NOAA Technical Memorandum NOS ORCA 100

 

Contaminants in Aquatic Habitats at Hazardous Waste Sites: Mercury

December 1996

Seattle, Washington

noaa NATIONAL OCEANIC AND ATMOSPHERIC ADMINISTRATION

National Ocean Service
Office of Ocean Resources Conservation and Assessment

National Ocean Service

National Oceanic and Atmospheric Administration

U.S. Department of Commerce

The Office of Ocean Resources Conservation and Assessment (ORCA) provides

decisionmakers comprehensive, scientific information on characteristics of the

oceans, coastal areas, and estuaries of the United States of America. The

information ranges from strategic, national assessments of coastal and estuarine

environmental quality to real-time information for navigation or hazardous materials

spill response. Through its National Status and Trends (NS&T) Program, ORCA

uses uniform techniques to monitor toxic chemical contamination of bottom-feeding

fish, mussels and oysters, and sediments at about 300 locations throughout the

United States. A related NS&T Program of directed research examines the

relationships between contaminant exposure and indicators of biological responses

in fish and shellfish.

Through the Hazardous Materials Response and Assessment Division (HAZMAT)

Scientific Support Coordination program, ORCA provides critical scientific support

to the U.S. Coast Guard for planning and responding to spills of oil or hazardous

materials into marine or estuarine environments. Technical guidance includes spill

trajectory predictions, chemical hazard analyses, and assessments of the sensitivity

of marine and estuarine environments to spills. To fulfill the responsibilities of the

Secretary of Commerce as a trustee for living marine resources, HAZMAT’s

Coastal Resource Coordination program provides technical support to the U.S.

Environmental Protection Agency during all phases of the remedial process to

protect the environment and restore natural resources at hundreds of waste sites

each year. As another part of its marine trusteeship responsibilities, ORCA

conducts comprehensive assessments of damages to coastal and marine resources

from discharges of oil and hazardous materials.

ORCA collects, synthesizes, and distributes information on the use of the coastal

and oceanic resources of the United States to identify compatibilities and conflicts

and to determine research needs and priorities. It conducts comprehensive,

strategic assessments of multiple resource uses in coastal, estuarine, and oceanic

areas for decisionmaking by NOAA, other Federal agencies, state agencies,

Congress, industry, and public interest groups. It publishes a series of thematic data

atlases on major regions of the U.S. Exclusive Economic Zone and on selected

characteristics of major U.S. estuaries.

ORCA implements NOAA responsibilities under Title II of the Marine Protection,

Research, and Sanctuaries Act of 1972; Section 6 of the National Ocean Pollution

Planning Act of 1978; the Oil Pollution Act of 1990; the National Coastal Monitoring

Act of 1992; and other Federal laws. It has four major line organizations: Coastal

Monitoring and Bioeffects Assessment Division, Hazardous Materials Response and

Assessment Division, Strategic Environmental Assessment Division, and the Damage

Assessment Center.
NOAA Technical Memorandum NOS ORCA 100

Contaminants in Aquatic Habitats at Hazardous Waste Sites: Mercury

Nancy Beckvar 1

Jay Field 1

Sandra Salazar 2

Rebecca Hoff 1

1NOAA/HAZMAT

2EVS Consultants

 Seattle, Washington

Seattle, Washington

United States National Oceanic and National Ocean Service

Department of Commerce Atmospheric Administration W. Stanley Wilson

Mickey Kantor D. James Baker Assistant Administrator

Secretary Under Secretary for Oceans for Ocean Services and

and Atmosphere Coastal Zone Management
Hazardous Materials Response and Assessment Division

Office of Ocean Resources Conservation and Assessment

National Ocean Service

National Oceanic and Atmospheric Administration

U.S. Department of Commerce

Silver Spring, Maryland

CITATION

Please cite this report as “Beckvar, Nancy, Jay Field, Sandra Salazar, and Rebecca Hoff. 1996.

Contaminants in Aquatic Habitats at Hazardous Waste Sites: Mercury. NOAA Technical

Memorandum NOS ORCA 100. Seattle: Hazardous Materials Response and Assessment Division,

National Oceanic and Atmospheric Administration. 74 pp.

NOTICE

This report has been reviewed by the National Ocean Service of the National Oceanic and

Atmospheric Administration (NOAA) and approved for publication. Such approval does not signify

that the contents of this report necessarily represent the official position of NOAA or of the

Government of the United States, nor does mention of trade names or commercial products constitute

endorsement or recommendation for their use.
Contents

Executive Summary................................................................................................................................................... i

Environmental Chemistry................................................................................................................i

Bioaccumulation.............................................................................................................................ii

Toxicity..........................................................................................................................................iii

Applications ...................................................................................................................................iv

Summary........................................................................................................................................vi

Acronyms ................................................................................................................................................vii

Glossary..................................................................................................................................................viii

Introduction ................................................................................................................................................................1

Environmental Chemistry.....................................................................................................................................1

Chemical Speciation .......................................................................................................................2

Distribution in the Environment...................................................................................................9

Bioaccumulation of Mercury..........................................................................................................................12

The Effect of the Form of Mercury on Bioaccumulation..........................................................13

Exposure Pathways......................................................................................................................17

Biological Factors Affecting Accumulation of Mercury...............................................................21

Other Factors Affecting Accumulation ........................................................................................23

Toxicity of Mercury...............................................................................................................................................25

Toxicity of Mercury in Water.....................................................................................................27

Toxicity of Mercury in Sediment.................................................................................................32

Toxicity Associated with Mercury in Tissues..............................................................................33

Interactions with Other Metals...................................................................................................36

Criteria and Guidelines ......................................................................................................................................36

Ambient Water Quality Criteria..................................................................................................36

Sediment Guidelines.....................................................................................................................38

Tissue............................................................................................................................................38
Contents, cont.

Applications ...............................................................................................................................................................38

Sampling and Monitoring Considerations...................................................................................39

Approaches to Remediation........................................................................................................41

Analytical Considerations.............................................................................................................43

Summary ..............................................................................................................................................44

References......................................................................................................................................................................45

Figure

1 Mercury speciation.........................................................................................................................3

Tables

1 Toxicity of Mercury in Water.....................................................................................................28

2 Toxicity Associated with Mercury in Tissues..............................................................................35
EXECUTIVE SUMMARY

This document reviews published literature on mercury chemistry,

bioaccumulation and toxicity, and is intended to serve as guidance for NOAA

Coastal Resource Coordinators in their work with EPA on hazardous waste sites.

The purpose of this document is to highlight factors to consider in designing and

evaluating ecological risk assessments; and, in the sampling, monitoring and

analyses of environmental media in aquatic habitats affected by mercury. Though

many questions about mercury remain, the science is evolving rapidly. This paper

should be reviewed with the knowledge that information can change as new

studies are published.

Environmental Chemistry

The fate of mercury in the environment depends on the chemical form of mercury

released and the environmental conditions. Elemental mercury, inorganic mercury,

and methylmercury are the three most important forms of mercury in natural

aquatic environments. Most mercury is released into the environment as inorganic

mercury, which is primarily bound to particulates and organic substances and may

not be available for direct uptake by aquatic organisms. The process of methylation

of inorganic mercury to methylmercury, which is highly bioavailable, is thus an

important key to the fate of mercury in the environment.

Elemental mercury has a high vapor pressure, a low solubility, does not combine

with inorganic or organic ligands, and is not available for methylation. The

mercurous ion (Hg[I]) combines with inorganic compounds only and cannot be

methylated. The mercuric ion (Hg[II]) combines with both inorganic and organic

ligands, and can be methylated. Methylation in aquatic habitats is primarily a

biological process. Mono- and dimethylmercury are formed by microorganisms in

both sediment and water through the methylation of inorganic mercuric ions

(Hg[II]). Dimethylmercury, which is highly volatile, is generally not persistent in

aquatic environments.

Methylation is influenced by environmental variables that affect both the

availability of mercuric ions for methylation and the growth of the methylating

microbial populations. Methylation rates are higher under anoxic conditions, in

i
freshwater compared to saltwater, and in low pH environments. The presence of

organic matter can stimulate growth of microbial populations (and reduce oxygen

levels), thereby enhancing the formation of methylmercury. Sulfide can bind

mercury and limit methylation. Methylmercury production can vary due to

seasonal changes in nutrients, oxygen, temperature, and hydrodynamics. In most

studies, methylation increased during the summer months when biological

productivity was high, and decreased during the winter months.

Measurements of total mercury concentrations in the sediment do not provide

information on the form of mercury present, methylation potential, or availability to

organisms locally and downstream. If environmental conditions are conducive for

methylation, methylmercury concentrations may be high in proportion to the

supply and distribution of total mercury.

Bioaccumulation

Mercury is accumulated by fish, invertebrates, mammals, and aquatic plants and

the concentration tends to increase with increasing trophic level (mercury

biomagnifies). Although inorganic mercury is the dominant form of mercury in the

environment and is easily taken up, it is also depurated relatively quickly.

Methylmercury accumulates quickly, depurates very slowly, and therefore

biomagnifies in higher trophic species. The percentage of methylmercury, as

compared to total mercury, also increases with age in both fish and invertebrates.

Uptake and depuration rates vary between tissues within an organism. Partitioning

of mercury between tissues within aquatic organisms is influenced by the chemical

form of mercury and route of exposure (ingestion or via the gills). Due to its

preferential uptake, ability to be transferred among tissues, and slow depuration,

most of the mercury in fish muscle tissue ( 99%) is methylmercury.

Marine mammal tissues have some of the highest concentrations of mercury found

in all marine organisms, with the liver generally having the highest total mercury

concentration. Although many juvenile and adult marine mammals primarily feed

on fish, which contain high percentages of methylmercury, high concentrations of

inorganic mercury are found in adult specimens. Apparently, adult marine

mammals can mineralize methylmercury into inorganic mercury. Juvenile marine

mammals have lower concentrations of total mercury than adults; but unlike fish

and invertebrates, the percentage of methylmercury is higher in juvenile mammals.

ii
Invertebrates generally have a lower percentage of methylmercury, as compared to

total mercury, in their tissues than do fish and marine mammals. The percentage of

methylmercury in invertebrates varies greatly and can range from one percent in

deposit-feeding polychaetes, to close to 100% in crab.

Bioconcentration factors (BCFs) reflect uptake from water in laboratory experiments.

BCFs for mercury are variable, with the highest factors determined for

methylmercury. BCFs for methylmercury in brook trout range from 69,000 to

630,000, depending on the tissue analyzed. BCFs for inorganic mercury (mercuric

chloride) in saltwater species range from 129 for adult lobster (Homarus

americanus) to 10,000 for oysters (Crassostrea virginica).

While sediment is usually the primary source of mercury in most aquatic systems,

the food web is the main pathway for accumulation. High trophic level species

tend to accumulate the highest concentrations of mercury, with concentrations

highest in fish-eating predators. Mercury concentrations in higher trophic species

often do not correlate with concentrations in environmental media. Correlations

have been made between sediment and lower trophic species that typically have a

high percentage of inorganic mercury, and between mercury concentrations in

higher trophic species and their prey items. The best measure of bioavailability of

mercury in any system can be obtained by analyzing mercury concentrations in the

biota at the specific site.

Toxicity

Toxicity is influenced by the form of mercury, the environmental media,

environmental conditions, the sensitivity or tolerance of the organism, and the life

history stage. Inorganic mercury is less acutely toxic to aquatic organisms than

methylmercury, but the range in sensitivity among individual species for either

compound is large. Toxicity was found to be greater at elevated temperatures,

lower oxygen content, reduced salinities in marine environments, and in the

presence of metals such as zinc and lead.

In general, toxic effects occur because mercury binds to proteins and alters protein

production or synthesis. Toxicological effects include reproductive impairment,

growth inhibition, developmental abnormalities, and altered behavioral responses.

iii
Reproductive endpoints are generally more sensitive than growth or survival, with

embryos and the early developmental stages the most sensitive. Mercury can be

transferred from tissues of the adult female to developing eggs. Exposure to low

concentrations of mercury may not result in mortality directly, but may retard

growth thereby increasing the risk of predation.

Data available on the effects of mercury-contaminated sediment on aquatic

organisms reviewed by Long and MacDonald (1992) resulted in effects range-low

(ERL) and effects range-median (ERM) concentrations of 0.15 mg/kg and

0.71 mg/kg, respectively. However, these numbers were less accurate than other

metals in predicting adverse effects, highlighting the need for site-specific effects

data to determine concentrations of mercury in sediment that pose a threat to

aquatic biota.

Few studies report both tissue residues and effects in long-term exposure to low

concentrations of mercury. However, results from studies on different freshwater

species indicate that reproductive effects could be expected to occur in sensitive

fish species at tissue concentrations close to the FDA action level of 1 mg/kg (ppm).

The interaction of mercury and other trace elements (e.g., cadmium, copper,

selenium, and zinc) can be both antagonistic and synergistic, primarily depending

on exposure concentrations and form of mercury. Effects were generally less than

additive (antagonistic) at lower exposure levels and greater than additive

(synergistic) at higher levels. Zinc and cadmium were reported to reduce the

teratogenic effects of methylmercury to killifish while selenium reduced mercury’s

toxic effects on development in medaka embryos.

Applications

Ecological assessments of waste sites with elevated concentrations of mercury in

the aquatic environment are particularly challenging due to the complexity of the

factors that affect the availability of mercury to aquatic organisms. Depending on

the magnitude of the problem (local versus system-wide), the level of effort

necessary to evaluate mercury contamination may range from simple monitoring of

chemical concentrations to more complex programs including monitoring of

numerous physical, chemical and biological parameters. The distribution of total

mercury in sediment, which in most cases is predominantly inorganic mercury, may

not by itself provide useful information about the bioavailability of mercury to

iv
aquatic species. Concentrations of total mercury in sediment that decrease with

increasing distance from the source may still pose a threat to organisms if the

bioavailability of the mercury increases (i.e., environmental conditions are more

conducive for methylation). Mercury concentrations in aquatic organisms,

particularly higher trophic-level organisms, may provide the best measure of the

availability of mercury in a particular area.

In sites where a whole system has been affected, evaluation of remedial alternatives

may need to be based on an understanding of the system-specific processes that

lead to increased methylation and the pathways to resources of concern. An

assessment of environmental parameters that affect the activity of methylating

microbes (e.g., nutrients, temperature, pH, and dissolved oxygen) and the factors

affecting the availability of inorganic mercury for methylation (e.g., the

resuspension of sediment, total organic carbon, and sulfides) may be warranted

when designing sampling plans for a remedial investigation.

To establish protective sediment target cleanup concentrations and remedial

options for mercury-contaminated sites, we must understand the extent of

contamination, the major pathways of transport, and bioavailability. Therefore,

data on the accumulation of mercury in tissues of aquatic organisms should be

included in assessment studies because it addresses potential human health

concerns and availability to aquatic receptors. In addition, studies should assess

toxicity to aquatic organisms, focusing on early life stages and reproductive effects.

Detection limits should reflect the program objectives. Contract Lab Program (CLP)

methodology may be appropriate for screening level assessments; however,

biologically relevant detection limits are often required and not available at CLP

laboratories. Thus, analytical laboratories that can achieve lower detection limits

may need to be used. Quality control is an important aspect of any testing program

but is particularly important when analyzing mercury in environmental matrices.

In water, very low concentrations need to be measured; the separation of the

different forms of mercury requires special analytical techniques. Matrix effects in

the extraction of mercury from tissue may interfere with accurate analyses for

methyl and total mercury. When analyzing mercury in water, sediment, and tissue,

analysis of certified standards for the appropriate matrix must be included as part of

the quality control plan.

v
Summary

NOAA recommends a site-specific approach that focuses on determining the

availability of mercury and the potential for toxic effects. The accumulation of

mercury in aquatic biota is often the primary concern at mercury sites and is

useful for assessing availability. Bioaccumulation studies should measure tissue

concentrations in more than one resident and/or transplanted caged species,

preferably with species representing different trophic levels or different food web

pathways. It may not be possible to correlate sediment mercury concentrations

with concentrations in biota. However, correlations between mercury

concentrations in predator and prey species may be useful in determining

pathways of mercury transfer.

Toxicity tests such as the standard amphipod tests should also be conducted to

assess mercury toxicity to benthic organisms. At major mercury sites, chronic

toxicicty endpoints should be included in the assessment—in particular, fish early

life stage or reproductive endpoint tests. Because of the persistence of mercury in

aquatic systems, source control alone may not be sufficient to permit recovery.

Additional remedial actions may be required to reduce the total mercury burden in

the system. Long-term monitoring of tissue concentrations of mercury in aquatic

biota is needed to assess remedial effectiveness.

vi
ACRONYMS AND ABBREVIATIONS

AVS acid-volatile sulfides

AWQC Ambient Water Quality Criteria

BAFs Bioaccumulation factors

BCFs Bioconcentration factors

Cd cadmium

CLP Contract Lab Program

Cu copper

DOC dissolved organic carbon

Eh oxidation-reduction potential

ERL effects range-low

ERM effects range-median

FDA U.S. Food and Drug Administration

Fe iron

Hg mercury

Mn manganese

MT metallothioneins

μeq/l micro equivalent per liter

μg/l micrograms per liter

μm micrometer

mg/kg milligrams per kilogram

mg/l milligrams per liter

ppm parts per million

Se selenium

SRB sulfate-reducing bacteria

TOC total organic carbon

Zn zinc

vii
GLOSSARY

acid labile mercury

Determined by SnCl2 reduction on acidified samples, includes inorganic

complexes, labile organic associations, elemental mercury, and labile

particulate mercury. Doesn't measure organic forms (C-Hg bound) of

mercury such as methylmercury. Same as reactive mercury.

acid soluble mercury

Mercury that passes through a 0.45 μm membrane filter after the sample is

acidified to pH1-2.0 with nitric acid (EPA 1984). Strongly sorbed Hg is not

measured, but all toxic forms as well as some non-toxic forms are measured.

alkylmercury

Includes phenyl-, monomethyl-, and dimethylmercury

bioaccumulation

Net uptake of a contaminant into tissue from all pathways

bioaccumulation factors (BAF)

Ratio of tissue concentration to concentration in medium, with exposure

from the food web and the medium

bioconcentration factors (BCF)

Ratio of tissue concentration to concentration in medium, with exposure

only through the medium

biomagnification

Tissue concentration increases as trophic level increases

DOC (dissolved organic carbon)

Includes all sources of carbon, including humic and fulvic matter as well as

carbohydrates, proteins, etc.

demethylation

Conversion of methylmercury back to an inorganic form.

depuration

Elimination of a contaminant from the body of an organism

dimethylmercury

Organic form of mercury consisting of a single mercury atom and two

methyl groups [(CH3)2Hg], highly volatile and not persistent in the

environment.

