at 75µg/g (dry weight), 30 mM BESA plus Hg2+, and 20 mM Na2MoO4 plus Hg2+. Error bars represent the range of measurements on duplicate samples and are omitted when smaller than the symbols. Source from (Coppeau and Bartha 1985).

18

NADH NAD+

CH3-CO-COO - CH3-CHOH-COO -

ATP Fd FdH2 AMP Pi FdH2 CH3-CO-SC0A CH2= CO(P)-COO - [CO2] CODH NAD(P)H Fd NAD(P) + [CO] CH3-CO-SC0A

HOCH2-CHO(P)-COO - HCOO - CoA-SH ATP THF Acetyl-CoA synthase ADP + Pi (P)OCH2-CHOH-COO- N10-HCO-THF [ H+

]

CH3-Hg+ (P)OCH2-CO-COO - N5 ,N10=CH-THF NADPH Propyl iodine (P)OCH2-CH(NH2)-COO - NADP

+ Hg 2+

Pi CH2(NH2)-COO - CH3-Corrin

FdH2 Fd MeTr HOCH2-CH(NH2)-COO - N5, N10-CH2-THF N5-CH3-THF SHMT THF THF Co-Corrin Co-Corrin FdH2 Fd Fig 2.3. Proposed metabolic pathway involved in mercury methylation by Desulfovibrio desulfuricans. Where AMP, adenosine monophosphate; ADP, adenosine diphosphate; ATP, adenosine triphosphate; NAD+, nicotinamide ademine dinucleotide; NADP+, nicotinamide ademine dinucleotide phosphate; Fd , ferredoxin; FDH , formate dehydrogenase; CODH, CO dehydrogenase; Metr; methyltransferase; SHMT, serine hydroxymethyltransferase. Source from(Choi et al. 1994).

19

CHAPTER 3

Mercury Methylation Reaction Rates

Sulfate reducing bacteria (SRB) have been identified as the primary methylators in

aquatic sediments. Mercury methylation rate (MMR) is naturally correlated with the

sulfate reduction rate (SRR) and SRB species. To quantify the MMR, three models are

presented in this chapter.

3.1. Model describing the relationship between MMR and SRR

Anoxic sediment slurry incubations were performed by (King et al. 1999) to examine the

correlation between MMR and SRR in salt marsh sediments from Savannha, Georgia,

USA. Both the maximum MMR and SRR were found located in the top 4cm of the

sediments and decreased similarly with respect to depth. A series of stimulation-

inhibition incubations were performed. MMR increased with the addition of substrate

(100 mM pyruvate or 100 mM acetate) which stimulate SRB. Whereas MMR changed

insignificantly with the addition of 100 mM molybdate, which is an inhibitor of SRB.

This information suggested that a correlation exists between MMR and SRR. The

sediments were also treated with various concentrations of inorganic mercury. Results of

incubation illustrated that the MMR observed in the initial 12 hours depended on the

concentrations of inorganic mercury added to the sediments. The sediments were also

incubated at three temperatures, namely, 4, 25 and 37 ºC, to examine the influence of

temperature. The MMR for the initial 12 h increased with temperature. The rates

measured at 25 and 37ºC exceeded that at 4 ºC by factors of 2 and 30. MMR was

20

observed significantly slower after the initial 12 hours following inorganic mercury

addition suggested that sorption or precipitation reduced the availability of mercury for

methylation. A preliminary model describing the relationship between MMR and SRR

was developed based on the Michaelis-Menten equation which is given by

MMRMMR Hg

K HgHg=

+

+

++

max[ ]

[ ]

2

22 ( 3.1 )

Where MMR max represents maximum mercury methylation rate; [Hg2+] represents

aqueous inorganic mercury concentration; KHg2+ represents the inorganic mercury

concentration at which the reaction is half the maximum value. The developed model

utilized the dose of inorganic mercury concentration added to the slurry sediments

replacing aqueous [Hg2+] concentration. The alternative concentration is represented by

[Hg2+]a . MMR max was assumed to be proportional to SRR by the following relationship.

MMR max = [ constant] ( SRR ) ( 3.2 )

Applying a nonlinear, hyperbolic-fit program to the data in a plot of MMR versus [Hg2+],

the KHg value was determined to be 1,575 ng/g Hg2+ and MMR max was 648.5 pg/g.h.

Through a process of substitutions and rearrangements, the constant term that related the

SRR to the MMR max was determined to be 134.8 pg/nmol. Equation 3.3 illustrates the

mathematical model for predicting MMR in terms of the SRR for the initial 12 h after

inorganic mercury addition

21

MMR pg nmol SRR

Hg

ng g Hg=

+

+

+[ . / ]( )[ ]

[ , / ] [ ]134 8

1575

2

2 ( 3.3 )

A comparison of predicted MMR and the measured MMR indicated that the model

equation provided a good correlation for data reported with an SRR less than 30

nmol/g.h. The greatest discrepancy between predicted and observed MMR occured under

the conditions that are not encountered in situ. For example, the use of high

concentrations of substrates such as acetate or pyruvate (100 mM), respectively, could

affect normal cellular processes. The faster SRR generated larger discrepancies in

predicted and measured MMR.

In pure-culture studies, researchers primarily utilized one SRB, Desulfovibrio

desulfuricans, to determine the Hg methylation potential of the whole SRB population.

SRB capable of mercury methylation are currently thought to be much more

physiologically and phylogenetically diverse than originally thought. Each

phylogenetically distinct group could have a different potential to methylate Hg on a per-

cell basis. (King et al. 2000) conducted a research to determine if different SRB strains

methylate mercury at similar rates. Pure cultures of five genera of the SRB

(Desulfovibrio desulfuricans, Desulfobulbus propionicus, Desulfococcus multivorans,

Desulfobacter sp strain BG-8, and Desulfobacterium sp strain BG-33 ) were grown in a

strictly anoxic, minimal medium that received a dose of inorganic Hg 120 hours after

inoculation. Because all five SRB strains exhibited a lag in methylmercury production in

the first 24 h after inorganic Hg was added, the methylation rates (MMR) was obtained

for each group based on a plot of the methylmercury concentration versus time from 24

22

to 96 h. Sulfate reducing consortia was identified by using group specific oligonucleotide

probes that targeted the 16SrRNA. The rates at which SRB methylated Hg were

determined to be in the following order: Desulfobacterium>>Desulfobacter=

Desulfococcus>> Desulfovibrio = Desulfobulbus. The MMR normalized per cell were

up to 3 order of magnitude higher in pure culture members of SRB groups capable of

utilizing acetate( e.g., the family Desulfobacterionaceae ) than in pure cultures that are

not capable of utilizing acetate (e.g., the family Desulfovibrionaceae). Almost no

methylation was observed in cultures of Desulfobacterium or Desulfovibrio strains

without the presence of sulfate, indicating that Hg methylation was coupled to respiration

in these strains. The differential Hg methylation rates may be explained by the presence

of constitutive and induced methyl transferase pathways. SRB completely oxidize the

acetyl group of acetyl coenzyme A to CO2 by two entirely different mechanisms.

Desulfobacter strains employ the citric acid cycle, while Desulfobacterium strains

employ the carbon monoxide dehydrogenase pathways.

