Mercury Transport 

and Fate Through 

a Watershed

January 2006

Synthesis Report of Research from EPA’s Science to Achieve Results (STAR) Grant Program
P R E PA R E D F O R

U.S. Environmental Protection Agency, Office of Research and Development

National Center for Environmental Research

Washington, DC

P R E PA R E D B Y

ICF International

Research Triangle Park, NC, and Fairfax, VA

Under EPA Contract No. 68-C-03-137

Work Assignments 00-01, 01-01, and 02-03
Disclaimer

The research described in this document has been funded wholly by the United States Environmental Protection Agency 

(EPA) under the Science to Achieve Results (STAR) grants program. The information does not necessarily reflect the 

views of the Agency and no official endorsement should be inferred. Mention of trade names or commercial products 

does not constitute endorsement or recommendation by EPA for use.

The information presented in this synthesis report is intended to provide the reader with insights about the progress 

and scientific achievements of STAR research grants. The report lists the grantees whose research is discussed and also 

indicates where more detailed peer-reviewed scientific data can be found. This report is not sufficiently detailed nor is 

it intended to be used directly for environmental assessments or decision making. Readers with these interests should 

instead consult the peer reviewed publications produced by the STAR grantees and conduct necessary data quality 

evaluations as required for their assessments. 



Contents

Photo Acknowledgements ii

Executive Summary iii

1. Introduction 1

1.1 Purpose of this Report 1

1.2 Background 1

1.3 Organization of this Report 1

2. Overview of Star Mercury Grants 2

2.1 Overview of Mercury in the Environment 2

2.2 FY99 STAR Mercury Grants 2

2.3 Relationship of 1999 Mercury STAR Grants to Office of Research and Development’s Multi-Year Plans 3

2.4 Research Themes 4

3. Study Results 7

3.1 Biogeochemical Controls of Mercury Transformations 7

3.2 Biogeochemical Controls of Mercury Cycling in the Environment 11

3.3 Sources and Distribution of Mercury in Terrestrial and Aquatic Systems 13

3.4 Bioavailability and Bioaccumulation of Mercury in Aquatic Systems 16

3.5 New Methods for Mercury Analysis 17

4. Applications of Findings from Star Grants 18

4.1 Potential Improvements to Mercury Fate and Transport Models 18

4.2 Understanding Key Variables Affecting Mercury Fate and Transport 19

4.3 Understanding Variation in Key Variables by Ecoregion and Ecotype 19

5. Future Research Needs 22

6. References Cited 24

Appendix: Grant Publication Lists A-1
ii

Photo Acknowledgements

Cover page, page 2, and page 7

Experimental Lakes Area, Ontario, Canada (courtesy of Cynthia Gilmour)

Page iii

Florida Everglades (courtesy of Cynthia Gilmour)

Page 1

Adirondack lake (courtesy of Charles Driscoll)

Page 18

Calcine dumps at New Idria mercury mine, CA (courtesy of Gordon Brown)

Page 22

Steam plumes from smokestacks in Baltimore, MD (courtesy of Cynthia Gilmour)

Page A-1

Experimental spraying of sodium sulfate solution at the Marcell Experimental Forest in northern Minnesota 

(courtesy of the Minnesota Pollution Control Agency). 
iii

Executive Summary

Background

The U.S. Environmental Protection 

Agency’s (EPA’s) Science to Achieve 

Results (STAR) research grants 

program, managed by the National 

Center for Environmental Research, 

has funded significant research on 

the fate, transport, and transformation of mercury in aquatic and 

terrestrial environments. This report 

summarizes the research and findings generated through nine grants 

awarded under the 1999 Request for 

Applications (RFA) entitled “Mercury: 

Transport and Fate through a Watershed” and two other closely related 

STAR grants. The important scientific 

findings from these grants, data gaps, 

and additional research needs on this 

topic are described in this synthesis 

report.