Effects range-low

Concentration equivalent to that reported at the lower 10 percentile of the

viii
available screened sediment toxicity data for predominantly marine and

estuarine toxicity studies. This number represents the low end of the range

of concentrations at which effects were observed in the studies compiled

(Long and MacDonald 1992).

Effects range-median

Concentration equivalent to that reported at the 50th percentile of the

available screened sediment toxicity data for predominantly marine and

estuarine toxicity studies (Long and MacDonald 1992).

elemental mercury

Not in ionic form, cannot be methylated

halide

Binary compound of a halogen (fluorine, chlorine, bromide, iodine and

astatine)

inorganic mercury

Includes elemental mercury and mercury bound to other inorganic

molecules and compounds, including inorganic ligands and sulfides.

labile mercury

Includes Hg(OH)2, HOHgCl, HgCl2, and weakly bound organo-complex

forms

ligand

Any group, ion, or molecule that binds to another, called a receptor

methylmercury

Includes both mono- and dimethylmercury

mercuric-Hg[II]

Forms both inorganic and organic complexes, this is the only ionic form that

can be methylated

mercurous-Hg[I]

Forms inorganic complexes, cannot be methylated

methylation

Addition of a methyl (CH3) group

monomethylmercury

Organic form of mercury with one methyl group attached to a mercury atom

(CH3Hg) - highly toxic and readily accumulated by living organisms

organic mercury

Includes mercury complexes with organic ligands (e.g. humic/fulvic acids,

ix
amino acids, but without a Hg-carbon bond) and organic mercury bound via

a carbon atom (CH3-HgOH, CH3HgCl, CH3HgCH3)

organo-mercury

Mercury compounds with a Hg-carbon bond

pinniped

Mammals of the family Pinnipedia, includes all seals and walruses

piscivorous

Feeds on fishes

reactive mercury

“Easily reducible,” determined by SnCl2 reduction on acidified samples,

includes inorganic complexes, labile organic associations, elemental

mercury, and labile particulate mercury, doesn't include C-Hg bound

mercury such as methylmercury and dimethylmercury. Same as acid labile

mercury.

total mercury

Includes all forms of mercury

x
INTRODUCTION

As a trustee for marine, estuarine, and anadromous resources, NOAA is responsible

for ensuring the well-being of those trust resources potentially affected by releases

from hazardous waste sites. Metals, in particular, pose a threat because of their

persistence and toxicity in aquatic environments. The importance of mercury in

many aquatic environments is underscored by the fact that 35 states have fish and

wildlife consumption advisories in place for mercury (EPA 1996).

Although the hazards of mercury to humans are well-known, less information is

available on the risks to aquatic organisms. In order to define and address the

potential effects and extent of mercury contamination at hazardous waste sites, a

number of questions are often asked: Is the mercury present in a bioavailable

form? What concentrations in water, sediment, and tissues are potentially harmful

to aquatic resources? What types of sampling and analysis are necessary to define

potential risks to aquatic organisms? Is there a relationship between sediment

concentrations and tissue concentrations in aquatic organisms? What level should

be used for cleanup? What are important factors to consider when selecting a

remedy and designing monitoring plans at a site?

ENVIRONMENTAL CHEMISTRY

Mercury is among the most toxic of the heavy metals, has complex behavior in the

environment, and may persist for decades following abatement of the source.

Mercury’s environmental persistence is due in part to its high affinity for

particulates and organic matter. Even if mercury concentrations in sediment and

water decrease over time, concentrations in organisms may not decrease due to the

slow rate of elimination of the highly bioavailable methylmercury form. The

physical properties, bioavailability, and toxicity of mercury are governed by

speciation into both organic and inorganic forms.

Elemental mercury, bivalent inorganic mercury, and monomethylmercury are the

three most important forms of mercury occurring in natural aquatic environments

(Battelle 1987). Elemental mercury in aquatic environments has a high vapor

pressure, a low solubility in water, and an octanol-water partition coefficient

(Kow)=4.15 (Shoichi and Sokichi 1985 as cited in Major et al. 1991). Elemental and

1
dimethylmercury can occur as dissolved gaseous mercury. Mercury can also occur

as particulate and dissolved ionic and monomethylmercury species. In natural

water, ionic mercury is consumed by methylation, reduction, and particulate

scavenging (Mason et al. 1995a). Bivalent inorganic mercury binds to inorganic and

organic ligands, especially sulfur-containing ligands, and forms both inorganic and

organic complexes. Figure 1 shows a schematic of some common pathways of

mercury speciation in the environment.

Although most mercury occurs in the inorganic form, methylmercury, an organic

form, is the most toxic and readily bioaccumulated form of mercury.

Methylmercury normally occurs in the environment at extremely low

concentrations; however, it is taken up easily by aquatic organisms and

bioaccumulated. Consequently, methylmercury may comprise more than 95% of

the mercury in fish tissue while only 5-15% of the total mercury burden in

sediments and water of contaminated lakes is methylmercury (Saroff 1990).

Chemical Speciation

Inorganic and total mercury versus methylmercury

Chemical speciation terms commonly used include total, inorganic, organic, and

methylmercury and are based on the oxidation state and associated compounds.

Mercury has three stable oxidation states: the native element (Hg[0]), mercurous

(Hg[I]), and mercuric (Hg[II]). Inorganic mercury includes elemental Hg and some

complexes of the mercurous and mercuric oxidation states. Hg [I] forms inorganic

compounds only and, like Hg[0], cannot be methylated. Hg [I] compounds include

mercurous salts (halides) and mercurous chlorides such as calomel. Both Hg[0] and

Hg[I] can be oxidized to form Hg[II]. Hg[II] (bivalent mercury), the form that can be

methylated, forms both organic and inorganic compounds. Mercury [II] can

combine with inorganic ligands including chloride, hydroxide, nitrate, and sulfate

anions (Benes and Havlik 1979) to form inorganic mercury compounds that

include mercuric halides, mercuric chloride (cinnabar), and mercuric sulfides.

Chloride concentration and pH affect the proportions of the uncharged inorganic

species in solution. For example, at low chloride concentrations most of the

inorganic mercury occurs in the form of mercuric hydroxide (Hg(OH)2), with

mercuric chloride (HgCl2 ) and mercury hydroxide chloride (HgOHCl) also

important (Mason et al. 1996). As chloride concentration increases

2
(CH3)2 Hg

(CH3)2 Hg

X = Cln, (OH)n, (SH–), humic/fulvic aci

CH

(CH3)2 Hg

dimethyl mercury

CH3 

p H

bacteria

p H

bacteria

CH3 HgX

methyl mercury

Figure 1. Important pathways of mercury speciation in the aq

3

3
(e.g., high-chloride lake water), the proportion of HgCl2 increases and the other two

species constitute only a few percent of the total inorganic mercury. As pH

increases, more Cl is needed than at lower pHs to increase the percent of HgCl2. At

even higher chloride concentrations, HgCl4

2- becomes the dominant species. This

speciation chemistry affects the accumulation and toxicity as described later in this

report.

The term organic mercury can include different types of organically bound mercury.

Hg [II] combines with organic compounds (humic/fulvic acids, amino acids) via an

organic ligand bond to form organomercury salts. The Hg [II]-organic ligand bond

is relatively weak compared to a C-Hg bond. The organomercury salts resemble

their corresponding inorganic mercuric salts in their properties and reactions. The

organomercury compounds methylmercury, dimethylmercury, and phenylmercury

have a C-Hg bond; methyl (CH3) and phenyl groups (C6H5) link to a mercury

atom via a carbon atom. Some authors group the organically bound Hg [II]

complexes (without a C-Hg bond) with inorganic mercury compounds, while others

group all organically bound mercury together as organic mercury.

Mercury compounds may be grouped according to their form based on chemical

speciation discussed above, or based on the analytical technique used to measure

the mercury. The analytical technique determines which forms of mercury are

detected. Mercury terms based on analytical procedures include names such as

acid-soluble, reactive or acid labile, and calcium chloride-extractable.

Measurements of reactive mercury include Hg [II] bound to inorganic substances

and weakly bound to organic matter (however, methylmercury is not included due

to the strong C-Hg bond). To estimate methylmercury concentrations, some

authors measure total mercury and reactive mercury, then subtract the reactive

from the total measurement. Total mercury measurements include all the various

inorganic and organic forms of mercury, including dissolved, colloidal, and/or

particulate states. Analytical groupings are defined in the glossary. Unless the

specific type of mercury is mentioned, the reader should assume that the term

“mercury” refers to total mercury.

Speciation of mercury is affected by environmental conditions such as pH,

oxidation-reduction potential (Eh), oxygen content, sulfide content, chloride

4
concentration, organic matter content, and microbial activity. Similarly, biological

and chemical processes control the conversion of inorganic mercury to

methylmercury. The factors that enhance and inhibit methylation, and affect the

distribution of inorganic and methylmercury are discussed in this section of the

paper.

Methylation of Mercury

In both freshwater and saltwater environments, mercury is converted from

inorganic bivalent mercury (Hg[II]) to methylmercury primarily by microorganisms

(Berman and Bartha 1986), although chemical methylation also occurs (Craig and

Moreton 1985; Weber 1993). Two forms of monomethylmercury, methylmercuric

hydroxide (CH3HgOH) and methylmercuric chloride (CH3HgCl) occur in both

fresh and saltwater, with the former dominant at low chloride concentrations (low

chloride freshwater) and the latter dominant at high chloride concentrations (high

chloride lakes, seawater). As with inorganic mercury, the organic chloride species

(Kow = 1.7) is more hydrophobic than the hydroxide species (Kow = 0.07; Major et

al. 1991; Faust 1992; Mason et al. 1996). Dimethylmercury (Kow = 182) readily

volatilizes from surface water and is generally not persistent in aquatic

environments at concentrations of concern; therefore, discussions of

methylmercury in this review refer to the monomethylmercury species, unless

otherwise stated.

Methylmercury production depends on both the availability of Hg[II] for

methylation and microbial activity. Methylation is usually greatest at the sedimentwater interface, but also occurs in the water column. Net methylmercury

production is a function of both the rate of methylation and the rate of

demethylation (Korthals and Winfrey 1987). Methylmercury is not readily

decomposed so the methylation rate is usually higher than the demethylation rate.

Degradation of methylmercury is also primarily a microbial process.

Methylation is influenced by the availability of Hg[II], oxygen concentration, pH,

redox potential (Eh), presence of sulfate and sulfide, type and concentrations of

complexing inorganic and organic agents (Parks et al. 1989), salinity (Blum and

Bartha 1980), and organic carbon (Jackson 1989; Winfrey and Rudd 1990). Strongly

bound Hg[II] is not available for methylation. For example, insoluble mercuric

5
sulfide (HgS) will be methylated in aerobic sediments at rates 100 to 1,000 times

slower than for the less strongly bound HgCl2 (Olson and Cooper 1976).

Anaerobic, sulfate-reducing bacteria (SRB) are the primary methylators of mercury

in both lacustrine and estuarine sediments (Compeau and Bartha 1985; Gilmour

and Henry 1991). The primary methylators of mercury in the water column have

not been identified. SRB are common in sulfate-rich estuarine sediments (Hines et

al. 1989) but are more limited in freshwater sediment with lower sulfate

concentrations. A sulfate concentration of 200-500 μM in the water column is

optimal for mercury methylation by SRB in sediment (Gilmour and Henry 1991).

The activity of the methylating microbes is affected by environmental conditions

(Jackson 1986) with nutrient availability and seasonality particularly important.

The concentration of inorganic mercury in environmental media may not be a good

indication of the concentration of methylmercury present due to the influence of

environmental variables and biological activities. The importance of environmental

factors in the production of methylmercury is as follows:

pH: Neutral or low pH conditions favor the production of monomethylmercury

over dimethylmercury (Beijer and Jernelov 1979). An alkaline (high) pH favors the

formation of dimethylmercury, which tends to escape into the atmosphere.

Elevated tissue concentrations of methylmercury have been noted in numerous

pristine lakes of the northern United States and Canada that receive acid rain and

no point sources of mercury (Xun et al. 1987; Bloom et al. 1991). The mechanism(s)

causing increased bioaccumulation in low pH lakes are not understood (Ramlal et

al. 1985; Winfrey and Rudd 1990; Richardson and Currie 1996). The factors

primarily responsible for net methylmercury production in lakes are, in decreasing

order of importance, pH, dissolved organic carbon (DOC) concentration, and

microbial respiration (Miskimmin et al. 1992). The importance of pH and sediment

properties (Fe and Mn content) on methylation rates in saltwater environments has

not been well studied.

Sulfide, Sulfate, and Other Ions: In the presence of sulfides, the mercuric ion

(Hg[II]) becomes tightly bound to sulfide as insoluble mercuric sulfide and is not

available for methylation. Sulfide activity may be the main factor influencing the

availability of Hg[II] (Bjornberg et al. 1988) and the concentration of methylmercury

in sediment (Craig and Moreton 1983). If pH is high or Eh is low, sulfide activity

6
will be high and mercury will be precipitated as insoluble mercuric sulfide. If the

sulfide is oxidized to sulfate, the mercuric ion will become available for

methylation. Both free sulfides and acid-volatile sulfides (AVS) appear to inhibit

methylation (Gilmour and Capone 1987). The presence of other minerals may

affect this relationship. Excess ferrous iron has been found to bind the sulfide and

limit its Hg-binding effectiveness such that no difference in methylation rates are

noted between sulfide-rich and sulfide-poor sediments (Rudd et al. 1983). Selenium

similarly binds the Hg [II] ions and reduces their availability for methylation. The

redox cycling of manganese in lakes may be more important than iron-scavenging

of mercury (Bonzongo et al. 1996).

Addition of sulfate to anoxic lake sediment slurries or the overlying water column

can increase methylmercury production by stimulating the SRB population

(Gilmour et al. 1992). SRB can both methylate mercury and produce sulfide, which

inhibits methylation: the kinetics of this are not understood. Mercury methylation

was once viewed as a detoxification process in SRB, but it may ultimately serve

some other function (Gilmour and Henry 1991).

Oxygen Conditions/Eh : Although methylation occurs under both aerobic

(oxidizing) and anaerobic (reducing) conditions, methylation is greater under

anaerobic conditions (Callister and Winfrey 1986; Weis et al. 1986; Regnell 1994).

In addition, demethylation rates are lower under anaerobic conditions, so the net

methylmercury production is higher in oxygen-depleted environments (Jackson

1987). Over 90 percent of methylmercury is formed biochemically in anaerobic

sediment (Berman and Bartha 1986).

In anoxic lake bottoms containing hydrogen sulfide, mercury is bound to sediment

as insoluble mercuric sulfide. If conditions become aerobic due to a decrease in the

organic load or seasonal turnover, sulfide can be oxidized to sulfate, releasing the

mercury in the ionic form Hg[II], which is available for methylation (Jernelov 1968).

Nutrients /Organic content :/DOC : Nutrients can enhance the rate of

methylmercury production by stimulating the methylating bacteria. Decaying

organic matter can enhance microbial activity and create low oxygen conditions,

both of which cause higher methylation rates (Olson and Cooper 1974; Gilmour

and Henry 1991). In freshwater areas with a high organic input, methylation rates

7
can become locally elevated if other environmental conditions do not inhibit

methylation (i.e., high sulfide levels; Jackson 1986). When sulfide and sulfate

concentrations are not limiting, organic matter may be the major factor controlling

mercury methylation rates in estuarine sediments (Choi and Bartha 1994).

Increased DOC levels may inhibit methylation due to the binding of free mercury

ions (Jackson 1989; Winfrey and Rudd 1990) even though supplemental DOC

increases microbial respiration (Miskimmin et al. 1992). In clear freshwater lakes,

DOC and pH may interact such that less of the Hg [II] is bound by DOC at low pH,

resulting in higher methylation rates. Acid rain may also limit the amount of DOC

transported into a system because at lower pH, DOC solubility and mobility is

reduced (De Haan 1992; Schindler et al. 1992).

Humic/Fulvic material: The geochemistry of Hg in lake and stream water may be

dominated by humic material interactions (Mierle and Ingram 1991). Hg complexes

with humic and fulvic substances and Hg retention and export from watersheds in

Canada have been correlated with the export of humic substances. Hintelmann et

al. (1995) assumed that the methylmercuric ion is bound to sulfidic binding sites of

humic acid. At lower pHs, the amount of free unbound methylmercury ion was

higher in their laboratory study of humic and fulvic acids. Acidification could

potentially release bound methylmercury from humic acids into the aqueous phase

where it would be readily bioavailable.

Salinity : There appears to be a negative correlation between the rate of

methylmercury formation and salinity in estuarine sediments (Blum and Bartha

1980). The rate is lower in more saline environments because the bicarbonate

component of seawater slows methylation of Hg [II] under both aerobic and

anaerobic conditions (Compeau and Bartha 1983). The release of reactive Hg [II]

and Hg [0] is slowed when chloride ions bind to mercury, thereby inhibiting

methylmercury formation (Craig and Moreton 1985). Salinity also affects

methylation due to the high pore-water sulfide concentrations as a result of rapid

sulfate reduction in saline water compared to sulfate-limited freshwater

environments (Gilmour et al. 1992). Along a salinity gradient in the lower Hudson

River, methylation rates decreased downriver with increasing salinity and sediment

sulfide concentrations (Gilmour and Capone 1987).

8
The percentage of total mercury that is methylmercury is higher in freshwater

sediments (up to 37%) and water (up to 25% in aerobic water and 58% in anoxic

bottom water) than in estuarine and marine water (<5%) and associated sediments

(<5%) (Gilmour and Henry 1991). Dissolved reactive mercury (inorganic species)

forms the majority of the total mercury in open oceans (Bloom and Crecilius 1983;

Gill and Fitzgerald 1987).

Season : Biological productivity of methylating microbes is affected by seasonal

changes in temperature, nutrient supply, oxygen supply, and hydrodynamics

(changes in suspended sediment concentrations and flow rates). Methylmercury

concentrations varied seasonally by an order of magnitude at most sites studied

(Parks et al. 1989). Methylation may tend to increase during the summer months

when biological productivity and temperature are high and decrease during winter

months when biological productivity and temperature are low (Callister and

Winfrey 1986; Jackson 1986; Weis et al. 1986; Korthals and Winfrey 1987; Parks et

al. 1989; Kelly et al. 1995; Leermakers et al. 1995). Although the potential

methylmercury production is greatest during the summer, actual production may

not peak during this time (Kelly et al. 1995). In Onondaga Lake, New York, the

mercury species in the water column varied temporally (Battelle 1987; Bloom and

Effler 1990). Total mercury concentrations may also vary seasonally due to

physical factors such as winter storms resuspending mercury-contaminated

sediments (Gill and Bruland 1990).