3.2. Model estimating MMR based on SRR and SRB species

As mentioned above, SRB are very abundant in marine sediments and different strains of

SRB methylate mercury at variable rates. With the help of new genetic method, the

composition of SRB group has been elucidated by using oligonucleotide probes that

target the 16S rRNA. A quantitative framework was developed by (King et al. 2001) to

estimate mercury methylation rates in marine sediment cores based on measured sulfate

reduction rates and the community composition of SRB. Equation 3.4 defines the MMR

normalized to the SRR as the net incidence term, f .

23

MMR

SRRf

Hg

K Hgf

Hg=

+=^[

[ ]

[ ]]

[ ] ( 3.4 )

Where f ^ represents a rate constant; K [Hg] represents a half-saturation constant for

cellular internalization / transport of mercury for methylation; the net rate function ( f )

includes both a rate constant ( f ^ ) and a Hg bioavailability term. Since the metabolic

activity is potentially different for each phylogenetic group, the individual contribution of

each group is considered in equation 3.5.

MMR total = ∑ MMR i ( 3.5 )

Therefore, the relationship between individual MMRi and SRRi can be rewritten from

equation 3.4.

MMR f SRR

Hg

K Hgf SRRi i i

ii i

Hg

=+

=^ [[ ]

[ ]]

[ ]

( 3.6 )

Equation 3.7 illustrates the calculation of the individual contributions of SRRi

SRR SRRSrRNA SRR

Celli

SrRNASRR

Celli

i total

ipc

pc

ipc

pc

=∑

16

16

( 3.7 )

Where 16SrRNA represents the gene associated with SRB; SRRpc / Cellpc represents

the SRR observed in pure culture normalized to cell number. Equation 3.8 defined the

24

total MMR in sediments as the sum of individual MMRi contributions from the various

SRB phylogenetic groups found within the sediment cores.

MMR f SRR

SrRNASRR

Celli

SrRNASRR

Celli

i total

ipc

pc

ipc

pc

=

∑

∑

16

16

( 3.8 )

Using the field data collected in saltmarsh sediment where sulfate reduction activity is

high, calculated and measured MMR results were consistently within one order of

magnitude. The calculated and measured MMR diverged by greater than one order

magnitude in an estuarine sediment where sulfate reduction activity was low. The

quantitative framework elucidates the coupling of mercury methylation to sulfate

reduction based on the calculated rates of mercury methylation on the activity and

community composition of SRB.

3.3. Model measuring MMR by using stable isotope tracer

Methylmercury (MeHg) demethylation also occurs in aquatic sediments, environmental

MeHg concentrations measured actually reflect net methylation rather than actual rates of

mercury methylation. Traditionally, methylation and demethylation studies have been

conducted using radiotracers. Commercially available tracers such as 203Hg exhibit low

specific activities. In order to see any experimental effect, sediments had to be spiked

with radiotracer at high level which were usually well above natural background

concentration for MeHg ( Korthals, 1987., Kerry, 1991., Steffan, 1988., and Oremland,

1991). Thus, it was questioned whether the results obtained were indeed representative

for the behavior of the ambient mercury species. An alternative method was developed

25

to measure methylation and demethylation rate constants simultaneously in aquatic

samples(Hintelmann et al. 2000). Solutions containing stable isotope tracer of 199Hg2+

and CH3202Hg+ were spiked into lake sediments. The spike tracer increased the

concentrations of total Hg and MeHg by only 10 to 80%. The formation of CH3199 Hg+

and the decrease in CH3202Hg+ were measured simultaneously in time series experiments

using gas choromotography and by isotope-specific detection inductively coupled plasma

mass spectrometry (ICP-MS). A model based on first-order kinetics for methylation and

demethylation was used to describe these transformation reactions. The net production of

MeHg is given by the following equation.

d CH Hg

dtK Hg K CH Hgm d

[ ][ ] [ ]

3 199 23

199

199

++ += − ( 3.9 )

Where [ CH3199 Hg+ ] represent concentration of CH3

199 Hg+ newly generated from the

199Hg 2+ tracer ( in ng/g), [199Hg2+] represent concentration of added 199Hg 2+ (in ng/g ),

Km represents specific methylation rate constant (in d-1), Kd represent specific

demethylation rate constant (in d -1), t represents incubation time ( in days ). For those

data for which [ CH3199 Hg+ ] is low enough so that the second term in equation 3.9 was

ignored. Thus, equation 3.9 can be simplified to following equation.

d CH Hg

dtK Hgm

[ ][ ]

3 199 2199 +

+= ( 3.10 )

Integration of equation 3.10 leads to

26

[ CH3199 Hg+ ] = Km [ 199Hg 2+ ] t ( 3.11 )

K

CH Hg

Hg tm =

+

+

[ ]

[ ]

3

199 2

199

( 3.12 )

When CH3202Hg+Cl internal standards was spiked to the sediments, the 202Hg2+

resulting from the demethylation was virtually zero at the beginning of the experiment.

Equation 3.9 reduces to :

d CH Hg

dtk CH Hgd

[ ][ ]

33

202

202

++= − ( 3.13 )

Integration of equation 3.13 leads to:

[ CH3202 Hg+ ] = [ CH3

202 Hg+ ]0 e - Kd t ( 3.14 )

Where [CH3202Hg+ ]0 represents the initial concentration of CH3

202 Hg+ in the sediments;

Kd is obtained by linear regression of ln [CH3202Hg+] versus time( t ). The demethylation

rate constant is often expressed as the half life of methylmercury in the sediment. The

two constants are related by t 1/2 = 1/Kd • ln2 . To calculate the specific methylation and

demethylation rate constants for ambient mercury species, equation 3.9 was written for

natural mercury species without specific isotopes and then integrated to yield :

27

[ ] [ ]( )CH HgK

KHg e

m

d

Kdt3

2 1+ + −= − ( 3.15 )

Specific rate constants for methylation and demethylation were calculated and compared

to rate constants obtained by monitoring changes in concentration of the ambient

methylmercury in the same sample. The inorganic tracer 199Hg2+ was methylated at a

faster rate as compared to the ambient Hg2+, indicating that the added tracer Hg2+ is more

available for transformation reaction than the ambient Hg 2+. The degradation of the

tracer and ambient methylmercury proceeded at a similar rate, showing no significant

differences between added tracer and ambient MeHg. The calculated half life for MeHg

in sediments was 1.7d, suggesting a rapid turnover and low persistence of MeHg in

sediments.

28

CHAPTER 4

Mercury Speciation

Although MeHg production is a function of the activity of methylating bacteria, it also

depends on the availability of Hg for methylation. MeHg may be formed from inorganic

ions, divalent Hg compounds, and organic Hg compounds as well as metallic Hg. To be

methylated by SRB, mercury must first be transported across the lipid membrane of the

microorganisms. Therefore, microbial uptake of mercury is a key step in mercury

methylation. At high concentration, Hg2+ is transported into SRB with the help of a

critical enzyme, methyltransferase; at low concentration, the uncharged mercuric

complex is relatively non polar and has better lipid solubility(Morel et al. 1998). Thus

the cellular uptake of mercury is chiefly affected by diffusion through the cell membrane

of lipid-soluble mercury complexes. Consequently, many research studies have been

conducted to study mercury speciation and lipid solubility of mercury. Hg 2+ ion exhibits

high affinity for sulfide. The speciation of dissolved Hg2+ in sulfidic environment is

primarily determined by sulfide and polysulfides. Five models regarding mercury

speciation are summarized in this chapter.