The 1999 RFA funded fundamental 

research on the complex chemical 

and physical transformations and 

movement of mercury through the 

environment. The overall goal of the 

research solicited on mercury transport in a watershed was summarized 

in the original RFA as follows: 

“...to develop a better understanding of terrestrial and aquatic fate 

and transformation processes 

(especially microbial) that mediate ecological and human exposures to mercury. The development 

of improved models of the fate of 

mercury in aquatic and terrestrial systems in order to estimate 

ecosystem response to decreased 

anthropogenic inputs of mercury 

is also needed.”

The RFA specifically invited grant 

applications addressing the following 

critical research questions:

(1) For a given amount of mercury 

transported into a watershed, 

what is the predicted concentration of methylmercury in fish? 

How do mercury and methylmercury spatially distribute 

across the terrestrial and aquatic 

components of a watershed? 

What controls bioavailability of 

mercury in the food chain?

(2) What environmental and 

biochemical variables control 

transformation of mercury to 

methylmercury? What environmental variables control the 

reduction of divalent mercury 

to elemental mercury in soils, 

sediments, and surface waters?

(3) How does mercury cycling vary 

within different geographic regions of the United States (e.g., 

south Florida, Great Lakes, 

northeast, or west)? How might 

the variability be accounted for 

(e.g., resource types (wetlands), 

temperature regimes, microbial 

communities)? 

The eleven STAR grants discussed in 

this report examined how methylmercury transports into and within a 

watershed and concentrates in fish. 

Projects included studies of factors 

that control the partitioning and distribution of different mercury species 

in different environmental media in 

aquatic and terrestrial ecosystems 

(Question 1), factors that influence 

bioavailiability of mercury (Question 1), and studies of the factors 

controlling mercury transformations 

(Question 2). Most studies identified controlling factors that are likely 

to vary by geographic region, but a 

systematic examination of regional 

influences was not attempted (Question 3). 

Note that the critical research questions are interrelated. For example, 

partitioning of mercury into different 

environmental media influences its 

availability to various transforming 

processes, and transforming processes 

can create mercury species that move 

through the environment via different 

pathways. Transformation processes 

also can create “sources” of given 

mercury species in the environment. 

Exhibit ES-1 illustrates these concepts. 

This synthesis report discusses the 

STAR grant project results along four 

research themes. These themes are 

intended to correspond with components needed to develop an integrated 

multimedia modeling framework for 

understanding mercury fate from 

source to fish concentrations. These 

four themes are: (1) controls of 

mercury transformations, (2) controls 

of mercury partitioning and transport 

(i.e., cycling) in the environment, 

(3) sources of mercury species in the 

environment, and (4) bioavailability 

and bioaccumulation of mercury in 

food chains. Key findings of the STAR 

grant projects are organized according 

to these four themes.
iv

Study Results

(1) Biogeochemical controls of 

 mercury transformations. 

Net methylmercury (MeHg) production in aquatic systems is complex. 

Elemental mercury (Hg(0)) can be 

oxidized to Hg(II), the form that 

is methylated by sulfate-reducing 

bacteria (SRB), and Hg(II) is readily 

reduced to Hg(0) in surface waters 

via abiotic and biotic pathways. 

Hg(II) chemical complexes tend to 

become bound to sediment particles, whereas Hg(0) is lost to the 

atmosphere via volatilization. In 

addition, there are rapid conversions 

between inorganic mercury (Hg(II)) 

and MeHg. While SRB in sediments 

methylate Hg(II), MeHg is also 

demethylated via both abiotic (photochemical) and biotic pathways. 

STAR grant projects added to the 

understanding of these processes as 

highlighted below.

• STAR grant research has shown 

that, as mercury is deposited to 

surface waters from air, Hg(II) 

becomes progressively bound in 

forms that are less easily methylated, even with rapid cycling 

between Hg(II) and MeHg. Thus, 

newly deposited Hg(II) is more 

available for methylation than older mercury in some environments. 

This may not be true in every 

environment, but was observed in 

both the Florida Everglades and 

the Experimental Lake Area in 

Canada.