Distribution in the Environment

The distribution and abundance of inorganic mercury and methylmercury in the

environment may vary independently as they are controlled by different

physicochemical processes. The concentration of total mercury (which is mainly

inorganic) in the environment is generally not a good predictor of methylmercury

concentration (Gilmour and Henry 1991; Kelly et al. 1995). Inorganic mercury has

a high affinity for sediments; a significant portion of the total mercury in fresh

water is in particulate form (Gill and Bruland 1990). Most of the mercury in

estuaries was associated with particulate matter (Cossa and Noel 1987). The

environmental distribution of inorganic mercury appears to be controlled by

processes such as transport, sorting, and sedimentation as related to the hydrologic

regime. Resuspension and resettling of sediments caused persistently high

concentrations of mercury in the surface sediments of Lavaca Bay (Reigel 1990).

9
Total mercury concentrations in surface water may decrease as mercury bound to

particulate matter settles or is transported downstream (Bonzongo et al. 1996). The

distribution of biotically produced methylmercury initially depends on the

microbial populations that methylate the mercury. Although more abundant in the

sediment where it is formed, methylmercury forms a greater percentage of the total

mercury in the overlying water column (Gilmour et al. 1992).

The distribution of both inorganic and methylmercury is also affected by the largerscale physical characteristics of the environment such as type of system (river, lake,

estuary, ocean) and its physical configuration, water circulation patterns, catchment

type, sediment characteristics, rainfall, and the introduction of terrestrial sediments.

The physical characteristics of the system influence the mechanisms of mercury

distribution, availability, and cycling in varying degrees. Water flow regimes in

particular subenvironments (mainstream versus backwater), characteristics such as

the stratification cycle and amount of nutrients in lakes, or the configuration of the

estuary (open circulation versus restricted), may have very different, distinct

features that control both the persistence of mercury in the environment and how

the different species of mercury behave and are distributed among sediment, water,

and organisms.

In freshwater systems, Kelly et al. (1995) found a predictive, linear relationship

between total and methylmercury concentrations in unfiltered water samples from

some specific lake systems, but not from stream systems. Runoff from wetland

catchments contributed more methylmercury to lake systems than did runoff from

upland catchments (St. Louis et al. 1994). In estuarine systems, total dissolved

mercury concentrations were found to be enhanced where salinity was less than 10

ppt, coinciding with the maximum turbidity zone (Cossa and Noel 1987; Cossa et

al. 1988).

Terrestrial sediment influxes can also affect mercury availability. In the estuarine

Ala Wai Canal in Hawaii, total mercury increased over two orders of magnitude in

polychaetes and shrimp during the rainy season (Luoma 1977). Mercury bound to

freshwater sediments and introduced into the estuary from urban runoff during

rainfall was desorbed upon contact with saline water. The increased concentration

of dissolved mercury temporarily increased total mercury in filter-feeding worms

and shrimp. Total mercury concentrations in the water column and biota

10
decreased after the runoff stopped. In contrast, lower mercury concentrations were

found in plankton in a freshwater lake after input of high concentrations of clean,

fine-grained sediment. Sediments washed into the lake during rainfall bound the

mercury, inhibiting uptake by plankton (Jackson 1988).

Sediment composition can also affect the way that mercury is distributed in the

environment. Mercury concentrations in freshwater benthic organisms appeared to

be determined by the sediment composition, such as the concentration of hydrated

Fe and Mn oxides and carbon-rich humic matter in bottom sediment. The mercury

appeared to be less available when it was bound by iron hydroxide (FeOOH),

manganese hydroxide (MnOOH), and possibly by higher-molecular-weight humic

substances (Jackson 1988).

In a freshwater river-lake system in Canada (Parks et al. 1989) methylmercury

concentrations in surface water were highest 80 kilometers downstream from the

most contaminated sediments (contaminated with inorganic mercury). Fish were

contaminated as far as 270 km downstream from the inorganic mercury source,

with the most contaminated fish found more than 100 km downstream of this

source. Methylmercury concentrations in the water increased as inorganic mercury

concentrations in the sediment decreased. The most highly contaminated

sediments were located near a sewage outfall. The researchers surmised that the

mercury in these sediments was bound to sulfide and thus not available for

methylation. Further downstream the mercury became available for methylation

probably due to a decrease in sediment sulfide levels. Even though concentrations

of inorganic mercury in the sediment here were much lower than upstream

concentrations, methylmercury production was much higher and biota were more

contaminated.

There may be a similar situation in low-salinity water of Berrys Creek, New Jersey,

where high inorganic sediment mercury concentrations were also found next to a

sewage outfall (Weis et al. 1986). However, concentrations of mercury in fish

inhabiting the area were not as high as expected. This was attributed to the

presence of sulfide, which binds mercury and limits methylation. No downstream

studies have been conducted to determine whether the mercury is more

bioavailable further from the source of high sulfide concentrations.

11
An eight-ton cargo of elemental mercury located within the hold of the sunken

Empire Knight in offshore marine waters did not contaminate invertebrates living

outside the hold of the ship (Hoff et al. 1994). Only a small percentage of

invertebrates sampled from within the hold had elevated concentrations of total

mercury. However, it is not know whether the mercury was incorporated into the

tissue of the organisms. The large source of elemental mercury in this environment

was not bioavailable to organisms located away from the source.

BIOACCUMULATION OF MERCURY

Mercury bioaccumulates in aquatic plants, invertebrates, fish, and mammals.

Concentrations increase (biomagnify) in higher-trophic-level organisms. Even

though the different types of mercury have relatively low Kow values (compared to

organic compounds such as PCBs), they are readily accumulated. Inorganic

mercury (excluding elemental) and methylmercury’s strong reactivity with

intracellular ligands is thought to be responsible for their high degree of

accumulation. Uptake and accumulation of mercury are affected by the type of

mercury present, with neutral mercury species (e.g., HgCl2

0 and CH3HgCl0)

absorbed more efficiently than charged mercury species (e.g., HgCl- 3 CH3Hg+;

Mason et al. 1996).

Despite the fact that the neutral inorganic and organic complexes have similar lipid

solubilities, methylmercury is selectively accumulated (due to a higher transfer

efficiency and lower rate of elimination), resulting in biomagnification in higher

trophic levels (Mason et al. 1995b). Inorganic mercury species are not biomagnified

(Surma-Aho and Paasivirta 1986; Riisgård and Hansen 1990; Hill et al. 1996).

Environmental factors that enhance mercury methylation result in greater

bioavailability and accumulation of methylmercury. Environmental variables also

influence the bioavailability and accumulation of inorganic mercury. Although

concentrations of mercury in the environment may correlate with concentrations in

resident plants and biota, correlation is often difficult. Correlating total mercury in

sediment with total mercury in upper-trophic-level organisms is complicated by

high methylmercury concentrations in high-trophic-level organisms relative to low

methylmercury concentrations in the environment.

12
Tissue concentrations of mercury are often positively correlated with organism

length, weight, and/or age. Diet has a significant role in the overall body burden of

mercury, both between and within species. Differences in total mercury

concentrations between species reflect diet differences due to trophic position;

within-species differences are related to dietary requirements of various

developmental stages.

The Effect of the Form of Mercury on Bioaccumulation

Both inorganic and methylmercury are taken up directly from water and food (or

ingested sediment). However, methylmercury is more efficiently accumulated than

inorganic mercury for most aquatic organisms (Fowler et al. 1978; Julshamn et al.

1982; Riisgård and Hansen 1990; Mason et al. 1995b). The uptake and depuration

of mercury depends on the form of mercury, source of mercury (water or food), and

the type of receptor tissue, resulting in different patterns of accumulation.

Methylmercury is readily transferred across biological membranes. Within the

organism, methylmercury is strongly bound to sulfhydryl groups in proteins of

tissues such as muscle, and is much slower to depurate than inorganic mercury.

Thus, methylmercury has a much greater potential for bioaccumulation and a

longer half-life in organisms than inorganic mercury.

Fish

The accumulation of mercury from water occurs via the gill membranes. Gills take

up aqueous methylmercury more readily than inorganic mercury (Huckabee et al.

1979; Boudou et al. 1991). Methylmercury is eventually transferred from the gills to

muscle and other tissues where it is retained for long periods of time (Julshamn et

al. 1982; Riisgård and Hansen 1990).

Inorganic mercury taken up with food initially accumulates in the tissues of the

posterior intestine of fish (Boudou et al. 1991). Inorganic mercury is not easily

transferred through this organ to other parts of the body. After 15 days, 80% had

depurated from the fish intestine. Liver and kidney in fish tend to have higher

percentages of inorganic mercury than muscle tissue, although percentages vary by

organ and species (Windom and Kendall 1979; Riisgård and Hansen 1990).

13
Methylmercury ingested in food is efficiently transferred from the intestine to other

organs (Boudou et al. 1991). Methylmercury has been reported to constitute from

70 to 95% of the total mercury in skeletal muscle in fish (Huckabee et al. 1979; EPA

1985; Riisgård and Famme 1988; Greib et al. 1990; Spry and Wiener 1991).

Methylmercury accounted for almost all ( 99%) of the mercury in muscle tissue in

a wide variety of both freshwater and saltwater fish found in waters not highly

contaminated by other organomercurial species (Bloom 1992).

The ratio of liver to muscle total mercury concentration usually fluctuates around

one and can reflect the exposure history of the organisms. For example, the

liver:muscle ratio may be less than one in chronically exposed fish, while a recent

exposure to mercury may result in a ratio greater than one (Riisgård and Hansen

1990).

McKim et al. (1976) reported that mercury could be transferred from adult to

offspring in brook trout. Exposure of the parent population to aqueous

methylmercury concentrations of 0.03 to 2.93 μg/l in the laboratory resulted in

mercury concentrations as high as 2 mg/kg in their embryos. Total mercury

concentrations in eggs of several species of adult fish from Swedish lakes were

much lower than concentrations in other tissues; therefore, spawning did not lower

their total mercury body burden (Lindqvist 1991).

The main depuration pathway is through the kidney and liver in fish. Half-lives for

methylmercury in fish range from one to three or more years (McKim et al. 1976;

Pentreath 1976a, b; Riisgård and Famme 1986; Riisgård and Hansen 1990), while

estimates of half-lives for inorganic mercury are much lower, ranging from

approximately five days to five months (Pentreath 1976a, b; Huckabee et al. 1979).

Invertebrates

Invertebrates accumulate and partition inorganic and methylmercury in tissues

similar to the trends exhibited by fish ( Fowler 1978; Riisgård and Famme 1986;

Saouter et al. 1991; Saouter et al. 1993). However, invertebrates generally contain a

lower percentage of methylmercury than fish or mammals (Lasorsa and Allen-Gil

1995), with highly variable concentrations. This wide variation of mercury content

in invertebrates is most likely a function of different feeding strategies (and trophic

14
levels) and different environmental exposures. Reported percentages of

methylmercury compared to total mercury concentrations are less than 1% for the

polychaete Nereis succinea (Luoma 1977); 10% in copepods, mussels and shrimp

(Horvat 1991); 10-100% in the cockle (Møhlenberg and Riisgård 1988); 16% in

urchin gonads (Eganhouse and Young 1978); 30-90% in lake zooplankton

(Lindqvist 1991); 87% in crab muscle (Eganhouse and Young 1978); and 100% in

red rock crab, Dungeness crab, and spot shrimp (Bloom 1992). Becker and Bigham

(1995) found an increasing percentage of methylmercury compared to total

mercury in higher trophic levels in the Onondaga lake food web. Lake water

contained 5% of total mercury as methylmercury; phytoplankton 24%; benthic

macroinvertebrates 26%; zooplankton 40%; and fish fillets 96%.

Viscera in mussels contained the highest tissue concentration of total mercury

(Fowler et al. 1978). The total mercury concentration was highest in the midgut

and muscle tissue in crab (Bjerregaard and Christensen 1993) and in the viscera in

shrimp (Fowler et al. 1978). The shrimp molts had the lowest mercury content;

therefore, molting is not considered an important depuration pathway in

crustaceans (Fowler et al. 1978).

Half-lives for total mercury in salt-water mussels ranged from two months to one

year (Riisgård et al. 1985). Inorganic mercury was eliminated more rapidly than

methylmercury in mussels and shrimp (Fowler et al. 1978).

Marine mammals

Marine mammals have some of the highest tissue mercury concentrations of all

marine organisms investigated (Andre et al. 1991a); however, concentrations are

highly variable both within and among species. These variations have been

attributed to collection locations (Wren 1986), concentrations in prey items (Szefer

et al. 1993), and organism age (Julshamn et al. 1987; Thompson 1990). For

example, species that feed primarily on benthic invertebrates, such as walruses and

baleen whales, tend to have relatively low mercury concentrations. In contrast, fisheating species, such as porpoises and seals, exhibit relatively high mercury

concentrations (Born et al. 1981).

15
In contrast to fish, adult marine mammals have a much higher percentage of total

mercury as inorganic mercury, although the concentration of methylmercury may

also be elevated. Less than 10% of the total mercury content is methylmercury

(Eisler 1987). Juveniles tend to have higher percentages of methylmercury.

The liver generally exhibits the highest total and methylmercury concentration

(Holden 1978; Wagemann et al. 1983; Julshamn et al. 1987; Thompson 1990; Andre

et al. 1991a), followed by kidney and muscle tissues (Szefer et al. 1993). Julshamn

et al. (1987) measured the highest concentrations of methyl- and total mercury (13

and 150 mg/kg) in pilot whale livers (Globicephalus meleanus) compared to total and

methylmercury concentrations in muscle (2.8 mg/kg total mercury; 1.7 mg/kg

methylmercury) and kidney (15.3 mg/kg total mercury; 5.1 mg/kg methylmercury).

Andersen et al. (1987) also measured the highest methylmercury concentrations in

pilot whale liver (20 mg/kg; 14% of total mercury). The fraction that is methylated,

however, is usually lower in the liver compared to muscle and kidney. The

methylated fraction of total mercury ranged from 1% to 36% in seal liver (Holden

1978); 30% in older specimens to 100% in young specimens in the muscle of

harbor porpoise (Joiris et al. 1991); and 24% to 86% in the muscle of pilot whales

(Globicephalus meleanus; Julshamn et al. 1987). The liver:muscle ratio for

methylmercury concentration in harbor porpoises was approximately one, while the

ratio for liver:muscle for total mercury concentration was two. In some of the

harbor porpoises and some other species (sperm whale, common dolphin, and adult

bottle-nose dolphin) the liver:muscle ratio for total mercury ranged up to 20 (Joiris

et al. 1991) while the liver:muscle ratio for methylmercury was still one.

Schintu et al. (1992) observed an age-related change in the percentage of

methylmercury compared to total mercury in pilot whale livers. The liver of threeto seven-year old pilot whales (with a relatively low total mercury body burden)

contained 30% to 60% organic mercury, compared to 3% to 17% organic mercury

in livers of 30- to 40-year old pilot whales (with a relatively high total mercury

load). Porpoises exhibited a similar trend. Juveniles had a higher percentage of

methylmercury in liver (100%), while the percentage of methylmercury in adult

specimens decreased to 2 to 3% of the total mercury (Joiris et al. 1991).

Although mercury in the diet of many marine mammal species is predominantly

methylmercury, it has been proposed that the mammals are able to mineralize

methylmercury into the more harmless inorganic form, which then accumulates in

the liver of adult specimens (Holden 1978; Joiris et al. 1991). The estimated half-life

16
of total mercury in pinnipeds and dolphins is about 1.4 and 2.7 years, respectively

(Eisler 1987; Andre et al. 1991b).

Plants

Vascular plants accumulate both inorganic and methylmercury from sediment and

water in root, stem, and leaf sections (Alberts et al. 1990; Boudou et al. 1991). The

rooted macrophyte Elodea densa accumulated different concentrations of

methylmercury versus inorganic mercury from sediment (uptake was 40 times

higher for methylmercury). Tissue concentrations were similar throughout the

plant when the mercury source was water (ratio of 1.5 methylmercury to inorganic;

Boudou et al. 1991).

Chloride concentration and pH influenced uptake of inorganic mercury by a marine

diatom. Rates were low in seawater and low-chloride freshwater with neutral pH

(Mason et al. 1996). Methylmercury uptake rates were high in high-chloride waters

and were not influenced by pH. The uptake rate of methylmercury only became

limited when very low chloride concentrations decreased the concentration of

CH3HgCl. Elemental and dimethylmercury were not significantly accumulated.

Exposure Pathways

Aquatic organisms can accumulate mercury from water (including pore water) and

food sources (including sediment). Quantity accumulated is a function of the

exposure pathway and the physical and environmental factors such as temperature,

pH, salinity, total organic carbon, and sulfides. If conditions are favorable for

methylation, organisms can accumulate high concentrations of mercury even with

low concentrations in the water and sediment.

Water

Phytoplankton, invertebrates, fish (including eggs and larvae), and mammals take

up inorganic and organic mercury from the water column (McKim et al. 1976;

Pentreath 1976a; 1976b). In phytoplankton, algae, and microorganisms, mercury

uptake is primarily a passive process that occurs by adsorption to the cell surface

either through interaction with functional groups in the cell wall or through

sorptive properties associated with the extracellular matrices (Darnell et al. 1986;

Gadd 1988). Passive diffusion of lipid-soluble species (uncharged chloride

17
complexes) is responsible for mercury uptake in a marine diatom (Mason et al.

1996). Uptake in phytoplankton and aquatic plants has been correlated with the

concentration of mercury in the water (Windom and Kendall 1979; Lenka et al.

1990). Water is an important exposure pathway for mercury uptake by lower

organisms and thus into the food web (Francesconi and Lenanton 1992). Dissolved

mercury concentrations in water are typically very low; the major increase in

mercury concentrations occurs between water and phytoplankton of about a factor

of 105 to 106 (Mason et al. 1995b). In contrast to microorganisms, uptake is

primarily an active process for fish and invertebrates, and is related to respiration

rate and metabolic rate (Rodgers and Beamish 1981). Uptake of methylmercuric

chloride in water by different tissues of brook trout was found to be directly related

to the water concentration of the mercury (McKim et al. 1976).

BCFs are the concentration of mercury in tissue divided by the concentration in the

exposure water. They have been calculated from laboratory experiments for many

species of aquatic organisms to estimate uptake from water. However, BCFs have

limited use for several reasons. First, BCFs reported in the literature most likely

underestimate actual values because laboratory studies were done before the use of

trace-metal free protocols and used higher water concentrations than found in the

field (Zillioux et al. 1993). More recent BCF calculations for mercury have yielded

values one to two orders of magnitude higher than previous estimates. Second,

BCFs only reflect uptake of a contaminant from the water. Higher trophic species

accumulate mercury primarily through the food web. Reported BCFs for mercury

vary considerably due to differences between species, exposure concentration, and

duration. Further, BCFs for the same species may be several orders of magnitude

higher for methylmercury than for inorganic mercury.