(Benoit et al.1999) hypothesized that the availability of mercury methylation in sediments

with sulfidic pore waters was controlled by the dissolved neutral Hg complexes such as

HgS0 rather than Hg2+ or total dissolved inorganic mercury (HgD). HgS0 is the dominant

neutral dissolved complex in sulfidic sediments. They developed a chemical equilibrium

model for mercury speciation in sulfidic pore water to predict observed HgD in sediments

29

from two different ecosystems, the Patuxent River Estuary and the Florida Everglades.

At first, they developed pure phase cinnabar solubility models (Table 4.1) for dissolved

Hg in the presence of excess cinnabar (HgS). The predicted HgD curve generated by this

model indicated that dissolved Hg increases with increasing sulfide concentration in the

presence of excess HgS(S).

In order to more adequately predict HgD trends in sediment pore water, the simple models

in Table 4.1 were modified by including adsorption to the solid phase (see Table 4.2). In

Table 4.2, reaction 4.6 is a simplified expression of the early diagnetic formation of

solids such as FeS or organic thiols. Reaction 4.9 and 4.10 are the overall reactions for

solid phase Hg complex formation. The constants for the overall reactions can not be

directly estimated from known constants due to the unknown formation constants for

RSH . The model was fit to the pore water data to derive K values for reaction 4.9 and

4.10. All of the dissolved Hg was present as sulfide complexes in this model. The

model included HgS0 in explaining Hg solubility and predicted that concentration of

HgS0 decreases with increasing sulfide, and concentration of HgD increases or keep

constant with increasing sulfide. The HgS0 concentration trends estimated by the model

consistent with the observed MeHg distributions in both the Patuxent River Estuary and

the Florida Everglades. This suggests that, HgS0 is the major neutral complex in sulfidic

sediments.

To help understand the mechanism and control of Hg uptake in mercury methylating

bacteria, (Benoit et al, 2001) investigated the effect of sulfide on Hg methylation by pure

30

cultures of the sulfate-reducing bacterium Desulfobulbus propionicus (1pr3). In this

study, the chemical speciation of Hg in culture media was manipulated by growing

Desulfobulbus propionicus across a range of sulfide concentrations. Inorganic Hg was

added in the form of ground ores. A solid-phase rather than a dissolved source of Hg was

used to simulate the controls on Hg partitioning between solid and aqueous phases found

in natural sediments. The results indicated that MeHg production by cultures was not

related to the absolute solid-phase Hg concentration in the ores but was related to the

dissolved inorganic Hg concentration in the medium. Methylation production, however,

was linearly related to the calculated concentration of the dominant neutral complex in

solution. The diffusive membrane permeability of HgS0 was found to be sufficient to

support MeHg production by cells. The experimental results also support the hypothesis

that sulfide influences methylation by affecting the speciation of dissolved inorganic Hg

and its uptake via passive diffusion.

The solubility of cinnabar (HgS(s)) and the mercury-sulfide speciation have been

extensively studied. The formation reactions along with the equilibrium constants are

given in Table 4.3 (Jay et al. 2000). A study performed by (Paquette and Helz 1997)

addressed the possible formation of complexes between mercury and polysulfides and

their influences on mercury solubility. They explained the solubility of cinnabar without

dissolved zero-valent sulfur (S0) (see Table 4.4). HgS solubility increased in the

presence of S0 . The additional Hg solubility in S0 saturated solutions with pH up to 9.5

can be explained by :

HgS ( cinn) + HS - + ( n-1) S0 (rhombic) = Hg (S n) ( SH )-

31

In sulfidic natural water, near neutral condition, Hg (S n) ( SH ) - usually exceeds other

inorganic mercury species due to the extensively existence of S0 , which is generated by

oxidation of sulfide.

(Jay et al. 2000) extended the study of (Paquette and Helz 1997) by studying the effect of

polysulfides on cinnabar solubility at lower S(-II )T ( ≤1 mM ), which represents the total

concentration of H2S, HS -, S2 -, S32-, S 4

2-, S5 2-, S6

2- , HS 4- and HS5

-. They proposed a

new chemical speciation scheme for mercury in the presence of polysulfides. At high

S(-II )T (≥1mM) , the data obtained were generally consistent with those of (Paquette

and Helz 1997). When the pH is low (≤8) , the dissolved mercury concentration in the

presence of S0 was 3-fold higher than in the absence of S0; the dissolved mercury

methylation concentration in the presence of S0 increased up to about 200- fold higher

than in the absence of S 0 with the increase of pH. The trend of increasing solubility

with pH was more obvious when S(-II )T is low ( ≤1 mM ). When the pH is high (≥8),

the data obtained by (Jay et al. 2000) significantly differed from those obtained by

(Paquette and Helz 1997). The observed solubility was 100 times larger than predicted

by(Paquette and Helz 1997). In order to improve the speciation model, (Jay et al. 2000)

provided new mercury speciation models in Table 4.5. They proposed that the complex,

HgSxS2-, dominated the mercury speciation in the water at high pH in the presence of

elemental sulfur. The model incorporating the formation of the species HgSxOH -

predicted the measured data reasonably good. (Benoit et al. 1999) conducted octanol-

water partition experiment at pH 8 in the absence of S0. The results of the octanol-water

distribution coefficient Dow (Dow = Dissolved Hg in octanol / Dissolved Hg in water)

32

confirm the charged nature of the dominant mercury-polysulfide complexes and imply

the presence of a uncharged species, HgS5. In addition, the octanol-water partition results

of this experiment support the model estimated data by (Benoit et al.1999) , they

hypothesized that the hydrophobic mercury sulfide species such as HgS0(aq) dominate in

aquatic sediments at low sulfide concentration.

33

Table 4.1. Mercury-sulfide Complexes and Equilibrium Constants ( K f ) in the

Speciation Models for Dissolved Hg in the presence of Excess Cinnabar

Complex log Kf

Hg2+ + 2 HS- = Hg (SH )20 37.5

Hg2+ + 2 HS- = HgS2H - + H+ 32.0

Hg2+ + 2 HS- = HgS22- + 2H + 23.5

Hg2+ + HS- = HgSH+ 30.5

Simulation 1

HgS (S) + H + = Hg2+ + HS - log KSP = -38,-37,-36

HgS(S) = HgS 0 log KS1 = - 10

Hg2+ + HS - = HgS0 + H+ log KS0 = 28, 27, 26

Simulation 2

HgS(S) + H + = Hg2+ + HS- log KSP = -37

HgS(S) = HgS0 log KS1 = - 11,-10, -9

Hg2+ + HS - = HgS0 + H+ log KS0 = 28, 27, 26

Where Kf represent equilibrium constants; KSP represents the solubility product of cinnabar nabar; KS0 represents the intrinsic solubility of cinnabar; KS0 = KS1/KSP. Simulations 1 and 2 show how the different values of KSP and KS1 affect HgD in equlibrium with excess cinnabar, respectively. Source from (Benoit et al.1999). .