• Mercury that is deposited directly 

to surface waters is readily methylated in the sediments. In contrast, 

it is unclear how long it takes mercury deposited to a watershed to 

be methylated (e.g., via transport 

to sites of methylation in surface 

water sediments).

• Studies demonstrated that the rate 

at which SRB methylate Hg(II) 

depends on a variety of factors, 

including bacterial activity levels 

in general (e.g., seasonal), bacterial community structure (i.e., 

metabolic pathways of species 

present), and the bioavailability of Hg(II), which depends on 

several factors (e.g., sulfides and 

dissolved organic carbon (DOC) 

concentrations) as discussed later. 

• The addition of sulfate to freshwater environments can enhance 

rates of methylation, which is 

expected because sulfate and 

organic matter are the primary 

substrates used by SRB to produce 

energy. However, a build-up of 

sulfides as reaction products can 

reduce the availability of environmental Hg(II) to the bacteria.

• Demethylation is less well studied 

than methylation, yet is important 

in determining net methylation 

rates. It occurs via both biotic and 

abiotic (photochemical) pathways. 

STAR grant research demonstrated 

photochemical demethylation 

requires UV radiation and is confined to the upper surface water 

layers.

• The availability of Hg(II) for 

SRB also depends on the cycling 

between Hg(0) and Hg(II). Photochemical reduction of Hg(II) 

appears to be an important reaction in surface waters, resulting in 

a diurnal cycle of Hg(0) formation 

and release from surface water 

via volatilization. The rate of net 

reduction is related to several 

Exhibit ES-1: Sources and Paths of Mercury in the Environment
factors including light intensity 

and DOC. Oxidation of Hg(0) to 

Hg(II) is enhanced by free radicals 

in surface waters, which tend to 

be formed via photolysis of DOC.

• Recently, understanding of the 

major processes that impact Hg 

methylation has improved so that 

it is now possible to model Hg 

methylation in a more comprehensive way than accomplished in 

previous models. Such approaches are now being developed and 

will soon be incorporated into 

models. However, the current 

understanding of biotic demethylation is not as complete, and 

models for this reaction remain 

relatively crude.

(2) Biogeochemical controls of 

mercury cycling in the environment.

Mercury cycling (i.e., phase partitioning and inter-media transport) is 

key to the distribution of mercury in 

the environment. Most of the STAR 

research in this area focused on processes controlling mercury transfer 

between surface water and sediment, 

surface water and soils and the atmosphere, and plants and the soil and 

atmosphere. 

• STAR grant studies indicated that 

inorganic mercury (Hg(II)) in 

aquatic systems generally is bound 

with DOC. Complexes of reactive 

Hg(II) can coagulate and adsorb 

to suspended sediment particles 

which are deposited to the sediment bed. The influence of DOC 

on mercury partitioning in the 

sediment bed is complicated and 

is influenced by sulfide levels. 

• Studies indicated that the bioavailability of Hg(II) to SRB depends on 

partitioning of Hg(II) between particulate and dissolved phases of the 

sediments. As expected, methylation 

rates vary inversely with the partition coefficient, KD, of Hg(II) and 

positively with the concentration of 

Hg(II) in sediment pore water. 

• A model of Hg(II) partitioning between sediment particles and pore 

water that accounts for adsorption 

and sulfide concentrations has 

been developed and validated for 

two ecosystems (Patuxent River, 

Florida Everglades). The model’s 

predictions of neutral mercurysulfide complexes, which are the 

bioavailable forms, correlated well 

with the methylmercury concentrations measured in field samples.

• STAR projects demonstrated 

that volatilization of Hg(0) from 

surface waters can be as important 

as sediment burial and methylation of Hg(II) in sediments as a 

mechanism of loss of inorganic 

mercury from different types of 

aquatic ecosystems (e.g., Long 

Island Sound, wetlands).

• Mercury in soils can emit to the 

air via volatilization of Hg(0) and 

light-enhanced emissions of mercury sulfide (and other) complexes, perhaps via photoreduction of 

Hg(II) to Hg(0) with subsequent 

Hg(0) volatilization.