Brook trout exposed to varying concentrations of methylmercury for 28 to 38

weeks had bioconcentration factors ranging from 69,000 to 630,000 (McKim et al.

1976). The wide range of BCFs reported for brook trout are related to the tissue

analyzed. Bioconcentration factors for muscle tissue in brook trout were higher at

lower water concentrations of mercury. BCFs for other tissues remained the same

when water methylmercury concentrations were varied.

18
Bioconcentration factors for inorganic mercury (mercuric chloride) in saltwater

species were 129 for adult lobster (Homarus americanus), 1,000 for mussels, and

10,000 for oysters (Crassostrea virginica; Kopfler 1974; Roesijadi et al. 1981).

Bioaccumulation factors (BAFs) were calculated from field studies for yearling

yellow perch from five freshwater lakes. These factors ranged from 106-107 for

methylmercury and more than 104 for other mercury species (Bloom 1992).

Sediment

Sediment is an important exposure pathway for all forms of mercury to aquatic

organisms. High concentrations of organic substances and reduced sulfur that

complex free Hg[II] ions in sediment can reduce the availability of mercury to biota

(Luoma 1977; Rubinstein et al. 1983). Correlating mercury concentrations in

sediment with concentrations in biota may be difficult, particularly for highertrophic-level species.

The bioavailability of total mercury to benthic invertebrates was reported to be

inversely correlated to the organic content of the sediment (Langston 1982, 1986).

Normalizing sediment mercury concentrations to percent organic matter improved

the correlation between total mercury concentrations in sediment and invertebrate

species (including gastropods, polychaetes, and deposit- and suspension-feeding

bivalves) in a marine environment (Bryan and Langston 1992). Good sedimenttissue correlations for mercury have been found in amphipods from a freshwater

lake (Becker et al. 1993). Breteler et al. (1981) studied mercury uptake by plants

and invertebrates from several types of sediments in salt marsh environments.

Concentrations of total mercury in mussels, fiddler crabs, and Spartina alterniflora

increased as organic matter in sediments decreased.

Many investigators report no correlation between sediment and tissue

concentrations of mercury for higher-trophic-level species (Nishimura and Kumagai

1983; Jackson 1988; Rada et al. 1989b; Lindqvist 1991; Duckerschein et al. 1992).

Organic carbon normalization of sediment concentrations did not improve the

correlations for pike, a high trophic level species (Lindqvist 1991). The difficulty in

correlating mercury in sediment with mercury in organisms reflects the complexity

of variables that affect both the methylation of mercury in surface sediments and

the transfer of mercury between trophic levels. Since methylation occurs primarily

19
in surface sediments, the physical factors that affect the rate of methylation (and

demethylation) also affect the availability of mercury for uptake by organisms.

Sediment total-mercury concentrations alone may not provide information on the

exposure potential of resident organisms.

Food web

Though sediment may be the ultimate source of mercury for many higher trophic

species, the food web is the primary pathway to most organisms (Lindqvist 1991;

Bryan and Langston 1992). Most of the differentiation between inorganic and

methylmercury accumulation occurs during trophic transfer (Mason et al. 1995b)

because of the differences in assimilation of the different mercury forms and how

efficiently the different forms are transferred to predators.

Mason et al. (1995b) detected an assimilation efficiency four times greater for

methylmercury compared to inorganic mercury from phytoplankton to

zooplankton, and ten times greater between phytoplankton and planktivorous fish.

The transfer efficiency of methylmercury over inorganic mercury in zooplankton

was attributed to mercury partitioning in the algal cell. Methylmercury

accumulated in the algal cytoplasm, which zooplankton digest, with 62% of the

methylmercury transferred, while inorganic mercury was primarily bound to thiols

in the algal cell membrane. Therefore, a smaller percentage (15%) of inorganic

mercury was transferred to zooplankton.

As methylmercury increases in prey items, the transfer efficiency also increases

(Windom and Kendall 1979). Since methylmercury concentrations are highest in

fish, piscivorous fish will be exposed to higher concentrations of methylmercury

than fish that feed on invertebrates. For example, walleye accumulated mercury at

a faster rate and at higher concentrations than pike from the same freshwater lake

(Mathers and Johansen 1985). A high proportion of the diet of walleye was smelt,

the most contaminated prey item, whereas pike ate only a small proportion of this

prey item. Dietary changes during life history development, or due to season or

habitat differences can change exposure. Dietary shifts in prey items of similar

trophic levels but from different habitats, or dietary shifts due to a different size

structure of prey, can also affect the mercury concentrations in top-level predators

(Lindqvist 1991).

20
The relative importance of dietary versus aqueous mercury uptake pathways is

unclear. Probably less than 10% of the mercury in fish tissue residues is obtained

by direct (gill) uptake from water (Francesconi and Lenanton 1992; Spry and Wiener

1991). Methylmercury concentrations used in laboratory studies of aqueous uptake

are 1,000 to 10,000 times the ambient concentration of methylmercury in natural

water (Spry and Wiener 1991), thereby overestimating the significance of direct

aqueous uptake. The proportion of mercury taken up from dietary sources versus

water in invertebrates has not been estimated. Suspension-feeding bivalves may

principally accumulate mercury by consuming algal cells (Riisgård and Hansen

1990).

Although mercury correlations are complicated by the importance of the foodchain exposure pathway, mercury concentrations in predators and prey have been

correlated (e.g., Allard and Stokes 1989; Lindqvist 1991; Spry and Wiener 1991). For

example, mercury concentrations in smallmouth bass from Ontario lakes were

directly correlated with mercury in crayfish, which comprised 60% of their diet.

Detritus can be a very important source of mercury, particularly in estuarine

habitats. Organic detritus from Spartina alterniflora may contain 30 times more

mercury than plankton. Organisms in detritus-based food webs are thus exposed to

higher mercury concentrations than are animals feeding on plankton (Lindberg and

Harriss 1974). Mercury associated with humic matter in lakes is fed upon by

bacteria and zooplankton, which incorporate mercury into the detrital food web

(Lindqvist 1991).

Mercury in the fecal matter of marine mammals can also be a significant source to

other aquatic organisms near breeding colonies or haul-out areas (Eisler 1987).

Biological Factors Affecting Accumulation of Mercury

The primary biological factors governing the accumulation of mercury include age,

weight, and diet. Differences in accumulation between the sexes have been

attributed to differences in diet.

21
Fish

Numerous field studies have shown that the concentration of total mercury in fish

positively correlates with length, age, and weight (Hall et al. 1976a, b; Huckabee et

al. 1979; Rada et al. 1986; Møhlenberg and Riisgård 1988; Greib et al. 1990; Leah et

al. 1992). However, total mercury concentrations may not always correlate with

size due to differences associated with diet, residence time in a contaminated

habitat, and type of mercury (Francesconi and Lenanton 1992). The percentage of

methylmercury increases with age in both fish and invertebrates (Møhlenberg and

Riisgård 1988; Riisgård and Hansen 1990).

In some species of fish and invertebrates, sex differences in mercury tissue

concentrations have been reported. For example, total mercury concentrations in

the muscle tissue of freshwater sunfish were greater in females than males at ages 2

to 3 (Nicoletto and Hendricks 1987). This may be due to increased food demands

for females related to reproduction. In contrast, there was little relationship

between sex and bioaccumulation of mercury in three species of fish (roach, perch,

and pike) collected from Swedish lakes (Lindqvist 1991).

Bloom (1992) did not find a relationship between lipid content and methylmercury

concentrations in a variety of fresh- and saltwater fish.

Invertebrates

Cockles (Cardium edule) from a polluted estuary were found to have a positive

linear correlation between their age and the percentage of organic mercury in their

tissues (Møhlenberg and Riisgård 1988). Organic mercury comprised 30% of the

total mercury in two-year old cockles; 60% in three-year olds; and 90% in four-year

olds. This relationship was attributed to the rapid loss of inorganic mercury and

continued uptake of organic mercury over time. However, the correlation was not

as strong when weight was used instead of age due to variations in growth rate at

different locations. Total mercury concentrations in mussels were found to be

higher in 27 mm than 31-mm sized individuals (Riisgård and Hansen 1990). This

difference was perhaps due to a decrease in both weight-specific filtration rate and

surface area-to-volume ratio in larger mussels (Fowler et al. 1978; Riisgård et al.

1985).

22
Total mercury concentrations may (Allard and Stokes 1989) or may not (Rada et al.

1986) correlate with weight or age in crustaceans.

Concentrations of mercury in male and female emergent mayflies (Hexagenia

bilineata) in the upper Mississippi River differed. The authors recommend sampling

male and female mayflies separately (Dukerschein et al. 1992).

Marine Mammals

Mercury concentrations (both organic and inorganic) are positively correlated with

body length in marine mammals (Arima and Nagakura 1979; Wagemann et al.

1983; Joiris et al. 1991). Hansen et al. (1990) found a highly significant correlation

between age and tissue content of mercury in whales from West Greenland. This

correlation has been used to separate immature specimens from adults. Joiris et al.

(1991) found that the concentration of methylmercury in muscle and liver tissue in

harbor porpoises did not increase with increasing length as strongly as did total

mercury.

Leonzio et al. (1992) suggest that the elevated concentrations of inorganic mercury

measured in mammals, as compared to fish, may be related to differences in

respiratory systems. In contrast to fish, where the gills allow contaminants to be

lost to the environment because blood flow has contact with the water, the

mammalian respiratory system does not have a similar exchange. Mammals have

developed different defense mechanisms. For example, selenium combines with

mercury to form the non-toxic compound tiemannite that is stored within cells.

The processes of intracellular storage tend to increase concentrations of the metal

in certain organs while reducing the toxicity. In marine mammals, intracellular

storage of mercury occurs as complexes of both selenium and metallothioneins

(MTs).

Other Factors Affecting Accumulation

Temperature and season influence the availability and accumulation of mercury in

addition to the factors already discussed. Changes in temperature can affect

mercury concentrations in organisms either directly by affecting metabolic rate and

thereby exposure, or indirectly by influencing the methylation of mercury and

therefore enhancing availability. Rates of methyl- or inorganic mercury uptake

23
increase with increasing aqueous concentrations and/or increasing temperature in

the water for some species (e.g., phytoplankton, gastropods, fish; Windom and

Kendall 1979; Rodgers and Beamish 1981; Tessier et al. 1994). A rise in

temperature (and a corresponding rise in respiratory volume) can increase the rate

of uptake via the gills (EPA 1985).

Total concentrations of mercury in killifish from an estuarine wetland were five

times higher in spring and summer than in other seasons (Weis et al. 1986),

presumably due to higher methylation rates in summer. Zooplankton mercury

concentrations peaked in June in Swedish lakes and fish tissue levels varied by a

factor of two, reaching a maximum in spring (Lindqvist 1991). Mercury content of

mussels from the Gulf of St. Lawrence estuary varied seasonally by a factor of two

(Cossa and Rondeau 1985).

The relationship of pH, conductivity, and salinity to mercury accumulation is not

well understood. Elevated mercury concentrations have frequently been found in

piscivorous fish in poorly buffered (alkalinity < 55 μeq/l and calcium < 2 mg/l), lowpH lakes (pH 6.0-6.5) in areas removed from industrial inputs of mercury (Rada et

al. 1989a; Winfrey and Rudd 1990; Spry and Wiener 1991). Total mercury

concentrations in yellow perch were inversely correlated with pH in ten Wisconsin

lakes (Cope et al. 1990). Mercury concentrations in zooplankton in Swedish lakes

were correlated with pH but the relative importance of this correlation changed

over time (Lindqvist 1991).

In freshwater lakes removed from direct sources of mercury, conductivity explained

54% of the variability in mercury concentrations in crayfish (Allard and Stokes

1989). Conductivity was also highly correlated with calcium, magnesium,

alkalinity, pH, and sodium. This correlation suggests that the buffering capacity of

the lake was an important influence on crayfish accumulation of mercury. Low

calcium ion concentrations enhanced the efficiency of methylmercury uptake

across the gills of rainbow trout (Rogers and Beamish 1983).

24
TOXICITY OF MERCURY

The toxicity of mercury to aquatic organisms is affected by both abiotic and biotic

factors including the form of mercury (inorganic versus organic), environmental

conditions (e.g., temperature, salinity, and pH), the sensitivity of individual species

and life history stages, and the tolerance of individual organisms. Toxicological

effects include neurological damage, reproductive impairment, growth inhibition,

developmental abnormalities, and altered behavioral responses. Wiener and Spry

(1996) concluded that neurotoxicity seems to be the most probable chronic

response of wild adult fishes to methylmercury exposure, based on observed effects

such as uncoordination, inability to feed, diminished responsiveness, abnormal

movements, lethargy, and brain lesions. In laboratory studies, reproductive

endpoints are generally more sensitive than growth or survival, with embryos and

the early developmental stages being the most sensitive (Hansen 1989). Impaired

reproduction in sensitive aquatic organisms has been shown to occur at aqueous

concentrations of mercury between 0.03 and 1.6 μg/l (Eisler 1987). Long-term

mercury exposure to adult fish also has been shown to result in retarded growth of

offspring (Snarski and Olson 1982) and teratogenic effects (Weis et al. 1981).

Chronic exposure to low concentrations of mercury may result in populations that

become tolerant to the toxic effects of mercury contamination (Weis and Weis

1989).

The toxic concentration of mercury compounds can vary by an order of magnitude

or more depending on the exposure condition. For example, toxicity is greater at

elevated temperatures (Armstrong 1979), at lower oxygen content (Sloof et al. 1991),

and at reduced salinities in marine environments (McKenney and Costlow 1981).

Site-specific factors (such as TOC) affect the bioavailability and toxicity of mercurycontaminated sediment (Langston 1990). Even though correlations exist between

toxicological observations and sediment pollution gradients, Langston (1990)

recommends collecting site-specific data because biological responses can not

always be satisfactorily predicted from chemical data or modeling results.

The sensitivity of aquatic organisms to either inorganic or methylmercury varies

considerably between species — more than the difference in sensitivity of a

particular species to various mercury compounds (EPA 1985). Methylmercury is

more acutely toxic to aquatic organisms than inorganic mercury, but the range

25
among different species in sensitivity to either compound is quite large. For

example, the concentration of inorganic mercury inducing acute toxicity was

observed to range over almost three orders of magnitude from 0.1 μg/l to more

than 200 μg/l when results from tests with different species were compared (Eisler

1987). Tests on the same freshwater species with both inorganic and

methylmercury showed that methylmercury was more than 30 times more acutely

toxic than inorganic mercury (EPA 1985).

The general mechanism of action for toxic effects for inorganic mercury which has

the form Hg (II), the divalent mercury cation, is believed to be the high affinity for

thiol or sulfhydryl groups of proteins (Clarkson 1972; Hughes 1957, Passow et al.

1961) resulting in altered protein production or synthesis (Syversen 1977).

Methylmercury is lipid soluble, allowing rapid penetration of the blood-brain

barrier (Feltier et al. 1972, Giblin and Massaro 1973; McKim et al. 1976; Olson et al.

1978; Beijer and Jernelov 1979). Injury to the central nervous system results from

accumulation of methylmercury in the cerebellum and cerebral cortex where it

binds tightly to sulfhydryl groups resulting in pathological changes (Sastry and

Sharma 1980). Inside the cell, methylmercury inhibits protein synthesis/RNA

synthesis (Yoshino et al. 1966; Chang et al. 1972).

Zillioux et al. (1993) suggest that, prior to the mid-1980s, few data are available on

the biological effects of mercury at environmental concentrations because

laboratory studies used exposure concentrations that were much higher than actual

concentrations in the field. This was in part due to contamination during sample

collection and analysis. Improvements in trace-metal-free clean protocols during

sample collection, handling, and processing as well as lower analytical detection

limits have resulted in lower environmental concentrations of mercury and lower

concentrations reported to elicit adverse effects. Although pre-1980 data are useful

in identifying modes of effect and relative toxicity of the various mercury

compounds, these data should be used with caution.

The following sections review available literature on the toxicity of mercury in

water, toxicity in sediment, and toxicity associated with mercury in tissues. A wide

range of toxic concentrations have been reported.

26
Toxicity of Mercury in Water

Nearly all of the studies evaluating the toxicity of mercury compounds where the

route of exposure is through water have been conducted under laboratory

conditions. Due to the nature of laboratory studies and differences in experimental

design and technique, a wide range of toxic concentrations have been reported for

a given species (Table 1). For example, toxicity tests using flow-through systems

generally show higher toxicity at lower concentrations than static-renewal systems

using the same (nominal) concentrations and the same species. This difference is

probably due to loss of mercury from the test container in the static-renewal tests

(Birge et al. 1979; Biesinger et al. 1982; WHO 1989).

Fish

Fish tend to be more sensitive to sublethal effects from chronic exposure to both

inorganic and organic mercury than invertebrates, but fish are less sensitive to

acute effects (EPA 1985; Hansen 1989). The early life stages of fish are generally

the most sensitive to mercury. Birge et al. (1979) conducted several tests designed

to evaluate embryo survival, hatching success, teratogenic effects, and the effects of

mercury on six species of freshwater fish. The sensitivity of the embryo-larval stage

for various species was correlated with the length of time for eggs to develop and

hatch and the duration of exposure. Trout eggs treated in a flow-through system

experienced approximately 40% mortality after a five-day exposure and 100%

mortality after an eight-day exposure to an average mercury concentration of

0.12 μg/l.

Birge et al. (1979) also evaluated the long-term effects of mercury exposure on fish

reproduction by conducting chronic bioassays with rainbow trout. Their results

suggest that exposure of adult fish to mercury can have significant adverse effects

on their offspring, with the effects enhanced if the embryos are also reared in a

mercury-contaminated environment. Their data show a dose-dependent response

in both bioaccumulation of mercury in gonadal tissues and toxic effects on

embryos. Short-term exposures of embryos to high concentrations of mercury can

also elicit significant adverse effects (Sharp and Neff 1980) and such exposures

should be taken into account in the evaluation of potential long-term impacts to

receiving environments. Although time-integrated concentrations may be within

accepted guidelines, a short-term exposure to an elevated mercury concentration

27
Table 1. Toxicity resulting from exposure to mercury in water.

Fish Species Hg (μg/l) Hg Form Exposure Effect(s) ReferenceBrachydanio rerioZebrafish(adults & fertilized eggs)

0.1, 1, 5,

10, & 20

phenylmercuricacetate

25 days 83% spawning inhibition at 5 ug/l;

99% inhibition at 20 ug/l; significantdecrease in hatching frequency at0.2 and 1 ug/l.