34

Table 4. 2. Mercury-sulfide Complexes and Equilibrium Constants ( Kf ) Used in

the Speciation Models for Dissolved Hg with Sorption to the Solid Phase

Dissolved species log Kf Equation

Number

Hg2+ + 2HS- = Hg(SH)20 37.5 ( 4.1 )

Hg2+ + 2HS- = HgS2H - + H+ 32.0 ( 4.2 )

Hg2+ + 2HS- = HgS22- +2 H+ 23.5 ( 4.3 )

Hg2+ + HS- = HgSH+ 30.5 ( 4.4 )

Hg2+ + HS- = HgS0 + H+ 26.5 ( 4.5 )

Solid species Reaction Type

ROH + HS- = RSH + OH - Solid-phase thiol formation ( 4.6 )

RSH + Hg2+ = RSHg+ + H + Sorption to solid ( 4.7 )

2RSH + Hg2+ = (RS)2Hg +2H + Sorption to solid ( 4.8 )

Net Reaction log Kf

ROH + HS - + Hg2+ = RSHg+ + H2O Unknown ( 4.9 )

2ROH + 2HS- + Hg2+ = (RS)2Hg + 2

H2O

Unknown ( 4.10 )

Where ROH may represent inorganic precipitates or organic particles. Source from (Benoit et al.1999).

35

Table 4.3. Reactions and Constants for Mercury-Sulfide Interactions

Formation reactions log Kf

HgS (cinn) + HS- = HgS22- + H+ - 13.0

HgS (cinn) + HS- = HgS2H - - 4.5

HgS (cinn) + HS- + H + = Hg(SH)2 + 1.0

HgS (cinn) + H + = HgSH + -16.81

HgS (cinn) = Hg2 + + S2 - - 53.5

HgS (cinn) = HgS( aqueous) - 9.3

Source from (Jay et al. 2000).

36

Table 4. 4. Reactions used to explain the cinnabar solubility.

Reaction log Kf (298 k, Ionic strength = 0.7M )

HgS ( cinn) + H2S (aq) = Hg(SH)20 - 5.36 ± 0.1

HgS ( cinn) + HS - = HgS(SH) - - 5.34 ± 0.30

HgS ( cinn) + 2HS - = HgS22- + H2S -7.14 ± 0.16

HgS ( cinn) + 2 H2S (aq ) = Hg ( H2S )(SH)2 - 3.43

HgS ( cinn) + HS - + H2S = Hg(SH)3- - 2.27

Source from (Paquette and Helz 1997).

37

Table 4. 5. Formation Reactions and Constants for Hg -Sx Species

Reactions log Kf

HgS ( cinn) + HS - + (x-1) S 0 = HgSxHS - - 3.8

HgS ( cinn) + (x-1)S 0 = HgSx0 -5.9

HgS ( cinn) + HS - + (x-1) S0 = HgSxS2- + H+ -11.7

HgS ( cinn) + HS - + 2(X-1) S0 = Hg(Sx)22- + H+ -11.7

HgS ( cinn) + (x-1) S0 + H2O = HgSxOH - + H + - 15.7

The model proposed here is shown in the last two rows. Source from (Jay et al. 2000).

38

CHAPTER 5

Factors Affecting Mercury Methylation

The efficiency of microbial mercury methylation in aquatic sediments primarily depends

on the microbial activity and the bioavailability of mercury, which are influenced by a

wide variety of environmental factors such as organic matter, temperature and pH as well

as salinity and concentration of sulfide. In this chapter, the impact of these factors on

mercury methylation are presented.

5.1. Organic matter

The role of organic matter in the methylation of Hg is not well understood. Increasing

MeHg concentrations were observed in water, sediments with increasing levels of organic

carbon. A stimulating effect of organic nutrients on microbial methylation activity may

be the explanation contributed to this phenomenon (Ullrich et al. 2001). On the other

hand, (Barkay et al. 1997) demonstrated that dissolved organic matter decreases the rate

of MeHg formation by reducing the availability of the Hg 2+ to methylating bacteria.

However, the exact interaction between organic matter and mercury remains unclear.

5.2. Complexation

Dissolved organic matter (DOM) is believed to form strong complexes with MeHg.

(Amirbahman et al. 2002) used an equlibrium dialysis technique to study the extent of

association of MeHg with humic acids. They modeled the experimental data at different

39

MeHg and humic concentrations and at various pH values. The reactive thiol functional

groups of humic acids were modeled as multisite acids

H + + RS(i) - = RS(i) H; Ka (i) ( 5.1 )

Where RS(i) - is the deprotonated form of the ith thiol function group of humic acids;

RS(i) H represent the protonated form of the ith thiol function group of humic acids; Ka(i)

is the corresponding acidity constant. The reaction between MeHg and the humic

functional group was written as

CH3Hg+ + RS(i) - = RS(i ) Hg CH3 ; Kf (i ) ( 5.2 )

Where Kf (i) is the equlibrium formation constant. Table 5.1 lists all the relevant

reactions. Three humic acid samples were used in this experiment, which is Suwannee

River humic acid (SRHA), peat humic acid (PHA) and the humic acid from Baker Brook

(BBHA). The experiments were performed in dialysis cells with two chambers separated

by a dialysis membrane. A known concentration of MeHg was added to one cell and a

known concentration of the humic acid was added to the other cell. When the sorption of

MeHg to the humic acids reached equilibrium, concentration of MeHg was measured in

both cells. Table 5.2 lists the equilibrium binding constants and binding capacities for

the adsorption of MeHg with humic acids. The values of the equilibrium binding

constants listed in Table 5.2 were similar to those of MeHg with thiol-containing

compounds, suggesting MeHg associates primarily with the thiol group in humic acids.

40

The adsorption between MeHg and all humic acids exhibited pH dependence. At pH

values from 5.2-9.2, adsorption was relatively constant. The extent of adsorption

decreased when the pH value was below 5.2 due to the competition of MeHg with H+

for binding to the thiol groups.

Competitive complexation of inorganic Hg2 + with inorganic and organic ligands as well

as colloids is considered to be one of the principal factors controlling the bioavailability

of inorganic Hg2 + in aquatic ecosystems. (Farrel et al. 1998) assessed the effect of

mineral colloids common in freshwater sediments on the biomethylation of Hg2 + in a

synthetic growth medium (M-IIY), which was a minimal salts medium amended with

0.1% yeast extract and 0.1% glycerol and was made chloride free by substituting

appropriate nitrate salts. Three types of mineral colloids were used in the study, which is

kaolinite, montmorillonite and brinessite, respectively. The addition of kaolinite or

montmorillonite to the medium containing mercuric nitrate [Hg(NO3)2; 12µg Hg/ml]

have no significant effect on the production of methylmercury. The addition of

brinessite results in a significant decrease in the production of MeHg. It was found that

the adsorption of Hg2 + onto montmorillonite and brinessite before they were added to

the medium decreased the bioavailability of Hg2+. The amount of MeHg produced from

the mineral colloid added medium were significantly lower than those in the medium

without mineral colloids. MeHg production was decreased by 21% in the case of

montmorillonite. Moreover, in the case of brinessite, no MeHg was detected in the

medium after 25 hours incubation. The experimental results suggested that bioavailability

and methylation of mercury in aqueous system is affected by the mineral colloids.

41

Brinessite(MnO2 ) is an effective inhibitor of mercury methylation due to its ability of

adsorbing large amount of Hg2 +.