• Studies demonstrated that plants 

take up Hg(0) directly from the 

atmosphere via their stomata, and 

plants can accumulate Hg in their 

leaves. The uptake of mercury 

from soils via plant roots appears 

to be a minor uptake pathway. At 

the end of the growing season, 

litterfall from deciduous plants to 

soils and to surface waters is a significant source of mercury input 

to those systems.

(3) Sources and distribution of mercury 

in terrestrial and aquatic systems.

Several STAR grant studies identified 

major sources and means of distribution of mercury in given study areas. 

Researchers identified and investigated important processes influencing movement of mercury throughout a watershed (e.g., production of 

methylmercury, predominant mercury fate and transport processes). In 

addition, several researchers performed general mass balance studies 

to quantify sources and inputs within 

a given watershed. 

• Mass balance models of different 

watersheds indicated that atmospheric deposition is the primary 

source of total mercury to most 

aquatic systems. As indicated by 

mercury loading from tributaries, 

in some watersheds, runoff and 

erosion also can be large sources 

of mercury to a water body. Mass 

balance models indicated that 

watersheds and lakes serve as sinks 

for total mercury, while wetlands 

and lakes serve as sources for 

methylmercury. These models also 

indicated that dry deposition (i.e., 

via litterfall) and throughfall (i.e., 

via precipitation passing through 

the plant canopy) can be more important than wet deposition. Total 

ecosystem mercury deposition is 

a combination of wet deposition, 

litterfall, and throughfall.

• Studies indicated that in situ

methylation of Hg(II) within 

aquatic systems is the dominant 

source of MeHg, at least for freshwater lakes, wetlands (e.g., Florida 

Everglades), and for estuarine 

systems (e.g., Long Island Sound 

and the Chesapeake Bay). 

• Mercury deposited to the terrestrial portions of watersheds is slow 

to move to sites of methylation via 

erosion or runoff.

• Data from one study suggested 

that methylation may also occur in 

estuarine and coastal areas, perhaps in deep waters, sediments, or 

hydrothermal vents where mercury 

concentrations are not greatly affected by human activities. 

• There is some indication from 

studies in the Adirondacks that the 

retention capacity of watersheds 

for mercury has decreased over 

the last century, although possible 

reasons remain obscure.
vi

• STAR grant studies also demonstrated the importance of subsurface transport and transformation 

of Hg into MeHg in overall mercury cycling. The interface between 

ground water and streams appears 

to be an important methylation 

site and source of MeHg, and may 

be a more important source of 

MeHg from forested ecosystems to 

surface waters than expected.

• Transport of MeHg through watersheds may be strongly mediated 

by the transport of colloids. 

(4) Bioavailability and bioaccumulation 

of mercury in aquatic systems.

Several STAR grants investigated 

factors that influence the bioavailability of inorganic mercury to 

SRB for methylation and transport 

up the food chain. Some of these 

factors have been noted briefly 

above. These and other findings are 

described below.

• Newly deposited mercury can be 

more bioavailable than “older” 

mercury.

• Mercury appears to be less bioavailable to algae in surface waters 

with higher DOC levels, perhaps 

because mercury complexed with 

DOC is less bioavailable to SRB for 

methylation.

• Mercury concentrations in biota 

in lakes are generally higher in the 

spring than summer, a trend that 

might result from bio-dilution by 

growing phyto- and zooplankton 

during the summer. 

• Studies showing higher concentrations of methylmercury in 

zooplankton from surface waters 

than deeper waters suggest that 

mercury in surface waters deposited from the atmosphere may be 

more bioavailable than mercury in 

deeper waters. 

• In a study of fish in an Adirondack lake, bioconcentration factor 

(BCF) values of fish increased 

with each trophic level, indicating 

bioaccumulation through the food 

chain consistent with previous 

studies.