Kihlstrom et al. 1971

Brachydanio rerioZebrafish(fertilized eggs)

10, 20 & 50 phenylmercuric

acetate

~ 6 days increased hatching success at

10 ug/l; increased time to hatch andreduced hatching success at20 ug/l; no hatch at 50 ug/l

Kihlstrom and Hulth1972

Poecilia reticulataMedaka guppies

1.8 methylmercury 3 months impaired spermatogenesis Wester 1991

Oryzias latipesMedaka(fertilized eggs)

15

20 & 30

mercuric chloride 16 days 80% egg mortality; cardio-vascular

abnormalities in eggs(hemorrhaging, blood vesseldeterioration, loss of circulatingblood cells and blood vessels),teratogenic effects in fry (i.e., nonfunctioning caudal fins, skeletaldefects, spinal curvature)100 % egg mortality

Heisinger and Green1975

Fundulus heteroclitusKillifish(fertilized eggs)

0, 4, 10, 20, 30,

40, 60, 80 &

100

mercuric chloride 32 days decreased hatching at [Hg] >10 ug/l;

lateral curvature of the spine inlarvae at [Hg] 30 ug/l.

Sharp and Neff 1980

Fundulus heteroclitusKillifish(fertilized eggs)

30 & 40 methylmercuric

chloride

up to 48days

reduced axis formation,development of cyclopia, defectivecardiovascular system, skeletalmalformations

Weis and Weis 1977

Pimephales promelasFathead Minnow

0.26

1.022.01

mercuric chloride 41 days decreased growth in females and

reduction in number of spawningpairscessation of spawningsevere stunting and scoliosis

Snarski and Olson 1982

Oncorhynchus mykisRainbow trout(fertilized eggs)

0.12 mercuric chloride 5 & 8 days 40% mortality in fertilized eggs;

100% mortality after 8 days

Birge et al. 1979
Table 1. continued

Molluscs/Bivalves Species Hg (μg/l) Hg Form Exposure Effect(s) ReferenceCrepidula fornicataLimpet

>0.25

0.42

1.0

mercuric chloride 16 weeks growth impairment, diminished

conditionsuppressed fecundity, reductions inlarval settlementsignificant growth reductions

Thain 1984

Mercenaria mercenariaClams

15 mercuric chloride 8 & 10

days

reduced growth Calabrese et al. 1977

Crassostrea virginicaOysters

12 reduced growth

Mytilus edulisMussels

0.3

>1.6

mercuric chloride 10-22 days

>3 days

reduced growthcessation of growth

Stromgren 1982

Echinoderm SpeciesAnthocidaris crassispinaSea urchin

10 mercuric chloride Not

available

fertilization and developmentinterference

Kobayashi 1984

Crustacean SpeciesCallinectes sapidusBlue crab (megalopae through2nd crab stage)

10 to 20 mercuric chloride 22 days reduced survival in megalopae McKenney and Costlow

1981

Mysidopsis bahiaMysids

1.6

2.5

mercuric chloride full life

cycle

delayed sexual maturation, delayedbrood release, decreased broodproductionincreased brood development time;aborted developing juveniles

Gentile et al. 1983
could result in inhibition of hatching, teratogenic development, and possible

population effects.

Low concentrations of mercury in freshwater reportedly result in olfactory and

chemoreceptor impairment in salmonids and other fish, which may interfere with

normal migratory behavior (Hara et al. 1976; Rehnberg and Schreck 1986). For

example, Hara et al. (1976) reported that rainbow trout exposed to inorganic

mercury concentrations as low as 0.1 mg/l for two hours showed reduced olfactory

responses. Further physiological and behavioral studies by Rehnberg and Schreck

(1986) showed that mercury exposure reduced the ability of coho salmon to detect

natural odors and disrupted simple upstream movement in laboratory experiments.

Weis and Weis (1989) suggest that prior exposure to mercury may produce

populations that are more tolerant to the toxic effects of mercury contamination.

Differences in tolerance to the effects of methylmercury were observed between

organisms from mercury-contaminated and clean environments. Eggs collected

from killifish in a contaminated area were mostly resistant to the teratogenic effects

of methylmercury, while eggs of fish from the clean area showed a range of

sensitivity. The susceptible eggs from the clean area also accumulated higher levels

of mercury than did the eggs from the contaminated area (Weis et al. 1981; 1982).

Offspring from fish previously exposed to mercury contamination were more

tolerant to environmental mercury concentrations than offspring from clean

environments (Weis and Weis 1984).

The situation is complicated by the fact that some fish that build up a tolerance to

low concentrations of mercury can also detoxify the free metal within cells via the

production of metallothioneins (MTs) and other metal-binding proteins. Brown et

al. (1983) proposed that toxic effects occur as the binding capacity of MT becomes

saturated, due to the interaction of excess free metal in the cell with the enzyme

pool.

Invertebrates

Calabrese et al. (1977) suggest that marine bivalves embryos are more sensitive

than the larvae in their susceptibility to mercury. They further indicated that

growth of fully-developed larvae may be retarded at concentrations too low to elicit

30
significant mortality, thus prolonging the pelagic stages and increasing the risk of

predation, disease, and dispersion. Several endpoints have been used to measure

the effect of mercury exposure on bivalves, including biomarkers. The prophyrin

precursor -aminolevulinic acid (ALA) may be useful as a biomarker of mercury

exposure in bivalves (Brock 1993).

The effects of salinity on the toxicity of mercury have been demonstrated in a study

conducted with the megalopae of the blue crab, Callinectes sapidus (McKenney and

Costlow 1981). Their results indicated that as salinity was reduced below 20 partsper-thousand, less mercury was required to produce equivalent toxicity among

megalopae. This is significant for blue crab and other estuarine species which

inhabit, migrate through, and use areas of lower salinity for foraging, spawning, and

nursery grounds. Their data imply that the impact to a given population of fish or

invertebrates is highly dependent on the life stage and surrounding environmental

conditions.

The significance of experimental design and exposure period on evaluating the

toxicity of mercury was demonstrated in a series of studies conducted by Biesinger

et al. (1982). In acute flow-through toxicity tests with Daphnia magna,

methylmercuric chloride was about 10 times more toxic than inorganic mercury,

but only about 4 times as toxic under static-renewal conditions. In the staticrenewal tests with methylmercury, it was discovered that about 90 percent of the

mercury had been converted to inorganic mercury during the testing period. In

chronic flow-through toxicity tests with Daphnia magna, methylmercuric chloride

was about 30 times more toxic than inorganic mercury.

Plants

Chronic toxicity (as demonstrated by reduced population growth) in a marine

diatom (Thalassiosira weissflogii), exposed to inorganic mercury, methylmercury,

dimethylmercury, and elemental mercury, was related to the aqueous

concentration of a single mercury species, (the chloride species HgCl2 or CH3HgCl),

not to the total mercury or free mercury ion concentration (Mason et al. 1996).

Approximately the same concentration of CH3HgCl and HgCl2 reduced growth in

the diatom by 50 percent. Mason et al. (1996) explain the apparently higher

toxicity of methylmercury compared to inorganic mercury (expressed as a total

31
concentration of all inorganic forms) observed by numerous authors as a result of

the low percentage of the chloride form (HgCl2) in the inorganic mercury fraction

of seawater (3.3 %) compared to the high percentage of the methyl mercuric

chloride species (CH3HgCl), which forms 100% of the methylmercury in seawater.

Elemental and dimethylmercury, even though more hydrophobic than HgCl2 and

CH3HgCl respectively, were neither accumulated nor toxic to the diatom. The

hypothesis by Fisher et al. (1984), that the metals that are most bioaccumulated by

phytoplankton are the most toxic, may also be true about individual mercury

species.

Toxicity of Mercury in Sediment

The complex behavior of mercury in the environment makes it difficult to predict

toxic effects based on bulk sediment total mercury concentrations. All the available

data for effects of mercury in sediment are based on measurements of either

inorganic mercury or total mercury.

The concentrations of mercury in sediment associated with toxicity are primarily

derived from field studies, in contrast to the large number of laboratory toxicity

tests for water exposure. The results from only two spiked sediment bioassays are

available. Birge et al. (1979) reported reduced survival (70% and 45%) of rainbow

trout eggs exposed to sediment contaminated with inorganic mercury (mercuric

chloride) for 20 days at concentrations of 1.05 and 0.18 mg/kg respectively. Swartz

et al. (1988) reported an LC50 of 13.1 mg/kg for the marine amphipod (Rhepoxynius

sp.).

Considerable data are available, however, from field-collected samples that include

both measurements of mercury concentrations in sediment and adverse biological

effects. Long and MacDonald (1992) reviewed the concentrations of mercury that

were associated with measures of adverse biological effects in 169 studies that

included both marine and estuarine systems. Data from those studies were used to

calculate Effects Range-Low (ERL) and Effects Range-Median (ERM) concentrations

of 0.15 mg/kg and 0.71 mg/kg, respectively. The ERL and ERM concentrations are

the lower (10 percentile) and median (50 percentile) of the study concentrations

associated with toxic effects. Of the total number of studies in the data set, 8.3%

had biological effects below the ERL (Long et al. 1994). The incidence of effects

between the ERL and ERM concentrations was 23.5%. The incidence of effects

32
above the ERM concentration was 42.3% for mercury, while for other metals (e.g.,

chromium, copper, lead and silver) the incidence of adverse effects above the ERM

was in the range of 75%. The low accuracy of the ERL and ERM mercury guidelines

in predicting adverse effects compared to these other metals highlights the need for

site-specific effects-based data for determining sediment mercury concentrations

that are a threat to aquatic biota.

The Washington State Department of Ecology uses Apparent Effects Threshold

(AET) concentrations as the basis for sediment criteria for mercury. The AETs,

based on laboratory bioassays and benthic community studies, represent the

concentration of a contaminant above which significant adverse effects were always

observed for a specific biological indicator (PTI 1988). The AETs for mercury were

empirically derived from studies conducted with contaminated and reference

marine sediments collected from Puget Sound. AETs for total mercury are

0.41 mg/kg for the MicrotoxTM bacterial luminescence bioassay, 0.59 mg/kg for

oyster larvae abnormality, 2.1 mg/kg for amphipod (Rhepoxynius abronius) lethality,

and 2.1 mg/kg for reductions in the abundance of major taxa of benthic

macroinvertebrates.

A laboratory study by McGreer (1979) demonstrated that clams, Macoma balthica,

avoided burrowing into field-collected sediment containing a suite of metals. The

concentrations of both cadmium (1.4 ppm) and mercury (0.46 ppm) best explained

the behavioral responses. Avoidance of mercury-contaminated habitats by aquatic

species may be important ecologically. Inhibited burrowing response, relocation,

and lack of larval settlement can decrease population sizes and reduce overall

community composition. Species that avoid contact with contaminated sediment

and do not burrow into the sediments are more vulnerable to predators and adverse

environmental conditions (e.g., temperature extremes, wave action, and

contaminants in the water column).

Toxicity Associated with Mercury in Tissues

Few studies report both tissue residues and effects in either short- or long-term

exposure to low concentrations of mercury (Table 2). It is important to stress that

both the tissue concentration and the exposure time and route (i.e. water, food,

33
maternal transfer) are critical factors in producing toxic symptoms in aquatic

receptors.

According to Wiener and Spry (1996), mercury transferred from the female to the

eggs during oogenesis may pose a greater risk to embryos than exposure to

mercury in the water column. For rainbow trout, mercury residues in ovaries of 0.5

mg/kg were associated with a significant reduction in larval survival and abnormal

development (Birge et al. 1979). Whitney (1991) reported that hatching success and

embryonic survival in walleye were inversely correlated with mercury

concentrations in the egg (range 0.002 to 0.058 mg/kg). However, only one of 12

samples had hatching success or embryonic survival less than 90%, and there was

no apparent dose-response relationship.

Mercury concentrations in brain tissue associated with lethal effects appear to show

less variation than that of other tissues (e.g., muscle, whole body). For example,

mercury concentrations in most types of tissues of brook trout killed by exposure

to 2.9 μg/l of mercury in the water column varied among individuals, whereas

concentrations in the brain showed little variation (McKim et al. 1976). These

results are consistent with the hypothesis that the central nervous system, rather

than muscle tissue or other organs, is the site of the most harmful toxic action in

fish exposed to mercury (Wiener and Spry 1996). In their review of the literature,

Wiener and Spry (1996) concluded that mercury concentrations of 7 mg/kg or

greater in fish brain probably cause severe, potentially lethal effects. In sensitive

species such as the walleye, brain tissue concentrations of 3 mg/kg or greater

probably indicate significant toxic effects.

Based on a review of the literature, Niimi and Kissoon (1994) suggest that a total

mercury body burden of 1-5 mg/kg represents a threshold concentration for

chronic adverse effects in aquatic organisms. Wiener and Spry (1996) reviewed the

literature and provided guidance for interpreting mercury residues in the axial

muscle tissue in adult fish associated with toxicity; both field and laboratory studies

indicate that residues of 6 to 20 mg/kg are toxic. Whole body mercury

concentrations of about 5 mg/kg in brook trout and 10 mg/kg in rainbow trout

were associated with sublethal and lethal effects. Both of these papers are recent

examples of attempts to identify a threshold of mercury in tissue that is associated

34
Table 2. Toxicity associated with mercury in tissues (μg/g) wet weight.

Fish Species Tissue Type

Tissue Hg

(μg/g, ww) Hg Form Exposure Effect(s) 

Oncorhynchus mykis

Rainbow trout

ovary 0.5 mercuric chloride 400-528

days

significant reduction in alevin survival (4-day

post hatch); significant increase in

teratogenic effects

Oncorhynchus mykis

Rainbow trout

whole body 10-30

30-35

methylmercuric

chloride

84 days decreased growth and appetite

darkened skin and lethargy

Oncorhynchus mykis

Rainbow trout

brain

liver

muscle

whole body

16-30

26-68

20-28

19

total mercury in food 270 days darkened skin; loss in appetite, visual acuity,

and growth; loss of equilibrium

Oncorhynchus mykis

Rainbow trout

muscle 12-23 methylmercuric

chloride (4-24 ppm in

food)

105 days hyperplasia of gill epithelium 

Oncorhynchus mykis

Rainbow trout

brain

liver

muscle

7-32

32-114

9-52

methylmercuric

chloride (4 μg/l in water

column)

30-98 days decreased appetite and activity, mortality 

Oncorhynchus mykis

Rainbow trout

whole body 4-27 methylmercuric

chloride (9 μg/l in water

column)

12-33 days decreased appetite and activity, mortality 

Pimephales promelas

Fathead minnow

whole body 1.4 mercuric chloride 60 days impaired reproduction, retarded larval growth 

Mugil cephalus

Striped mullet

whole body 0.3

5.0

methylmercuric

chloride (0.001 mg/l in

water column)

methylmercuric

chloride (0.01 mg/l in

water column)

7,11, 14 days

inhibition of regeneration of amputated

caudal fin

Salvelinus fontinalis

Brook trout

brain

liver

ovaries/eggs

whole body

5

8

5

3

methylmercuric

chloride (0.27 μg/l in

water column)

273 days no apparent effects 

Salvelinus fontinalis

Brook trout

brain

liver

ovaries/eggs

whole body

17

24

10

5-7

methylmercuric

chloride (0.93 μg/l in

water column)

273 days increased mortality, decreased growth,

lethargy, and deformities

Salvelinus fontinalis

Brook trout

brain

liver

ovaries/eggs

whole body

42

58

24

24

methylmercuric

chloride (2.9 μg/l in

water column)

273 days loss of appetite, muscle spasms, and

deformities; mortaility

Stizostedion vitreum

vitreum

Walleye

brain

liver

muscle

3-6

6-14

5-8

methylmercury (5-13

ppm in food)

42-63 days mortality; emaciation; loss of locomotion,

coordination and appetite.

Stizostedion vitreum

vitreum

Walleye

brain

liver

muscle

15-40

18-50

15-45

methylmercury (5-13

ppm in food)

240-314

days

higher mortality; emaciation; poorer

locomotion, coordination and appetite.
with adverse effects. The "thresholds" presented in these papers are based on

effects in adult fish and probably do not represent a truly protective level for all

species and life stages, including maternal transfer. We begin to become concerned

about reproductive or early life stage effects when total Hg in whole bodies of fish

are between 0.5 and 1.0 ppm.

Interactions with Other Metals

The effects on aquatic organisms due to interactions of mercury with cadmium,

copper, selenium, and zinc were found to be dependent on exposure concentrations

(Birge et al. 1979). In general, effects were less than additive at lower exposure

concentrations and greater than additive (synergistic) at higher concentrations.

Zinc and cadmium were reported to reduce the teratogenic effects of

methylmercury to killifish (Weis et al. 1981). Cadmium added to methylmercury

reduced the retardation effect on fin regeneration in mullet (Weis and Weis 1978).

The percentage of embryos affected and degree of malformation observed due to

exposure of killifish eggs to 20-50 μg/l methylmercury was reduced when cadmium

or zinc was added. Selenium was reported to reduce the developmental effects of

inorganic mercury to embryos of the medaka (Japanese ricefish), but only after the

formation of the embryonic liver (Bowers et al. 1980). Interactions between

inorganic mercury and zinc, PCBs, and a PAH (fluoranthene) were observed to be

generally additive in sediment exposure to a marine amphipod (Swartz et al. 1988).

A mixture of an inorganic form of mercury (mercuric chloride) and the chlorides of

zinc and lead had a synergistic toxic effect on the water exposure of a marine ciliate

Uronema marinum (Parker 1979).

CRITERIA AND GUIDELINES

This section briefly discusses EPA’s AWQC for mercury in freshwater and marine

systems and various guidelines that have been proposed for evaluating the

potential toxicity of mercury in sediments.

Ambient Water Quality Criteria (AWQC)

The AWQC, promulgated by the U.S. EPA, are intended neither as rules nor

regulations, but present data and guidance on the effects of pollutants that can be

36
used to derive regulations based on considerations of water quality impacts (EPA

1993). The AWQC consist of two concentrations: the Criterion Maximum

Concentration (usually referred to as the acute AWQC) and the Criterion

Continuous Concentration (usually referred to as the chronic AWQC). The acute

AWQC are derived from short-term toxicity tests using statistical methods that

estimate the LC50 concentrations for the lowest 5 percent of the most sensitive

species tested. Acute AWQC for mercury are 2.4 and 2.1 μg/l for freshwater and

marine organisms, respectively. They are based on inorganic mercury because it is

the predominant form of mercury released into the environment (EPA 1985).

The chronic AWQC are defined by EPA as the lowest (most protective)

concentrations from three categories of tests: the final chronic value, derived from

chronic toxicity tests with animals; the final plant value, derived from toxicity tests

using aquatic plants; and the final residue value (FRV), derived from maximum

permissible tissue concentrations (for protection of human health) and

bioconcentration factors (BCF). For mercury, the chronic AWQC of 0.012 μg/L for

freshwater species and 0.025 μg/l for marine species represent FRVs, which are

based on the Food and Drug Administration’s (FDA) action level of 1 mg/kg and

BCFs for methylmercury.