5.3. Temperature

It has been observed that peak Hg methylation rates frequently appear during the summer

months (Hintelmann et al. 1995). Seasonal variations in MeHg production generally

have been found related to temperature effects. Temperature affects methylation most

likely as a result of its effect on the overall microbial activity (Bisogni and Lawrence

1975). It was reported that microbial methylation in surficial river sediments had an

optimal temperature of 35 °C (Callister and Winfrey 1986). At the roots of tropic

floating macrophytes, where high methylation was observed, the highest methylation was

found in the temperature range of 35 ~ 45°C . Above 55°C methylation was completely

inhibited. At such temperatures, many enzymes are inactivated and bacterial activity is

probably stopped, suggesting a biological control of mercury methylation

5.4. pH

The effect of acidification on mercury methylation in sediment was examined by

(Steffan et al. 1988). The sediment samples were obtained from northern Wisconsin lake.

Mercury methylation was inhibited by adding H2SO4, HCl or HNO3 into the samples.

Decreasing sediment pH value from 6.1 to 4.5 with H2SO4 or HCl inhibited methylation

by 65%. The acidification of surficial lake sediments resulted in a significant decrease in

Hg methylation rates. Aerobic methyaltion in surface sediments was also found to

decrease with decreasing water pH (Matilainen et al. 1991). Methylation at the roots of

42

macrophytes was stimulated at pH values between 6~7; however, a significant

methylation decrease was verified at pH 8 (Mauro et al. 1999)

5.5. Sulfide and sulfate

Hydrogen sulfide plays a very important role in the chemistry of anaerobic sediments

where it is produced as a result of bacterial sulfate reduction. High sulfide concentrations

appear to inhibit MeHg formation in soil, sediments and bacteria culture (Jacobs and

Keeney 1974). In order to investigate the relationship between bacterial sulfate reduction

and mercury methylation, (Gilmour et al. 1992) conducted an experiment in Quabbin

Reservoir, MA. The results suggested that there is an optimal sulfate concentration ( 0.2

~0.5 mmol/L) for mercury methylation by sulfate-reducing bacteria in sediments.

Production of sulfide would inhibit methylation above this optimal sulfate concentration,

while sulfate availability would limit microbial sulfate reduction and hence mercury

methylation below the maximum. However, SRB like Desulfovibrio desulfuricans LS

can produce MeHg under fermentative condition without the presence of sulfate as

electron acceptor.

5.6. Salinity

The methylating activity of marine and estuarine sediments is usually lower than that of

freshwater sediments. It is generally attributed to salinity effects. (Compeau and Bartha

1987) measured mercury methylation in anoxic estuarine sediments in which salinity

ranged from 0.03 to 2.4 % (salinity was measured with a salinometer). A strong inverse

relationship between the salinity of anaerobic sediments and their methylation ability

43

was observed. High salinity sediments methylated Hg at only 40% of the level observed

in low salinity sediments. Under reducing conditions, the effect of inhibition is

particularly pronounced.

44

Table 5.1. Reactions and Constants for MeHg and Ligands

Reaction log Kf

H+ + OH- = H2O 13.97

CH3Hg+ + H2O = CH3HgOH + H+ - 4.63

H+ + Ac- = HAc 4.73

CH3Hg+ + Ac- = CH3HgAc 2.95

H+ + HPO42- = H2PO4

2- 7.14

CH3Hg+ + HPO42- = CH3HgHPO4

- 5.41

H+ + RS(1) - = RS(1) H 4.0

H+ + RS(2) - = RS(2)H 7.0

H+ + RS(3)- = RS(3)H 10.0

CH3Hg+ + RS(i) - = RS(i)HgCH3 Log Ks

Where Ac is abbreviation for acetate ion used in experiments at pH 5.2 only. LogKs represent reaction constants for adsorption of MeHg to humic acids used in this experiment (See Table 5.2 ). Source from (Amirbahman et al. 2002).

45

Table 5. 2. Equilibrium Binding Constants and Binding Capacities for Formation

of MeHg Complexes with Humic acids

Humic

acids

log Ks (1)

*RS T (1)

nmol/mg

log Ks (2)

*RS T (2)

nmol/mg

log Ks (3) *RS T (3)

nmol/mg

SRHA 10.39 0.15 14.74 0.24 14.84 1.44

PHA 10.42 0.43 12.39 0.13 14.47 1.51

BBHA 10.54 0.25 14.77 0.10 14.96 0.76

*RST(i) represents the binding capacity of humic acids. Source from (Amirbahman et al. 2002).

46

CHAPTER 6

Inhibition of Mercury methyaltion

Mercury methylation is linked to microbial sulfate (SO42-) reduction and influenced by

the concentration of ambient sulfate. Some substances can also inhibit microbial sulfate

reduction. In this chapter, the factors affecting sulfate (SO42-) reduction and eventually

mercury methylation are presented.

Many group VI anions, MoO42- (molybdate), WO4

2- (tungstate), TeO42–( tellurate) and

SeO42- (selenate), are capable of inhibiting microbial sulfate reduction. They can pass

through cell membranes along the same pathways as SO4 2- because they resemble SO4

2-

in terms of size, charge and stereochemistry (Chen et al. 1997). The concentrations of

above anions in natural aquatic systems are many orders of magnitude lower than SO42-,

their influence with sulfate reduction is thus not obvious. Relatively high concentrations

of these anions ( 0.1 mM MoO4 2- and 0.7 mM WO4

2 - ) have been reported in terminal

lakes of rivers draining from the east front of the Serria Nevada. Although the

concentrations of total Hg (Hg (T)) is high, concentrations of MeHg are relatively low.

(Chen et al. 1997) conducted a series of experiments in order to identify the effect of

group VI anions on rates of Hg methyaltion. In the experiments, surface sediment

samples were obtained from a sampling site in the Carson River-Lahontan Reservoir

system, Neveda, U.S.A. The results of the experiments indicate that mercury methyaltion

was significantly inhibited ( > 5% ) by SeO42 - and TeO4

2 - at the nanomolar level ( > 50

nM of TeO42 -, > 270 nM of SeO4

2 - ); while methylation was inhibited by MoO42- and

47

WO42 - at the micromolar level ( ≥100 µM of MoO4

2 -, ≥700 µM of WO42 - ). MoO4

2 -

is unlikely to affect MeHg production in the sediments collected from the Carson River,

because its concentration is below 100 µM. Although WO42 - is less inhibitory than

MoO42 - to MeHg formation, the concentration of WO4

2- (700 µM) is high, thus it can

inhibit MeHg formation in the Carson River significantly. In the experiments, the rates

of MeHg production slightly increased when SeO42 - was added at concentrations below

10 nM, whereas SeO42 -significantly inhibited MeHg formation when concentrations

were higher than 270 nM. This difference might be attributed to the complex aqueous

geochemistry of selenium. Selenium may be present in natural waters in more than one

oxidation state (- 2, 0, + 4 and + 6 ). The effect of selenium to mercury methylation thus

is complicated. Se6+ and Se4+ in anoxic sediment may be reduced to elemental

selenium(Se0) and ultimately to selenide(Se2-) by anaerobic bacteria. High concentrations

of reduced Se may decrease the availability of Hg2 + by forming insoluble HgSe. In

these experiments, the mercury methylation rate decreased 34 % when 1 µM SeO42 - was

added. While the rate decreased 94 % when 20 mM SeO42 - was spiked. These results

suggest that high concentration of SeO4 2 - would inhibit MeHg production by forming

insoluble HgSe. The inhibitory effect of TeO42 - was more pronounced than that of

SeO42 - within the concentration ranges in the test, but the inhibitory trends of TeO4

2 - and

SeO42 - are similar. TeO4

2 - concentrations in most aquatic systems are less than 0.6 nM.