The use of the FDA action level to derive the chronic AWQC assumes that aquatic

organisms would not be adversely affected by methylmercury tissue concentrations

greater than or equal to 1 mg/kg. This is based on a study in which long-term

exposure of brook trout to methylmercury resulted in tissue residue concentrations

greater than 1 mg/kg but no significant effects on survival, growth, or reproduction

(McKim et al. 1976). However, other studies have demonstrated that tissue

concentrations close to the FDA action level may elicit adverse effects in some

species (Birge et al. 1979; Snarski and Olson 1982).

Eisler (1987), in his review of the hazards of mercury to aquatic organisms, noted

that the AWQC do not appear to be protective of aquatic organisms when

compared to results from toxicity tests with sensitive species. Assuming the BCFs

accurately reflect bioaccumulation under field conditions, the FDA action level

would be the average concentration in muscle tissue of exposed aquatic organisms.

Thus, many aquatic organisms exposed to mercury concentrations equivalent to

37
the chronic AWQC would be expected to have tissue concentrations above 1 mg/kg

(EPA 1985).

Sediment Guidelines

No sediment criteria are currently available for either methylmercury or total

mercury. Several approaches have been proposed for developing guidelines for

screening contaminated sediments, but two of the more frequently used approaches

are the National Status and Trends Effects Range approach (Long and MacDonald

1992) and the AET (PTI 1988) approach developed for screening sediments in Puget

Sound, Washington. The State of Washington used the AET approach as the basis

for marine sediment management standards in Puget Sound.

The ERL and ERM concentrations for mercury in marine and estuarine sediments

are 0.15 and 0.71 mg/kg (Long and MacDonald 1992), respectively, while the AET

concentrations for mercury range between 0.41 and 2.1 mg/kg (PTI 1988).

Tissue

No standards that would be protective of aquatic organisms have been established

for mercury concentrations in tissues. The current FDA action level for the

protection of human health, based only on methylmercury in the edible flesh of

fish and shellfish, is 1 mg/kg (U. S. FDA 1984).

APPLICATIONS

Ecological assessments of hazardous waste sites with elevated concentrations of

mercury in the aquatic environment are particularly challenging due to the

complexity of the factors that affect the availability of mercury to aquatic

organisms. The distribution of total mercury in sediment, which in most cases is

predominantly inorganic mercury, may not provide sufficient information about the

bioavailability and toxicity of mercury to aquatic species. Because of the

importance of methylation in determining the availability of mercury to aquatic

organisms, the sampling design, evaluation of remedial alternatives, and monitoring

program should be based on an understanding of the system-specific processes that

lead to increased methylation and the pathways to resources of concern. The effort

required and detail of this understanding should be determined by the magnitude

of the problem and the scope of the project. This section discusses some special

38
considerations to assist in sampling, risk assessment, monitoring, and remedial

decisions.

Sampling and Monitoring Considerations

Target Species and Analysis

Mercury concentrations in resident aquatic organisms may provide the best

measure of the availability of mercury in a particular area, both because of potential

human health concerns and because it is the best indicator of availability of

mercury under the specific conditions present at a site. In selecting target species,

the trophic level, size, age, sex, life habit, metabolism, and life span of organisms are

all important factors to consider. Higher trophic-level fish species are useful for

determining whether a problem exists since mercury biomagnifies, and for longterm monitoring, since mercury concentrations are slower to decrease. However,

even fish occupying the same trophic level, with similar diets and feeding habits,

may exhibit different temporal patterns of mercury accumulation due to differences

in habitat preferences, behavior, and metabolic rate causing different exposures

(Jackson 1991). Mercury concentrations in biota may not correlate with sediment

mercury concentrations. Correlations between mercury concentrations in predator

and prey species may be useful in determining the food web pathways that connect

the mercury in the sediment to the biota.

Whole body analyses of fish are typically done to determine food chain exposure,

while fish fillets are typically analyzed to assess human health exposure. Wholebody mercury concentrations may be less than the concentration in the fillet;

however the difference may not be statistically significant. For example, although

methylmercury concentrations were higher in fillet than in whole body samples

measured in four fish species (white perch, small mouth bass, bluegill, and gizzard

shad) in Onondaga Lake (New York), only concentrations in bluegill were

statistically different (Becker and Bigham 1995).

Invertebrates such as bivalve molluscs and mayflies (Hexagenia sp.) have also been

used for assessing the availability of mercury in a particular location. Depending

on the scope of the project, monitoring several different species from different

trophic levels may be appropriate.

39
In developing assessment endpoints and sampling objectives, the potential for

direct toxicity to aquatic organisms should not be overlooked. Laboratory and insitu toxicity testing are useful approaches for assessing the direct biological effects

of elevated mercury concentrations to aquatic organisms in sediment and water.

Toxicity testing at mercury sites should include standard toxicity tests. In addition,

early life-stage tests (exposure of test species from post-fertilization through

embryonic, larval, and early juvenile development) or partial life-cycle tests (early

juvenile through post-hatch of next generation with measurements of survival,

growth, and reproductive endpoints) are sensitive tests for mercury toxicity.

Monitoring changes in abundance and diversity in macrobenthic community

composition may also provide useful information in assessing the toxicity of

mercury in aquatic habitats.

Environmental Sampling

Investigations should be designed to include both spatial and temporal sampling.

Seasonal and spatial variations in mercury concentrations, including its forms and

partitioning, within a single waterbody can be significant (e.g., Gill and Bruland

1990; Parks et al. 1989). Mercury contents in mussels from different parts of the

Gulf of St. Lawrence estuary (normalized by shell length and soft-tissue dry weight)

were highest in areas with the greatest freshwater influence and lower in regions

where the marine influence was greatest (Cossa and Rondeau 1985). Allard and

Stokes (1989) found that total mercury in two species of crayfish was significantly

higher in specimens from lake inlets than in those from the lake basin.

Determining those environmental parameters that affect the activity of methylating

microbes such as nutrients, temperature, and dissolved oxygen, and the factors

affecting the availability of inorganic mercury for methylation such as the

resuspension of sediment, TOC, and sulfides, may be warranted for the design of

sampling and monitoring plans. Data on chloride concentration and pH may be

used to determine the relative proportions of the individual inorganic and

methylmercury species and their overall partition coefficients (Mason et al. 1996).

In addition to the form of mercury, its partitioning between dissolved and

particulate forms has an important effect on both uptake and transport. In

determining the extent of contamination at a site, it is important to consider that

both resuspended contaminated sediment and dissolved mercury may act as

important sources.

40
Accurately modeling the fate of mercury in aquatic environments and the

availability of mercury to aquatic organisms requires the collection of detailed

information on the forms of mercury and their relative concentrations in different

environmental compartments (e.g., the amounts of inorganic, methyl-, and

elemental mercury in dissolved and particulate forms in the water column,

sediment [particulate and pore water], and biota). The scope of a project like this is

enormous and the effectiveness of this comprehensive approach has yet to be

demonstrated. The transfer of mercury through the food web was modeled using a

descriptive approach to explain the high levels of mercury in Lavaca Bay fish and

shellfish (Evans and Engel 1994). Tissue burdens for multiple food web

components are required. This approach was useful in identifying critical factors

responsible for localized elevations in mercury concentrations, but also

demonstrated the limitations and large effort required for modeling.

Approaches to Remediation

The level of effort needed to characterize a mercury-contaminated site may range

from simple monitoring of mercury concentrations in biota and environmental

media at small sites, to biogeochemical modeling at major mercury sites (PTI 1991).

Although methylmercury is the form of most concern in aquatic systems, it has not

been routinely measured due to the lack of standard analytical methods and cost

considerations. Determining which form of mercury to measure, and in which

media and organisms, depends on the nature of the contamination and the

objectives of the study. Measuring only total mercury concentrations in sediment

and biota may give a general picture of the extent of contamination and the

magnitude of the problem, but only provides minimal information on the fate,

transport, and availability of mercury in the system. In order to select an effective

remedial alternative, it may be necessary to characterize the major pathways to

receptors of concern and the aspects of the aquatic system that enhance

methylation and influence mercury availability.

The role of speciation in determining concentrations in, and toxicity to, biota may

need to be understood prior to attempts to control the geochemical cycling of

mercury within a waterbody. Remediation attempts have been unsuccessful at sites

where these factors have been ignored. However, this approach requires analytical

41
techniques that are selective and sensitive enough to measure ambient

concentrations of the different mercury species.

A primary goal of many remedial investigations is to establish cleanup

concentrations for mercury in various environmental media that will be protective

of both human health and the environment. Establishing target cleanup

concentrations for mercury is extremely difficult due to the many environmental

factors that influence the transformation of inorganic mercury to methylmercury.

Target cleanup concentrations should be determined on a site-by-site basis due to

the variability in the bioavailability of mercury and conditions between sites.

Determining a cleanup concentration requires knowing the effect threshold and

translating that to a sediment concentration that is protective. Cleanup

concentrations should be chosen that both reduce the source of total mercury to

the system and its bioavailability to organisms. Confirmation of the effectiveness of

the target cleanup concentration requires long-term monitoring of both sediment

and biota.

Removing hot spots (by dredging or capping) may eliminate important sources of

inorganic mercury but may not provide substantial improvement in environmental

conditions if methylation rates are much higher in less contaminated areas, such as

freshwater wetlands (St. Louis et al. 1994). Other factors affecting the site also need

to be considered. Methylation of mercury left in sediments could be increased by

dredging, increased organic loading (without sulfides), and increased thermal

loading (Rada et al. 1986).

Where source control has not been feasible due to the volume and extensive

distribution of the mercury, remediation strategies that focus on limiting the

bioavailability of mercury have been studied. Dilution approaches, such as adding

uncontaminated sediments, may reduce the supply of mercury by an order of

magnitude (Rudd and Turner 1983). Complexation of mercury to “detoxify” (e.g.,

addition of selenium to the water column) may be a valid approach. Research

indicates that selenium interferes with bioaccumulation efficiency of mercury in

fish since selenium concentrates in the fish food source and excludes mercury.

This approach can yield up to a twofold decrease in mercury bioaccumulation rates.

However, selenium toxicity may add a new problem. Liming to increase pH in

freshwater lakes has also been attempted with variable results, with positive effects

42
taking two or more years (Lindqvist 1991; Meili 1995). The long-term effectiveness

of these remedies and the potential adverse effects on the environment have not

been well-studied.

Analytical Considerations

Detection Limits in Water, Sediment, and Biota

Detection limits should be chosen based on the objective of the study. Analysis of

mercury in water samples is particularly difficult due to the very low concentrations

(parts per trillion) that need to be measured. Achieving low detection limits is

further complicated by the possibility of external contamination of samples which

can be a significant problem (Fitzgerald 1990).

The chronic AWQC for methylmercury are 0.012 and 0.025 μg/l for freshwater and

saltwater, respectively. Using the AWQC as a detection limit may be difficult as few

labs have analytical procedures that can reach these low concentrations. The

detection limit for the EPA standard contract lab program (CLP) method for analysis

of total mercury in water is 0.2 μg/l. To achieve ecologically relevant detection

limits it may be necessary to employ analytical methodologies other than those

specified under the CLP, or to modify the CLP methods. For example, mercury

detection limits to determine the mass balance of mercury in the Onondaga Lake

Superfund site were established as follows: 0.00001 μg/l for methylmercury in

water; 0.0001 μg/l for total mercury in water; and 0.01 mg/kg wet weight in fish

(PTI 1991).

These detection limits were achievable by modifying standard procedures.

Unfiltered water samples are analyzed for comparison to criteria concentrations. In

the EPA criteria document, the measurement of acid-soluble mercury (the mercury

that passes through a 45-μm filter following acidification to a pH of 1.5 to 2.0) is

recommended, although no EPA-approved protocol has been established (EPA

1985). Some recent analytical approaches for the analysis of methylmercury are

described by the following authors: Bloom and Crecelius (1983), Gill and Fitzgerald

(1987), Bloom (1989), Gill and Bruland (1990).

Choosing appropriate detection limits for sampling of sediment is particularly

important for mercury sites because even low concentrations can cause significant

43
accumulations in biota. Detection limits for sediments should be below the ERL

value of 0.15 mg/kg developed by Long and MacDonald (1992).

Quality control is an important aspect of any testing program but is particularly

important for the analysis of mercury in environmental samples. It is highly

recommended that analyses of mercury in water, sediment, and tissue also include

analyses of certified standards for the appropriate matrix as part of the quality

control plan to verify the extraction and analytical processes.

Difficulties in the extraction and analysis of mercury residues in tissues are

apparently not uncommon. For example, in three recent Superfund projects in

three different regions, methylmercury concentrations in fish tissue were reported

to be higher than total mercury concentrations. In another study of contaminant

concentrations in tissues of aquatic organisms, spike-recovery values for mercury in

tissue samples were in the range of 50 percent (Tetra Tech 1988).

The National Research Council of Canada has certified standards for

methylmercury and total mercury in animal tissue samples (e.g., dogfish liver and

dogfish muscle), and for sediment and water samples. The U.S. National Bureau of

Standards has comparable standards for total mercury.

SUMMARY

NOAA recommends a site-specific approach that focuses on determining the

availability of mercury and the potential for toxic effects. The accumulation of

mercury in aquatic biota is often the primary concern at mercury sites and is

useful for assessing availability. Bioaccumulation studies should measure tissue

concentrations in more than one resident and/or transplanted caged species,

preferably with species representing different trophic levels or different food web

pathways. It may not be possible to correlate sediment mercury concentrations

with concentrations in biota. However, correlations between mercury

concentrations in predator and prey species may be useful in determining

pathways of mercury transfer.

Toxicity tests such as the standard amphipod tests should also be conducted to

assess mercury toxicity to aquatic organisms. At major mercury sites, chronic

44
toxicity endpoints should be included in the assessment—in particular, fish early life

stage or reproductive endpoint tests. Because of the persistence of mercury in

aquatic systems, source control alone may not be sufficient to permit recovery.

Additional remedial actions may be required to reduce the total mercury burden in

the system. Long-term monitoring of tissue concentrations of mercury in aquatic

biota is needed to assess remedial effectiveness.

REFERENCES

Alberts, J.J., M.T. Price and M. Kania. 1990. Metal concentrations in tissues of

Spartina alterniflora (Loisel.) and sediments of Georgia salt marshes. Estuarine,

Coastal and Shelf Science 30: 47-58.

Allard, M. and P. Stokes. 1989. Mercury in crayfish species from thirteen Ontario

Lakes in relation to water chemistry and smallmouth bass (Micropterus dolomieui)

mercury. Canadian Journal of Fisheries and Aquatic Sciences 46.

Andersen, A., K. Julshamn, O. Ringdal, and J. Morkore. 1987. Trace Elements Intake

in the Faroe Islands, II: Intake of Mercury and Other Elements by Consumption of

Pilot Whales (Globicephalus meleanus). Science of the Total Environment 65:63-68.

Andre, J. M., A. Boudou, and F. Ribeyre, and M. Bernhard. 1991a. Comparative

study of mercury accumulation in dolphins (Stenella coeruleoalba ) from French

Atlantic and Mediterranean coasts. Sci. Total Environ. 104:191-209.

Andre, J. M., A. Boudou, and F. Ribeyre. 1991b. Mercury accumulation in

Delphinidae. Wat. Air Soil Pollut. 56:187-201.

Arima, S., and K. Nagakura. 1979. Mercury and selenium content of Odontoceti.

Bull. Japanese Soc. Scientific Fish. 45(5):623-626.

Armstrong, F. A. J. 1979. Effects of mercury compounds on fish. In: J. O. Nriagu

(Ed.), The Biogeochemistry of Mercury in the Environment. pp. 657-670. New York:

Elsevier/North-Holland Biomedical Press.

45
Battelle. 1987. Measurement of bioavailable mercury species in fresh water and

sediments. Palo Alto, California: Electric Power Research Institute.

Becker, D.S. and G.N. Bigham. 1995. Distribution of mercury in the aquatic food

web of Onondaga Lake, New York. Water, Air, and Soil Pollution 80:563-571.

Becker, D. S., G. N. Bigham, and M. H. Murphy. 1993. Distribution of mercury in a

lake food web. Poster presented at 14th annual meeting of the Society of

Environmental Toxicology and Chemistry (SETAC), Houston, Texas, November 14-

18, 1993.

Beijer, K. and A. Jernelov. 1979. Methylation of mercury in aquatic environments.

In: J. O. Nriagu (Ed.), The Biogeochemistry of Mercury in the Environment. pp. 203-

210. New York: Elsevier/North-Holland Biomedical Press.

Benes, P. and B, Havlik. 1979. Speciation of mercury in natural waters. In: J. O.

Nriagu (Ed.), The Biogeochemistry of Mercury in the Environment. pp. 175-202. New

York: Elsevier/North-Holland Biomedical Press.

Berman, M. and R. Bartha. 1986. Levels of chemical versus biological methylation

of mercury in sediments. Bulletin of Environmental Contamination and Toxicology

36:401-404.

Biesinger, K.E., L.E. Anderson, and J.G. Eaton. 1982. Chronic effects of inorganic

and organic mercury on Daphnia magna: toxicity, accumulation, and loss. Arch.

Environ. Contam. Toxicol. 11: 769-774.

Birge, W. J., J. A. Black, A. G. Westerman, and J. E. Hudson. 1979. The effects of

mercury on reproduction of fish and amphibians. In: J. O. Nriagu (ed.). The

Biogeochemistry of Mercury in the Environment. pp. 629-655. New York:

Elsevier/North-Holland Biomedical Press.

Bjerregaard, P. and L. Christensen. 1993. Accumulation of organic and inorganic

mercury from food in the tissues of Carcinus maenas: effect of waterborne

selenium. Marine Ecol. Prog. Ser. 99: 271-281.

Bjornberg, A., L. Håkanson, and L. Lundbergh. 1988. A theory on the mechanisms

regulating the bioavailability of mercury in natural waters. Environmental Pollution

49:53-61.

46
Bloom, N.S. 1989. Determination of picogram levels of methylmercury by aqueous

phase ethylation, followed by cryogenic gas chromatography with cold vapor

atomic fluorescence detection. Canadian Journal of Fisheries and Aquatic Sciences

46:1131-1140.

Bloom, N. S. 1992. On the chemical form of mercury in edible fish and marine

invertebrate tissue. Can. J. Fish. Aquat. Sci. 49:1010-1017.

Bloom, N.S., C.J. Watras, and J.P. Hurley. 1991. Impact of acidification on the

methylmercury cycle of remote seepage lakes. Water, Air, Soil Pollut. 56:477-491.

Bloom, N.S. and S.W. Effler. 1990. Seasonal variability in the mercury speciation of

Onondaga Lake (New York). Water, Air, Soil Pollut. 53:251-265.

Bloom, N.S. and E.A. Crecelius. 1983. Determination of mercury in seawater at

subnanogram per litre levels. Marine Chemistry 14:49-59.

Blum, J. M. and R. Bartha. 1980. Effect of salinity on methylation of mercury.

Bulletin of Environmental Contamination and Toxicology 25:404-408.

Bonzongo, J.J., Heim, K.J., Y. Chen, W.B. Lyons, J.J. Warwick, G.C. Miller and P.J.

Lechler. 1996. Mercury pathways in the Carson River-Lahontan Reservoir System,

Nevada, USA. Environ. Toxicol. and Chem. 15(5):677-683.

Born, E. W., I. Kraul, and T. Kristensen. 1981. Mercury, DDT, and PCB in the

Atlantic walrus (Odobenus rosmarus rosmarus) from the Thule District, North

Greenland. Arctic 34:255.

Boudou, A., M. Delnomdedieu, D. Georgescauld, F. Ribeyre, and E. Saouter. 1991.

Fundamental roles of biological barriers in mercury accumulation and transfer in

freshwater ecosystems (analysis at organism, organ, cell and molecular levels).

Water, Air, and Soil Pollution 56:807-822.

Bowers, M. A., D. Dostal, and J. F. Heisinger. 1980. Failure of selenite to protect

against mercuric chloride in early developmental stages of the Japanese ricefish

(Oryzias latipes). Comparative Biochemistry and Physiology 66C:175-178.

47
Breteler, R.J., I. Valiela, and J.M. Teal. 1981. Bioavailability of mercury in several

North-eastern U.S. Spartina ecosystems. Estuarine, Coastal and Shelf Science 12:

155-166.

Brock, V. 1993. Effects of mercury on the physiological condition and content of

the biomarker ALA in the oyster Ostrea edulis. Mar. Ecol. Prog. Series. 96:169-175.

Brown, D. A., R. W. Gosset, P. Hershelman, H. A. Schaefer, K. D. Jenkins, and E. M.

Perkins. 1983. Bioaccumulation and detoxification of contaminants in marine

organisms from Southern California coastal waters. In: D. F. Soule and D. Walsh

(Eds.), Waste Disposal in the Oceans. p. 171. Boulder: Westview Press.

Bryan, G. W. and W. J. Langston. 1992. Bioavailability, accumulation and effects of

heavy metals in sediments with special reference to United Kingdom estuaries: a

review. Environ. Pollut. 76: 89-131.

Calabrese, A., J. R. MacInnes, D. A. Nelson, and J. E. Miller. 1977. Survival and

growth of bivalve larvae under heavy metal stress. Mar. Biol. 41:179-184.

Callister, S. M. and Winfrey, M. R. 1986. Microbial methylation of mercury in upper

Wisconsin River sediments. Water, Air and Soil Pollution 29: 453-465.

Choi, S.-C., and R. Bartha. 1994. Environmental factors affecting mercury

methylation in estuarine sediments. Bull. Environ. Contam. Toxicol. 53:805-812.

Clarkson, T. W. 1972. Recent advances in toxicology of mercury with emphasis on

the alkyl mercurials. Crit. Rev. Toxicol. 203-234:.

Compeau, G. and R. Bartha. 1983. Effects of sea salt anions on the formation and

stability of methylmercury. Bulletin of Environmental Contamination and

Toxicology 31:486-493.

Compeau, G. and R. Bartha. 1985. Sulfate-reducing bacteria: principal methylators

of mercury in anoxic estuarine sediment. Appl. Environ. Microbiol. 50: 498-502.

Cope, W. G., J. G. Wiener, and R. G. Rada. 1990. Mercury accumulation in yellow

perch in Wisconsin seepage lakes: relation to lake characteristics. Environmental

Toxicology and Chemistry 9:931-940.

48
Cossa, D. and J. G. Rondeau. 1985. Seasonal, geographical and size-induced

variability in mercury content of Mytilus edulis in an estuarine environment: a reassessment of mercury pollution level in the Estuary and Gulf of St. Lawrence.

Marine Biology 88:43-49.

Craig, P.J., and P.A. Moreton. 1983. Total mercury, methyl mercury and sulphide in

River Carron sediments. Marine Pollution Bulletin 14(11):408-411.

Craig, P.J., and P.A. Moreton. 1985. The role of speciation in mercury methylation

in sediments and water. Environmental Pollution Series B 10:141-158.

Darnell, D. W., B. Greene, M. T. Henzl, J. M. Hosea, R. A. McPherson, J. Sneddon, M.

D. Alexander. 1986. Selective recovery of gold and other metal ions from an algal

biomass. Environ. Sci. Technol. 20: 206-208.

De Haan, H. 1992. Impacts of environmental changes on the biogeochemistry of

aquatic humic substances. Hydrobiologia 229:59-71.

Dukerschein, J. T., R. G. Rada, and M. T. Steingraeber. 1992. Cadmium and mercury

in emergent mayflies (Hexagenia bilineata) from the upper Mississippi River. Arch.

Environ. Contam. Toxicol. 23: 109-116.

Eganhouse, R.P. and D.R. Young. 1978. Total and organic mercury in benthic

organisms near a major submarine wastewater outfall system. Bulletin of

Environmental Contamination and Toxicology 19:758-766.

Eisler, R. 1987. Mercury hazards to fish, wildlife, and invertebrates: a synoptic review.

U.S. Fish and Wildlife Service Biology Report 85(1.10). Washington, D.C.: U.S.

Department of the Interior. 90pp.

Evans, D.W. and D.W. Engel. 1994. Mercury bioaccumulation in finfish and shellfish

from Lavaca Bay, Texas: descriptive models and annotated bibliography. NOAA

Technical Memorandum NMFS-SEFSC-348. Beaufort, North Carolina: National

Marine Fisheries Service, NOAA.

Faust, B.C. 1992. The octanol/water distribution coefficients of methylmercuric

species: the role of aqueous-phase chemical speciation. Environ. Toxicol. Chem.

11:1373-1376.

49
Feltier, J. S., E. Kahn, B. Salick, F. C. Van Natta, and M. W. Whitehouse. 1972. Ann.

Intern. Med. 76:779-792.

Fisher, N.S., M. Bohé, and J.-L. Teyssié. 1984. Accumulation and toxicity of Cd, Zn,

Ag, and Hg in four marine phytoplankters. Mar. Ecol. Prog. Ser. 18:201-213.

Fitzgerald, W.F. 1990. Mercury in seawater. In: Mercury in the Marine

Environment. Workshop proceedings. U.S. Dept. Interior, Mineral Management

Service, Anchorage, Alaska. OCS Study MMS 89-0049.

Fowler, S. W., M. Heyraud, and J. La Rosa, 1978. Factors affecting methyl and

inorganic mercury dynamics in mussels and shrimp. Marine Biology 46: 267-276.

Francesconi, K. and R. C. J. Lenanton 1992. Mercury contamination in a semienclosed marine embayment: organic and inorganic mercury content of biota, and

factors influencing mercury levels in fish. Marine Environmental Research 33: 189-

212.

Gadd, G. M. 1988. Accumulation of metals by microorganisms and algae. In:

Rehm H-J and G. Reed (eds) Biotechnology. Vol 6b, 401-433. Weinheim, Germany:

VCH Verlagsgesellschaft.

Gentile, J. H., S. M. Gentile, G. Hoffman, J. F. Heltshe, and N. Hairston, Jr. 1983. The

effects of a chronic mercury exposure on survival, reproduction and population

dynamics of Mysidopsis bahia. Environ. Toxicol. Chem. 2:61-68.

Giblin, F. J. and E. J. Massaro. 1973. Pharmacodynamics of methyl mercury in

rainbow trout (Salmo gairdneri): tissue uptake, distribution and excretion. Toxicol.

Appl. Pharmacol. 24:81-91.

Gill, G. A. and K. W. Bruland. 1990. Mercury speciation in surface freshwater

systems in California and other areas. Environmental Science and Technology 24(9):

1392-1400.

Gill, G.A. and W.F. Fitzgerald. 1987. Picomolar mercury measurements in seawater

and other materials using stannous chloride reduction and two-stage gold

amalgamation with gas phase detection. Marine Chemistry 20: 227-243.

50
Gilmour, C.C., E.A. Henry, and R. Mitchell. 1992. Sulfate stimulation of mercury

methylation in freshwater sediments. Environ. Sci. Technol. 26(11): 2281-2287.

Gilmour, C. C. and E. A. Henry. 1991. Mercury methylation in aquatic systems

affected by acid deposition. Environmental Pollution 71(2-4): 131-169.

Gilmour, C. C. and Capone, D. G. 1987. Relationship between Hg methylation and

the sulfur cycle in estuarine sediments. EOS 68:1718.

Greib, T.M., C.T. Driscoll, S. P. Gloss, C.L. Schofield, G.L. Bowie, and D.B. Porcella.

1990. Factors affecting mercury accumulation in fish in the upper Michigan

peninsula. Environmental Toxicology and Chemistry 9:919-930.

Hall, A.S., F.M. Teeny, and E.J. Gauglitz. 1976a. Mercury in fish and shellfish of the

northeast Pacific. II. Sablefish, Anoplopoma fimbria. Fishery Bulletin 74:791-797.

Hall, A.S., F.M. Teeny, L.G. Lewis, W.H. Hardman, and E.J. Gauglitz, Jr. 1976b.

Mercury in fish and shellfish of the northeast Pacific. I. Pacific Halibut,

Hippoglossus stenolepis. Fishery Bulletin 74:783-789.

Hansen, C. T., C. O. Nielsen, R. Dietz, and M. M. Hansen. 1990. Zinc, cadmium,

mercury and selenium in Minke Whales, Belugas and Narwhals from West

Greenland. Polar Biology 10:529-539.

Hansen, D. J. 1989. U.S. Environmental Protection Agency regulations and criteria

for mercury in water. Summary presentation to the coordination team. In:

Mercury in the Marine Environment. Workshop proceedings. U.S. Dept. Interior,

Mineral Management Service, Anchorage, Alaska. OCS Study MMS 89-0049.

Hara, T.J., Y.M.C. Law, and S. McDonald. 1976. Effects of mercury and copper on

the olfactory response in rainbow trout (Oncorhynchus mykiss). J Fish Res Bd Can

33:1568-1573.

Heisinger, F. J. and W. Green. 1975. Mercuric chloride uptake by eggs of the rice

fish and resulting teratogenic effects. Bull. Environ. Contam. Toxicol. 14:665-673.

Hill, W.R., A.J. Stewart, and G.E. Napolitano. 1996. Mercury speciation and

bioaccumulation in lotic primary producers and primary consumers. Can. J. Fish.

Aquat. Sci. 53:812-819.

51
Hines, M. E., S. L. Knollmeyer, and J.B. Tugel. 1989. Sulfate reduction and other

sedimentary biogeochemistry in a northern New England salt marsh. Limnol.

Oceanogr. 34(3): 578-590.

Hintelmann, H., P. M. Welbourn, and R. D. Evans. 1995. Binding of methylmercury

compounds by humic and fulvic acids. Water, Air, and Soil Pollution 80:1031-1034.

Hoff, R., H. Curl, Jr., J. Farr, and N. Beckvar. 1994. Empire Knight: Assessing

environmental risk. NOAA Technical Memorandum NOAA ORCA 81. Seattle:

Hazardous Materials Response and Assessment Division, National Oceanic and

Atmospheric Administration. 31pp.

Holden, A. V. 1978. Pollutants and seals-A review. Mammal Rev. 8(1-2):53-66.

Horvat, M. 1991. Determination of methylmercury in biological certified reference

materials. Water, Air, and Soil Pollution 56:95-102.

Huckabee, J., J. Elwood, and S. Hildebrand. 1979. Accumulation of mercury in

freshwater biota. In: Nriagu (ed.) The Biogeochemistry of Mercury in the

Environment. pp 277-302. New York: Elsevier/North-Holland Biomedical Press

1979.

Hughes, W. L. 1957. A physicochemical rationale for the biological activity of

mercury and its compounds. Ann. New York Acad. Sci. 65: 454-460.

Jackson, T.A. 1991. Biological and environmental control of mercury accumulation

by fish in lakes and reservoirs of northern Manitoba, Canada. Can. J. Fish. Aquat.

Sci. 48:2449-2470.

Jackson, T. A. 1986. Methyl mercury levels in a polluted prairie river-lake system:

seasonal and site specific variations, and the dominant influence of trophic

conditions. Canadian Journal of Fisheries and Aquatic Sciences 43:1873-1877.

Jackson, T. A. 1987. Methylation, demethylation, and bio-accumulation of mercury

in lakes and reservoirs of northern Manitoba, with particular reference to effects of

environmental changes caused by the Churchill-Nelson River diversion. In:

Summary Report, Canada-Manitoba Agreement on the Study and Monitoring of

Mercury in the Churchill River Diversion. Ottawa: Governments of Canada and

Manitoba.

52
Jackson, T. A. 1988. Accumulation of mercury by plankton and benthic

invertebrates in riverine lakes of northern Manitoba (Canada): Importance of

regionally and seasonally varying environmental factors. Canadian Journal of

Fisheries and Aquatic Sciences 45:1744-1757.

Jackson, T. A. 1989. The influence of clay minerals, oxides, and humic matter on

the methylation and demethylation of mercury by micro-organisms in freshwater

environments. Applied Organometallic Chemistry 3:1-30.

Jernelov, A. 1968. Laboratory Experiments regarding the conversion of mercury

into its different forms of occurrence 1. Vatten 24(1):53-56. Translation by Canada

Dept. Sed. State Transl. Bur..

Joiris, C. R., L. Holsbeek, J. M. Bouquegneau, and M. Bossicart. 1991. Mercury

contamination of the harbour porpoise Phocoena phocoena and other cetaceans

from the North Sea and the Kattegat. Wat. Air Soil Pollut. 56: 283-293.

Julshamn, K. O. Ringdal, and O. R. Braekkan. 1982. Mercury concentration in liver

and muscle of Cod (Gadus morhua) as an evidence of migration between waters

with different levels of mercury. Bull. Environm. Contam. Toxicol. 29:544-549.

Julshamn, K., A. Andersen, O. Ringdal, and J. Morkore. 1987. Trace elements intake

in the Faroe Islands, I. Element levels in edible parts of Pilot Whales (Globicephalus

meleanus ). Sci. Total Environ. 65:53-62.

Kelly, C.A., J.W.M. Rudd, V.L. St. Louis, and A. Heyes. 1995. Is total mercury

concentration a good predictor of methyl mercury concentration in aquatic

systems? Water, Air, and Soil Pollution 80:715-724.

Kihlstrom, J. E. and L. Hulth. 1972. The effect of phenylmercuric acetate upon the

frequency of hatching in the zebrafish. Bull. Environ. Contam. Toxicol. 7:111.

Kobayashi, N. 1984. Marine ecotoxicological testing with echinoderms. In: G.

Persoone, E. Jaspers, and C. Claus (Eds.), Ecotoxicological Testing for the marine

Environment. Bredene, Belgium: State University Ghent and Institute Mar. Scient.

Res. pp. 341-381.

53
Kopfler, F. C. 1974. The accumulation of organic and inorganic mercury

compounds by the eastern oyster (Crassostrea virginica). Bull. Environ. Contam.

Toxicol. 11:275-280.

Korthals, E. T and M. R. Winfrey. 1987. Seasonal and spatial variations in mercury

methylation and demethylation in an oligotrophic lake. Applied and

Environmental Microbiology 53:2397-2404.

Langston, W. J. 1990. Chapter 7: Toxic effects of metals and the incidence of metal

pollution in marine environments. In: R. W. Furness and P. S. Rainbow (Eds.),

Heavy Metals in the Marine Environment. pp. 101-122. Boca Raton, Florida: CRC

Press, Inc.

Langston, W. J. 1986. Metals in sediments and benthic organisms in the Mersey

Estuary. Estuarine, Coastal and Shelf Science 23:239-261.

Langston, W. J. 1982. The distribution of mercury in British estuarine sediments

and its availability to deposit-feeding bivalves. J. Mar. Biol. Ass. U.K. 62: 667-684.

Lasorsa, B. and S. Allen-Gil. 1995. The methylmercury to total mercury ratio in

selected marine, freshwater, and terrestrial organisms. Water, Air, and Soil Pollution

80:905-913.

Leah, R.T., S.J. Evans, and M.S. Johnson. 1992. Mercury in flounder (Platichthys

flesus L.) from estuaries and coastal waters of the north-east Irish Sea.

Environmental Pollution 75:317-322.

Lenka, M., K.K. Panda, and B. B. Panda. 1990. Studies on the ability of water

hyacinth (Eichlornia crassipes) to bioconcentrate and biomonitor aquatic mercury.

Environ. Pollut. 66: 89-99.

Leonzio, C., S. Focardi, and C. Fossi. 1992. Heavy metals and selenium in stranded

dolphins of the Northern Tyrrhenian (NW Mediterranean). Sci. Total Environ.

119:77-84.

Lindberg, S. E. and R.C. Harriss. 1974. Mercury-organic matter associations in

estuarine sediments and interstitial water. Envir. Sci. Tech. 8(5):459-462.

54
Lindqvist, O. ed. 1991. Mercury in the Swedish environment: Recent research on

causes, consequences and corrective methods. Water, Air, and Soil Pollution 55(1-

2). Special Issue. 261pp.

Long, E. R., D. D. MacDonald, S. L. Smith, and F. D. Calder. 1995. Incidence of

adverse biological effects within ranges of chemical concentrations in marine and

estuarine sediments. Environmental Management 19(1):81-97.

Long, E. R. and D. D. MacDonald. 1992. National Status and Trends Program

Approach. In: Sediment Classification Methods Compendium. EPA 823-R-92-006.

EPA Office of Water (WH-556). Washington, DC.: U.S. Environmental Protection

Agency.

Luoma, S.N. 1977. The dynamics of biologically available mercury in a small

estuary. Estuarine Coast. Mar. Sci. 5:643-652.

Major, M.A., D.H. Rosenblatt and K.A. Bostian. 1991. The octanol/water partition

coefficient of methylmercuric chloride and methylmercuric hydroxide in pure

water and salt solutions. Environ. Toxicol. Chem. 10:5-8.

Mason, R.P., J.R. Reinfelder, and F.M.M. Morel. 1996. Uptake, toxicity, and trophic

transfer of mercury in a coastal diatom. Environ. Sci. Technol. 30:1835-1845.

Mason, R.P., K.R. Rolfhus and W.F. Fitzgerald. 1995a. Methylated and elemental

mercury cycling in surface and deep ocean waters of the North Atlantic. Water, Air,

and Soil Pollution 80: 665-677.

Mason, R.P., J.R. Reinfelder, and F.M.M. Morel. 1995b. Bioaccumulation of mercury

and methylmercury. Water, Air, and Soil Pollution 80:915-921.