Therefore, a significantly inhibition of TeO42 - on MeHg production is unlikely.

(Chen et al. 1997) also evaluated the effect of various sulfate concentration on the

inhibition of mercury methylation by group VI anions. The addition of high

48

concentration of sulfate (1,200µM ) reduced the WO42 -, MoO4

2- and TeO42 - inhibition on

Hg methylation, whereas increased SeO42- inhibition. Effect of various sulfate

concentrations on the methylation rate inhibited by group VI anions are listed in Table

6.1. When the concentration of SO42 - was increased from 76 to 1200 µM , the percent

inhibition of methylation rate by WO42- significantly decreased from 20.2% to 5.62% ,

suggesting WO42- inhibited mercury methylation by competing with SO4

2-. While the

percent inhibition of mercury methylation rates by TeO42 - and MoO4

2- decreased from

32.7% to 29.1% and 56.0% to 47.4%, respectively, suggesting MoO42- and TeO4

2 -

inhibited mercury methylation through a noncompetitive mechanism other than

competing with SO42 -. However, the increase of SO4

2 - concentrations stimulated the

inhibition of Hg methylation, suggesting that SeO42 - play the similar role as SO4

2 - .

In the experiments aimed to investigate the effects of acidification on mercury

methylation, (Steffan et al. 1988) observed that decreasing pH of sediments to 5.5, 4.5

and 3.5 with HNO3 or equal amount NaNO3 resulted in almost complete inhibition of

mercury methylation, indicating NO3- may inhibit mercury methylation.

49

Table 6.1. Effect of various sulfate concentration on the methylation rate inhibited

by group VI anions

Percent inhibition of Hg methylation rate by addition of group VI

anions SO4

2 – ( µM )

TeO42 –

( 1 µM ) SeO4

2 –

( 1 µM) MoO4

2 – ( 2,000 µM)

WO42 –

( 2,000 µM) 76 32.7 32.3 56.0 20.0

1,200 29.1 40.7 47.4 5.62

Ratio of inhibition 76 µM/1,200µM SO4

2 -

1.12 0.79 1.18 3.59

Source from (Chen et al. 1997).

50

CHAPTER 7

Specific Site Study

Methylmercury is widely distributed in the environment, the researches regarding

mercury methylation are conducted worldwide. Five studies on specific site are

summarized in this chapter.

7.1. Seine River (France)

The hydrographic basin of the Seine River includes 30 % of the French population and 30

~ 40% of the total economic activity. As a consequence of industrial and agricultural

development of industry and agriculture, Seine estuary is highly contaminated by

mercury. (Mikac et al. 1999) conducted a study to assess the distribution and behavior of

total mercury and MeHg within Seine estuary. In this work, sediment cores were

collected on several occasions in the period 1994 -1997 at marine (M), estuarine (E) and

riverine (R) locations. The level of total Hg was 380 ± 80 ng/g at location M, 410 ± 95

ng/g at location E, and 406 ± 98 ng/g at location R. These results suggested

concentrations of total Hg were uniform all over the estuary and did not show significant

spatial variation. Levels of MeHg, however, displayed a spatial variation. The

concentration of MeHg was 1.3 ± 0.2 ng/g at location M, 3.1 ± 1.2 ng/g at location E,

and 2.3 ± 0.6 ng/g at location R. Vertical distribution of MeHg at location M and

location R did not show a significant change with depth up to 25cm. Maximum MeHg

concentration were obtained just below the sediment and water interface at estuarine

location E and decreased with depth. The vertical distribution of MeHg in sediment was

51

shown to be dependant on the sulfate reduction rate (SRR). The sediment depth profiles

of bacterial sulfate reduction and Hg methylation rate were similar. No correlation

between MeHg and total Hg was observed among all samples, indicating factors such as

sulfate reduction, redox potential and salinity other than total Hg concentration affect the

MeHg level in sediments of the Seine estuary. It was postulated that there is an optimal

sulfide concentration (0.2~ 0.5 mmol/L), above which sulfide would inhibit methylation;

whereas at lower sulfate levels, the sulfate reduction and methylation would be limited

by available sulfate. Maximum MeHg production could exist at a redox potential of 0 ~

-100 mV, where the sediment is anaerobic but not too high in sulfide. MeHg

concentration decreased with increasing salinity if sulfide levels reached a critical level

(2 ~ 6 mg/g , depending on the area).

7.2. Carson River (Nevada)

The Carson River system in Nevada has been contaminated by mercury for a long period

of time. Concentrations of MeHg in surficial sediments along the Carson River and

factors controlling MeHg production in sediment were investigated by (Chen et al.

1996). In this study, the samples were collected in June 1994, January and July 1995.

Biotic and abiotic activity of sediment samples, defined as the potential of each

compartment to specifically reduce an alternative electron acceptor, were used to assess

the influence of biological and nonbiological process on mercury methylation in

sediments. The method for determining abiotic activity was a spectrophotometric test

based on the quantitative reduction of the dye resazurin by both chemically reducing

substances and dehydrogenase in microorganisms. Microbial activity was inhibited by

52

using m-cresol. The abiotic activity was expressed as µg of resazurin reduced per day

and per gram of sediment. Biotic activity was determined based on the measurement of

the electron transport system (ETS) activity of respiring microbes. In this method, the 2-

( p-nitrophenyl )-3-( p-nitrophenyl )-5-phenyl tetrazolium chloride(INT) is reduced to

iodonitrotetrazolium(INT- formazan). The biotic activity was expressed as µg of INT-

formazan produced per day and per gram of sediment. Concentrations of MeHg varied

from less than 2-28.5 ng Hg/g dry sediment, representing less than 3% of the total Hg

concentrations. The concentrations of MeHg were one order of magnitude higher than

those in uncontaminated sediments. Concentrations of MeHg were related to both biotic

activity (R2 = 0.95) and abiotic activity (R2 = 0.85) of the sediments. The biotic activity

was positively correlated to abiotic activity, suggesting that the abiotic activity was

probably linked to reductant substances produced by microorganisms. In order to assess

the effect of periodic inputs of inorganic Hg on the potential of MeHg formation and

microbial activity, increasing concentrations of HgCl2 were spiked into sediment

samples. The addition of inorganic Hg (HgCl2 ) in concentrations less than or equal to

15.3 µg / g dry weight resulted in an increase of methylation rate. The methylation rate

decreased when the spike concentration of inorganic Hg was above 15.3 µg / g dry

weight. These results suggested that seasonal inputs into the river of significant amounts

of inorganic Hg eroded from mill tailings could have an inhibiting effect on Hg

methylating microorganisms.