Mathers, R. and P. Johansen. 1985. The effects of feeding ecology on mercury

accumulation in walleye (Stizostedion vitreum) and pike (Esox lucius) in Lake Simcoe.

Canadian Journal of Zoology 63: 2006-2012.

Matida, Y., H. Kumada, S. Kimura, Y. Saiga, T. Nose, M. Yokote, and H. Kawatsu.

1971. Toxicity of mercury compounds to aquatic organisms and accumulation of

the compounds by the organisms. Bull. Freshwater Fish. Res. Lab. (Tokyo) 21:197-

227.

55
McGreer, E. R. 1979. Sublethal effects of heavy metal contaminated sediments on

the bivalve Macoma balthica (L). Mar. Pollut. Bull. 10:259.

McKenney, C. L., Jr. and J. D. Costlow, Jr. 1981. The effects of salinity and mercury

on developing megalopae and early crab stages of the blue crab Callinectes sapidus

Rathbun. In. F. J. Vernberg, A. Calabrese, F. P. Thurberg, and W. B. Vernberg (Eds.),

Biological Monitoring of Marine Pollutants. pp. 241-262. New York: Academic Press.

McKim, J.M., G. F. Olson, G. W. Holcombe, and E.P. Hunt. 1976. Long-term effects

of methylmercuric chloride on three generations of brook trout (Salvelinus

fontinalis): Toxicity, accumulation, distribution, and elimination. Journal Fisheries

Research Board of Canada 33:2726-2739.

Meili, M.. 1995. Liming effects on mercury concentrations in fish. In: L. Henrikson

and Y. W. Brodin (eds). Liming Acidified Sur. Waters. p. 383-398. Berlin: Springer.

Mierle, G., and R. Ingram. 1991. The role of humic substances in the mobilization

of mercury from watersheds. Water, Air, and Soil Pollution 56:349-357.

Miskimmin, B. M., J. W. M. Rudd, and C. Kelly. 1992. Influence of dissolved organic

carbon, pH, and microbial respiration rates on mercury methylation and

demethylation in lake water. Can. J. Fish. Aquat. Sci. 49:17-22.

Møhlenberg, F. and H.U. Riisgård. 1988. Partitioning of inorganic and organic

mercury in cockles Cardium edule (L.) and C. Glaucum (Bruguiere) from a

chronically polluted area: influence of size and age. Environ. Pollut. 55:137-148.

Nicoletto, P. and A. Hendricks. 1987. Sexual differences in accumulation of

mercury in four species of centrarchid fishes. Canadian Journal of Zoology 66:944-

949.

Niimi, A.J. and G.P. Kissoon. 1994. Evaluation of the critical body burden concept

based on inorganic and organic mercury toxicity to rainbow trout (Oncorhynchus

mykiss). Arch. Environ. Contam. Toxicol. 26:169-178.

Nishimura H., and M. Kumagai. 1983. Mercury pollution of fishes in Minamata Bay

and surrounding water: analysis of pathway of mercury. Water, Air, and Soil

Pollution 20:401-411.

56
Olson, B.H. and R.C. Cooper. 1976. Comparison of aerobic and anaerobic

methylation of mercuric chloride by San Francisco Bay sediments. Water Res. 10:

113-116.

Olson, K R., K. S. Squibb, and R. J. Cousins. 1978. Tissue uptake, subcellular

distribution, and metabolism of 14CH3HgCl and CH3

203 HgCl by rainbow trout,

Salmo gairdneri. J. Fish. Res. Board Can. 35:381-390.

Parks, J. W., A. Lutz, and J. A. Sutton. 1989. Water column methylmercury in the

Wabigoon/English River-Lake System: Factors controlling concentration, speciation,

and net production. Canadian Journal of Fisheries and Aquatic Sciences 46:2184-

2202.

Passow, H., A. Rothstein, and T. Clarkson. 1961. The general pharmacology of the

heavy metals. Pharmacol. Rev. 13:185-224.

Pentreath, R.J. 1976a. The accumulation of organic mercury from sea water by the

plaice, Pleuronectes platessa L. Journal of Experimental Marine Biology and Ecology

24:121-132.

Pentreath, R.J. 1976b. The accumulation of inorganic mercury from sea water by

the plaice, Pleuronectes platessa L. Journal of Experimental Marine Biology and

Ecology 24:103-119.

PTI. 1991. Onondaga Lake RI/FS Work Plan. PTI Environmental Services, Bellevue,

Washington.

PTI. 1988. Briefing report to the EPA Science Advisory Board: The Apparent Effects

Threshold approach. Seattle: Environmental Protection Agency, Region 10, Office of

Puget Sound. 57 pp.

Rada, R. G., J. G. Wiener, M. R. Winfrey, D. E. Powell. 1989a. Recent increases in

atmospheric deposition of mercury to north-central Wisconsin lakes inferred from

sediment analyses. Arch. Environ. Contam. Toxicol. 18:175-181.

Rada, R. G., J. E. Findley, and J. G. Wiener. 1989b. Environmental fate of mercury

discharged into the upper Wisconsin River. Water, Air, and Soil Pollution 29:57-76.

57
Rada, R. G., J. E. Findley, and J. G. Wiener. 1986. Environmental fate of mercury

discharged into the upper Wisconsin River. Water Air Soil Pollut. 29:57-76.

Ramlal, P. S., J.W.M. Rudd, A. Furutani, L. Xun. 1985. The effect of pH on methyl

mercury production and decomposition in lake sediments. Canadian Journal of

Fisheries and Aquatic Sciences 42: 685-692.

Regnell, O. 1994. The effect of pH and dissolved oxygen levels on methylation and

partitioning of mercury in freshwater model systems. Environmental Pollution 84:7-

13.

Rehnberg, B.C. and C.B. Schreck. 1986. Acute metal toxicology of olfaction in coho

salmon: behavior, receptors, and odor-metal complexation. Bull. Environ. Contam.

Toxicol. 36:579-586.

Reigel, D. V. 1990. The distribution and behavior of mercury in sediments and

marine organisms of Lavaca Bay, Texas. College Station, Texas: The Texas A&M

University. Masters Thesis.

Richardson, G.M. and D.J. Currie. 1996. Does acid precipitation exacerbate the

problem of fish mercury contamination? SETAC News 16(2): 14.

Riisgård , H.U. and P. Famme. 1988. Distribution and mobility of organic and

inorganic mercury in flounder, Platichthys flesus, from a chronically polluted area.

Toxicol. Environ. Chem. 16:219-228.

Riisgård , H. U. and P. Famme. 1986. Accumulation of inorganic and organic

mercury in shrimp, Crangon crangon. Marine Pollution Bulletin 17:255-257.

Riisgård, H.U., T. Kiorboe, F. Møhlenberg, I. Drabk, and P. Pheiffer Madsen. 1985.

Accumulation, elimination and chemical speciation of mercury in the bivalves

Mytilus edulis and Macoma balthica. Mar. Biol. 86:55-62.

Riisgård , H. U. and S. Hansen. 1990. Biomagnification of mercury in a marine

grazing food-chain: algal cells Phaeodactylum tricornutum, mussels Mytilus edulis

and flounders Platichthys flesus studied by means of a stepwise-reduction-CVAA

method. Marine Ecology Progress Series 62:259-270.

58
Rodgers, D.W. and F.W.H. Beamish. 1982. Dynamics of dietary methylmercury in

rainbow trout, Salmo gairdneri. Aquatic Toxicology 2: 271-290.

Rodgers, D.W. and F. W. H. Beamish. 1981. Uptake of waterborne methylmercury

by rainbow trout (Salmo gairdneri) in relation to oxygen consumption and

methylmercury concentration. Canadian Journal of Fisheries and Aquatic Sciences

38(11): 1309-1315.

Roesijadi, G., A.S. Drum, and J.R. Bridge. 1981. Mercury in mussels of Bellingham

Bay, Washington (U.S.A.): the occurrence of mercury-binding proteins. In:

Vernberg, F.J., Calabrese, A. Thurberg, F.P. and W.B. Vernberg (Eds.), Biological

Monitoring of Marine Pollutants. pp. 357-376. New York: Academy Press.

Rubinstein, N.I., E. Lores, and N.R. Gregory. 1983. Accumulation of PCBs, mercury

and cadmium by Nereis virens, Mercenaria mercenaria and Palaemonetes pugio from

contaminated harbor sediments. Aquatic Toxicology 3:249-260.

Rudd, J. W. M. and M. A. Turner. 1983. The English-Wabigoon River System:

V. Mercury and selenium bioaccumulation as a function of aquatic primary

productivity. Canadian Journal of Fisheries and Aquatic Sciences 40:2251-2259.

Rudd, J. W. M., M. A. Turner, A. Furutani, A. L. Swick, and B. E. Townsend. 1983.

The English-Wabigoon River System: I. A synthesis of recent research with a view

towards mercury amelioration. Canadian Journal of Fisheries and Aquatic Sciences

40:2206-2217.

Saouter, E., L. Hare, P. G. C. Campbell, A. Boudou, and F. Ribeyre. 1993. Mercury

accumulation in the burrowing mayfly Hexagenia rigida (Ephemeroptera) exposed

to CH3HgCl or HgCl2 in water and sediment. Wat. Res. 27(6):1041-1048.

Saouter, E. F. Ribeyre, A. Boudou, and R. Maury-Brachet. 1991. Hexagenia rigida

(Ephemeroptera) as a biological model in aquatic ecotoxicology: experimental

studies on mercury transfers from sediment. Environmental Pollution 69:51-67.

Saroff, S. T. 1990. Proceedings of the Onondaga Lake Remediation Conference.

Bolton Landing, New York: New York State Department of Law and New York State

Department of Environmental Conservation. 193 pp.

59
Sastry, K. V. and K. Sharma. 1980. Effects of mercuric chloride on the activities of

brain enzymes in a freshwater teleost, Ophiocephalus (Channa) punctatus. Arch.

Environ. Contam. Toxicol. 9:425-430.

Scherer, E., F. A. J. Armstrong, and S. H. Nowak. 1975. Effects of mercurycontaminated diet upon walleyes Stizostedion vitreum vitreum (Mitchell). Fish. Mar.

Serv. Tech. Rep. No. 597. Winnipeg, Manitoba: Fisheries Marine Service. 21 pp.

Schindler, D.W., S.E. Bayley, P.J. Curtis, B.R. Parker, M.P. Stanton and C.A. Kelly.

1992. Natural and man-caused factors affecting the abundance and cycling of

dissolved organic substances in precambrian shield lakes. Hydrobiologia 229:1-21.

Schintu, M., F. Jean-Caurant, and J. C. Amiard. 1992. Organomercury

determination in biological reference material: application to a study on mercury

speciation in marine mammals off the Faröe Islands. Ecotoxicology and

Environmental Safety 24:95-101.

Sharp, J. R. and J. M. Neff. 1980. Effects of the duration of exposure to mercuric

chloride on the embryogenesis of the estuarine teleost, Fundulus heteroclitus, Mar.

Environ. Res. 3: 195-213.

Shoichi, O. and S. Sokichi. 1985. The 1-octanol/water partition coefficient of

mercury. Bull. Chem. Soc. Japan. 58:3401-3402.

Slooff, W., P.F.H. Bont, M. van Ewijk, and J.A. Janus. 1991. Exploratory report

mercury. Report no. 710401006. Bilthoven, The Netherlands: National Institute of

Public Health and Environmental Protection .

Snarski, V. M. and G. F. Olson. 1982. Chronic toxicity and bioaccumulation of

mercuric chloride in the fathead minnow (Pimephales promelas). Aquatic Toxicology

2:143-156.

Spry, D. J. 1991. Metal bioavailability and toxicity to fish in low-alkalinity lakes: a

critical review. Environ. Pollut. 71:243-304.

Spry, D. J. and J. G. Wiener. 1991. Metal bioavailability and toxicity to fish in lowalkalinity lakes: A critical review. Environmental Pollution 71:243-304.

60
St. Louis, V.L., J.W.M. Rudd, C.A. Kelly, K.G. Beaty, N.S. Bloom and R.J. Flett. 1994.

Importance of wetlands as sources of methyl mercury to boreal forest ecosystems.

Can. J. Fish. Aquat. Sci. 51:1065-1076.

Stromgren, T. 1982. Effect of heavy metals (Zn, Hg, Cu, Cd, Pb, Ni) on the length

growth of Mytilus edulis. Mar. Biol. 72:69-72.

Surma-Aho, K. and J. Paasivirta. 1986. Organic and inorganic mercury in the food

chain of some lakes and reservoirs in Finland. Chemosphere 15(3):353-372.

Swartz, R. C., P. F. Kemp, D. W. Schults, and J. O. Lamberson. 1988. Effects of

mixtures of sediment contaminants on the marine infaunal amphipod, Rhepoxynius

abronius. Environmental Toxicology and Chemistry 7:1013-1020.

Syversen, T. 1977. Effects of methylmercury on in vivo protein synthesis in

isolated cerebral and cerebellar neurons. Neuropathol. Appl. Neurobiol. 3:225-236.

Szefer. P., W. Czarnowski, J. Pempkowiak, and E. Holm. 1993. Mercury and major

essential elements in seals, penguins, and other representative fauna of the

Antarctic. Arch. Environ. Contam. Toxicol. 25:422-427.

Tessier, L., G. Vaillancourt, and L. Pazdernik. 1994. Temperature effects on

cadmium and mercury kinetics in freshwater molluscs under laboratory conditions.

Arch. Environ. Contam. Toxicol. 26:179-184.

Tetra Tech, Inc. 1988. Health risk assessment of chemical contamination in Puget

Sound seafood. Seattle: Environmental Protection Agency, Region 10, Office of

Puget Sound. 102 pp + appendices.

Thain, J. E. 1984. Effects of mercury on the prosobranch mollusc Crepidula

fornicata: Acute lethal toxicity and effects on growth and reproduction of chronic

exposure. Mar. Envir. Res. 12:285-309.

Thompson, D. R. 1990. Metal levels in marine vertebrates. In: W. D. Furness and

P. S. Rainbow (Eds.), Heavy Metals in the Marine Environment. pp. 144-182. Boca

Raton, Florida: CRC Press.

61
U.S. Environmental Protection Agency (EPA). 1985. Ambient water quality criteria

for mercury - 1984. U.S. EPA 440/5-84-026. Washington, D.C.: Office of Water.

136 pp.

U.S. Environmental Protection Agency (EPA). 1996. Update: National listing of fish

and wildlife consumption advisories. EPA Fact Sheet EPA-823-F-96-006.

Washington, D.C.: Office of Water.

U.S. Food and Drug Administration (FDA). 1984. Action level for methyl mercury

in fish. Federal Register 49:45663- . November 19, 1984.

Wagemann, R., N. B. Snow, A. Lutz, and D. P. Scott. 1983. Heavy metals in tissues

and organs of the narwhal (Monodon monoceros ). Can. J. Fish. Aquat. Sci. 40(Suppl.

2):206-214.

Weber, J. H.. 1993. Review of possible paths for abiotic methylation of mercury (II)

in the aquatic environment. Chemosphere 26(11):2063-2077.

Weis, J. S. and P. Weis. 1977. Effects of heavy metals on development of the

killifish, Fundulus heteroclitus. J. Fish. Biol. 11:49-54.

Weis, P., and J.S. Weis. 1978. Methylmercury inhibition of fin renereration in fishes

and its interaction with salinity and cadmium. Estuarine Coastal Mar. Sci. 6:327-

334.

Weis, J. S. and P. Weis. 1984. A rapid change in methylmercury tolerance in a

population of killifish, Fundulus heteroclitus, from a golf course pond. Marine

Environmental Research 13:231-245.

Weis, J. S,. and P. Weis. 1989. Effects of environmental pollutants on early fish

development. Rev. Aquat. Sci. 1:45-73.

Weis, J. S., P. Weis, and M. Heber. 1982. Variation in response to methylmercury by

killifish (Fundulus heteroclitus) embryos. In: J.G. Pearson, R.B. Foster, and W.E.

Bishop (eds.), Aquatic Toxicology and Hazard Assessment: Fifth Conference. ASTM

STP 766. pp. 109-119. Philadelphia: American Society for Testing and Materials.

62
Weis, J. S., P. Weis, M. Heber, and S. Vaidya. 1981. Methylmercury tolerance of

killifish (Fundulus heteroclitus) embryos from a polluted vs non-polluted

environment. Marine Biology 65:283-287.

Weis, P. , J. S. Weis, and J. Bogden. 1986. Effects of environmental factors on release

of mercury from Berry’s Creek (New Jersey) sediments and its uptake by Killifish

Fundulus heteroclitus. Environmental Pollution 40:303-315.

Weis, P. 1984. Metallothionein and mercury tolerance in the killifish, Fundulus

heteroclitus. Marine Environmental Research 14:153-166.

Whitney, S. D. 1991. Effects of maternally-transmitted mercury on the hatching

success, survival, growth, and behavior of embryo and larval walleye (Stizostedion

vitreum vitreum). Masters Thesis. La Crosse, Wisconsin: University of Wisconsin. 71

pp.

Wiener, J. G. and D. J. Spry. 1996. Toxicological significance of mercury in

freshwater fish. In: Beyer, W.N., G.H. Heinz, and A.W. Redmon-Norwood (eds.).

Environmental Contaminants in Wildlife: Interpreting Tissue Concentrations. Special

Publication of the Society of Environmental Toxicology and Chemistry. Boca Raton,

Florida: Lewis Publishers. 494 pp.

Windom, H.L. and D. R. Kendall. 1979. Accumulation and biotransformation of

mercury in coastal and marine biota. In: J.O. Nriagu (ed.) The Biogeochemistry of

Mercury in the Environment. pp 277-302. Elsevier/North-Holland Biomedical Press

1979.

Winfrey, M. R. and J. W. Rudd. 1990. Environmental factors affecting the formation

of methylmercury in low pH lakes. Environ. Toxicol. Chem. 9:853-869.

Wobeser, G. 1975. Prolonged oral administration of methyl mercury chloride to

rainbow trout (Salmo gairdneri) fingerlings. J. Fish.Res.Board Can. 32:2015-2023.

World Health Organization (WHO). 1989. Mercury — Environmental Aspects.

International Programme on Chemical Safety (IPCS). Geneva, Switzerland. 115 pp.

Wren, C. D. 1986. A Review of Metal Accumulation and Toxicity in Wild

Mammals, I: Mercury. Environmental Research 40:210 -244.

63
Xun, L., N.E.R. Campbell, and J.W.M. Rudd. 1987. Measurements of specific rates of

net methyl mercury production in the water column and surface sediments of

acidified and circumneutral lakes. Can. J. Fish. Aquat. Sci. 44:750-757.

Zillioux, E. J., D. B. Porcella, and J. M. Benoit. 1993. Mercury cycling and effects in

freshwater wetland ecosystems. Environmental Toxicology and Chemistry 12:1-20.

64