53

7.3. Everglades sediments (Florida)

MeHg concentrations and production rates were examined by (Gilmour et al. 1998) in

Everglades sediments in March, July and December, 1995, aiming to determine the

importance of Hg methylation in controlling MeHg levels and evaluate the effect of

factors such as eutrophication, sulfate concentration and temperature on mercury

methylation. Samples were collected from the Everglade agricultural runoff, across the

Everglades Nutrient Removal (ENR) area and Water Conservation Area (WCA) 2A 2B

and 3. The sampling sites generated a roughly north to south nutrient gradient. MeHg

levels and % MeHg were lowest in the more eutrophic areas and increased dramatically

to the south. The highest MeHg concentrations were less than 0.1 ng/g dw in sediments

in the ENR area and around 5 ng/g dw in WCA3 sediments; MeHg constituted less than

0.2% of total Hg in ENR area, but up to about 2% in two sites in WCA2B and WCA3.

By using tracer-level injections of 203Hg(II) into sediment samples, mercury methylation

rates in surficial sediments (0 ~ 4 cm) were estimated to be in the range of 1 ~ 10 ng/g

day. Methylation rates generally increased from north to the south. The distributions of

MeHg and its production suggested that MeHg concentrations in sediments are

controlled by in situ methylation. Methylation rates were lowest among sites sampled in

December (average temprature is 18°C) and higher rates of mercury methylation were

measured in summer (average temprature is 28°C), suggesting lower temperatures may

inhibit microbioal activity and methylation. Sulfate concentrations in surficial pore

waters (up to 400µm), and microbial sulfide concentrations (up to 300µm) at the

eutrophic northern sites were all higher than those in southern sites. MeHg concentration

and production were inversely related to pore water sulfide and sulfate reduction rate.

54

7.4. Pine Barrens area (New Jersey)

Aiming to investigate the effect of environmental factors such as sulfate and sulfide

concentrations, pH, and organic matter on the mercury methylation, (Pak and Bartha

1998b) conducted experiments in sediments of three oligotrophic lakes in the Pine

Barrens area of southern New Jersey. The lake sediment characteristics are listed in Table

7.1. The samples collected from above lakes were spiked with 1.0 µg/ml of HgCl2 and

incubated for 15 days. The results indicated 15 ~ 22 ng/ml of MeHg was produced.

Initial formation of MeHg was rapid and approached equilibrium after 5 ~ 10 days. At

time zero, only the Atlantic City Reservoir sediment contained a detectable amount of

MeHg (4 µg/ml ). When the same sediments were spiked with 0.1 µg/ml of MeHg,

MeHg concentration decreased rapidly and most of the decrease occured in the first 5

days of incubation. The high mercury methylation and demethylation activities in

Atlantic City Reservoir sediments correlate positively with the level of organic matter

and the sulfate concentration in the pore water (Table 7.1 ). Large mouth bass

(Micropterus salmoids) from Atlantic City Reservoir were found to contain higher

concentrations of MeHg ( 3.0 ~ 8.9 µg/g) than those in Batsto Lake ( 0.7~1.3 µg/g) and in

East Creek Lake ( 0.8 ~ 2.8 µg/g ). All of the mercury found in fish is present in the form

of MeHg, so mercury methylation rates may be correlated with elevated methylmercury

levels in fish. At low sulfate levels in freshwater, addition of sulfate to the 200 mM level

was found to stimulate Hg2+ methylation but high sulfate concentrations of estuarine

sediment were found to correlate inversely with mercury methylation activity due to the

reaction of H2S with Hg2+ to form HgS, which is unavailable for methylation. The

sulfate concentrations of sediments in the Pine Barren Lakes were only 44 ~ 67 µg/ml of

55

pore water, thus mercury methylation activities positively correlate to the sulfate

concentrations. The pH values at sediments were adjusted to 7.0. Neither methylation

nor demethylation rates were affected by this moderate adjustment. Pure cultures of

sulfidogenic, methanogenic and acetogenic bacteria, Desulfovibrio desulfuricans LS,

Methanococcus maripaludis, and Eubacterium limosum, were incubated. Then activities

of mercury methylation and demethylation were measured. The results suggested that

Desulfovibrio desulfuricans LS both methylated and demethylated mercury, but

Methanococucs maripaludis only catalyzed demethylation and Eubacterium limosum

neither methylated nor demethylated mercury.

7.5. Pantanal floodplain (Brazil)

Potential 203Hg methylation was assayed by (Guimarâes et al. 1998) in the samples of

different substrates such as surface sediment, roots of floating macrophytes in Frazenda

Ipiranga Lake, 30km downstream the Poconé gold mining fields in the Pantanal

floodplain, Brazil. The influence of temperature, salinity on mercury methylation in

sediments was also studied. Samples of surface sediments and roots of dominant floating

macrophytes such as Eichhornia azurea, Salvinia sp were spiked with approximate 43 ng

Hg/g dry weight and incubated in situ for 3 days. Then Me 203Hg was extracted in

toluene and measured by beta counting. Net methylation was 0.4 ~ 1.2% in all sediment

samples. The methylation percentages were slightly lower in surface sediments at an

open lake site than those from the littoral site under the floating macrophytes. Some

sediment samples were treated with 0.2 ml of sulfate(Na2SO4) or molybdate (Na2MoO4 )

solutions and incubated for 3 days. Mercury methylation was stimulated with the

56

addition of sulfate and inhibited by molybdate. Sulfate and molybdate had effects on

methylation in samples from both open lake sites and littoral sites but more marked in the

littoral sites, suggesting that sulfate reducing bacteria may be important Hg methylators at

both sites and their activity is sulfate-limited in particular at the littoral sites. The highest

methylation was found in the temperature range of 35 ~ 45°C. Above 55°C ,

methylation was completely inhibited. At such temperatures, many enzymes are

inactivated and bacterial activity is probably stopped, suggesting a biological control of

mercury methylation. The effect of salinity on methylation was investigated by adding

NaCl solutions into sediment samples. The conductivity values of sediment samples was

116, 248 and 314 µm/cm, respectively. The highest MeHg production appeared at the

lowest conductivity (116 µm/cm). Me 203Hg was detected only in the upper layer ( 0 ~ 2

cm) of the sediments. Total 203Hg was detectable down to the 14 ~ 16 cm layer and high

concentrations of total 203Hg were found in the top 4 cm of the sediment, corresponding

to the depth reached by the swimming insects present in the sediment. It suggested

swimming insects caused 203Hg penetration down to 4 cm. The MeHg production in the

roots of two dominant floating macrophytes, E.azurea and Salvinia sp, were 5.6 times

and 9 times, respectively, higher than in the surface sediments. An average of 10.4% of

added Hg was methylated in roots of Salvinia sp and 6.5% in roots of E.azurea. These

results indicated the tropical aquatic floating macrophytes are important sites for mercury

methylation.

57

Table 7.1. Physiochemical characteristics of three Pine Barrens lakes

Lake PH Concentration

Orgnic matter ( %)

Sulfate ( µg/ml of pore water)

Sulfide ( µg/g )

Atlantic City Reservoir

6 31 67 86

Batsto Lake

6 24 30 69

East Creek Lake

5.5 25.8 44 75

Source from (Pak and Bartha 1998b)

58

CHAPTER 8

Methylmercury Demethylation

Both methylation and demethylation processes occur in the aquatic sediments, the

environmental methylmercury values measured reflect the equilibrium concentrations

between these two processes. In contrast to mercury methylation, much more is known

about the mechanisms for MeHg demethylation. Photolytic demethylation appears to be

the only significant abiotic mechanisms in water column(Sellers 1996). In sediments,

microbial demethylation is predominant.

The cleavage of MeHg via organomercurial-lyase (OML) is one of the pathways for

microbial MeHg demethylation. The merB gene encodes for the formation of OML, the

later cleaves MeHg into methane and Hg (II). The second mechanism for MeHg

degradation is oxidative demethylation (OD) pathway.

(Marvin-Dispasquale and Oremland 1998) conducted a research to evaluate MeHg

degradation along an eutrophication gradient in the Florida Everglades. [14C] MeHg

degradation via oxidative demethylation (OD) was observed at all sites in the Florida

Everglades during each sampling period indicating the OD was an important mechanism

of mercury degradation. Both methogens and sulfate reducers are involved in OD, they

can oxidize methyl group to CO2, either with or without CH4 production. The pathways

of OD by these two microbial groups were proposed as follows:

59

4 CH3Hg+ + 2H2O + 4H + methanogens 3CH4 + CO2 + 4Hg2+ + 4H2

SO42- + CH3Hg+ + 3H + sulfate reducers H2S + CO2 + Hg2+ + 2 H2O

60

CHAPTER 9

Bioavailability of Methylmercury

Many studies have been performed to study mercury methylation and demethylation

dynamics, few studies focus on the bioavailability of MeHg. (Nuutinen and Kukkonen

1998) conducted a laboratory study aiming to 1) measure the accumulation kinetics and

the body burden of added MeHg in an oligochate worm Lumbriculus variegatus in

different lake sediments and 2) evaluate the effect of selenium concentrations on the

bioavailability of MeHg to L. variegatus.

The sediments were collected from Lake Höytiäinen and Mekrijärvi, respectively. 14C -

labeled MeHg was added to the sediments at the nomial concentration of 95 ng/g dry

sediment. Groups of six oligochate worms were raised in glass beakers for two weeks.

The kinetics of MeHg accumulation were determined by fitting the data to a first-order

rate-constant model

CK C e

Ka

u sKet

e=

− −* *( )1

Where Ca is the MeHg concentration in the organism ( ng/g wet weight organisms ), Ku

is the uptake clearance coefficient (g dry sediment/g wet organism.h), Cs is the

concentration of MeHg in the sediment ( ng/g dw ), t is the time (h), Ke is the elimination

rate constant (0.0005/h). After two weeks exposure, the uptake rate constant Ku was

61

0.0089 ± 0.0004 and 0.0032 ± 0.0002 in Lake Höytiäinen and Lake Mekrijärvi sediment,

respectively. The organic carbon concentration was 3.4% in Lake Höytiäinen and around

10% in Lake Mekrijärvi. It suggested that the accumulation rate of MeHg in the worms

were much lower in sediments having a high organic carbon content. Different

concentrations of sodium selenite (Na2SeO3) were spiked into Lake Höytiäinen sediment,

and the effect of selenium concentration(0.1 ~ 50 mg/kg dry sediment ) on bioavailability

of MeHg by L.variegatus was measured. The two lowest selenium concentrations (0.1,

0.5 mg/kg dry sediment) did not affect the bioaccumulation of MeHg. 2.5 mg Se/Kg dw

resulted in a 25% reduction in the body residue after two weeks exposure. When the

concentrations of Na2SeO3 was 15 and 50 mg Se/Kg dry sediment, the accumulation of

MeHg in the organisms was decreased by 75% and 86%, respectively.

62

CHAPTER 10

Conclusions and Recommendation

10.1. Conclusions

Through literature review, the principal methylators of mercury methylation are

discussed. The biological pathways and reaction equations regarding mercury

methylation were presented. Mercury methylation rate along with affecting factors were

described. Environmental parameters controlling mercury methylation were studied.

The following conclusions can be made:

• Mercury methylation is primarily a microbially mediated process which is

catalyzed by various enzymes.

• Sulfate reducing bacteria (SRB) are identified as the principal methylators of

mercury.

• A corrinoid protein, cobalamin (vitamin B12), is the methyl carrier in SRB.

Methyl group is transfered to cabalamin, forming methylcobalamin. MeHg is

then formed through the reaction between methylcobalamin and Hg 2+.

• Different SRB species methylate mercury at various rates.

• MMR is tightly correlated with the sulfate reduction rate and SRB species.

• Mercury methylation in aquatic sediments primarily depends on the microbial

activity and the biloavailability of mercury, which are influenced by a wide

variety of environmental factors such as temperature, pH, salinity, sulfide

concentration and the concentration of organic matter.

63

• Chemical speciation of mercury determines the availability of mercury for

methylation. Uncharged mercuric complexes such as HgS0 are thought the

dominant species taken by microbial bacteria.

• Maximum MeHg concentrations were obtained at the sediment and water

interface.

• Temperature range 35 ~ 45 °C is the optimum for methylmercury formation.

Above 55°C, methylation was completely inhibited.

• The acidification of surficial lake sediments resulted in a significant decrease in

Hg methylation rates.

• There is a strong inverse relationship between the salinity of anaerobic sediments

and their methylation ability.

• Mercury methylation can be inhibited by NO3- and group VI anions such as

MoO42-, WO4

2-, TeO42 - and SeO4

2 -. Moreover, methylmercury formation could

be blocked by propyl iodine.

10.2. Recommendation for future work

The study of methylation pathway was based on SRB such as Desulfovibrio desulfuricans

LS. High degree of MeHg can be formed in D. desulfuricans LS under fermentative

condition rather than under sulfate reduction condition, especially when pyruvate was

added. Substrate like acetate was observed to suppress mercury methylation. However,

SRB such as Desulfobacterium can not form MeHg without the presence of sulfate. SRB

groups capable of utilizing acetate, like the family Desulfobacterionaceae, methylate

mercury faster than those not capable of utilizing acetate, like the family

64

Desulfovibrionaceae. Therefore, different types of SRB not necessarily methylate

mercury along the same pathway. More extensive research regarding methylation

pathways in microorganisms need to be conducted.

MeHg formation includes the following key steps : 1) Methyl group is transferred to

cobalamin, forming methylcobalamin. 2) Methylcobalamin reacts with Hg2+, forming

MeHg. The rate limiting steps for mercury methylation is not yet clearly documented in

literature.

The accuracy of the model describing the relationship between mercury methylation rate

(MMR) and sulfate reduction rate (SRR) developed by (King et al. 1999) decreased with

increasing temperature. At 37°C, the predicated MMR was only 45% of the observed

MMR. The addition of pyruvate and acetate also decreased the accuracy of the estimated

MMR. Further development of the model by considering environmental factors such as

temperature and substrate is required.

Application of the model developed by (King et al. 2001) has some limitations. First, the

model was based on pure culture hence its use to predict methylation rate is limited.

Second, the mercury concentrations used in the model are the concentrations of added

mercury rather than contaminated sediments. Considering these differences in the model

is necessary.

65

Methymercury concentrations in sediments are controlled by both the methylation and

demethylation process. The net methylmercury levels are determined by the relative

importance of each process. Therefore, further work using simultaneous assay is

necessary to evaluate the two opposite processes.

66

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