• When sulfide levels are high, neutral mercury-sulfide complexes, 

which are the bioavailable forms, 

are more likely to form if the pH 

is low than if the pH is high. This 

relationship may help explain why 

fish tend to have higher tissue 

mercury residues in waters of 

lower pH. 

• Bioavailability of inorganic 

mercury varies depending on its 

chemical form and solubility, with 

substantial variation apparent 

across mining sites in different 

geographic regions.

• Sorption of Hg(II) onto mineral 

particles may effectively sequester mercury in mineral soils, 

particularly at mining sites. Iron 

and aluminum (hydr)oxides are 

particularly effective in sequestering Hg(II) in colloids.

• Mixing zones between two water 

masses (e.g., river/lake interfaces, 

estuaries) demonstrate enhanced 

bioaccumulation. Although 

coagulation and settling of larger 

suspended particles occur in these 

zones, increased concentrations 

of smaller particles (with sorbed 

methylmercury) are available for 

ingestion by zooplankton.

Major Remaining Tasks/Uncertainties

Future research needs related to 

mercury fate and transport were 

identified by comparing the STAR 

research results to the goals outlined in the ORD Mercury Research 

Multi-Year Plan (MYP). The STAR 

research summarized in this report 

identifies fate and transport issues 

and provides information on mercury 

cycling in complex ecosystems. 

Other ORD Mercury MYP Research 

Goals and Measures that are related 

to the grants described here, but 

may require further work to fully 

accomplish, include development of 

a model for mercury in fish, identification of sources of mercury emissions, and the eventual creation of 

an integrated multimedia modeling 

framework for mercury in the environment. 

The research provides information 

that relates to all three of the STAR 

mercury RFA questions, but it is most 

responsive to parts of Question 1 and 

Question 2 by identifying variables 

that control partitioning of mercury 

in the environment and transformation of mercury to methylmercury 

as well as transformations between 

inorganic species of mercury. Further research, which builds on the 

findings presented here, into the 

relationship between mercury in the 

watershed to mercury in biota (Question 1) may be necessary to close 

information gaps regarding uptake 

by fish and other organisms. Also, 

further research into how mercury 

cycling and transformation processes vary by region and ecosystem 

(Question 3) is needed to support 

regional modeling of mercury in the 

environment.

The STAR grant investigators also 

identified general areas where followup work would be beneficial as well 

as specific tasks that would advance 

the science addressed by their projects. Some of the topics identified as 

important for follow-up work include 

processes such as demethylation and 

photooxidation and factors affecting 

methylation such as sulfate concentration, DOC, intracellular sequestration of mercury, and other bacterial 

processes, as well as observations 

that newly deposited mercury contributes more substantially to production of methylmercury than “older” 

mercury.


1. Introduction

1.1 Purpose of this Report

In 1999, the U.S. Environmental 

Protection Agency (EPA) published 

a Request for Applications (RFA) 

entitled “Mercury: Transport and 

Fate through a Watershed” under the 

Science to Achieve Results (STAR) 

program. The purpose of the RFA 

was to solicit proposals to research 

fate, transport, and transformation 

of mercury (Hg) in aquatic and terrestrial environments. Nine grants 

were awarded in 1999 based on this 

RFA, and the work under these has 

recently concluded. The purpose 

of this report is to summarize the 

important scientific findings from 

these and two related grants, to 

describe how these findings have 

improved understanding of mercury 

behavior in the environment, and 

to convey remaining data gaps and 

research needs on this topic. This 

report is designed to provide EPA 

program managers and staff, as well 

as state environmental agencies, 

researchers, and the public, with 

an overview of advances in science 

achieved through these grants and a 

description of areas where additional 

research may be needed.

This report is not intended to be 

used directly for environmental 

assessments or decision making. 

Readers with these interests should 

instead consult the peer reviewed 

publications produced by the STAR 

grantees and conduct necessary data 

quality evaluations as required for 

their assessments.

1.2 Background 

As part of its mission to improve public health and increase the reliability 

with which risks to public health and 

the environment are identified and 

measured, EPA’s Office of Research 

and Development (ORD) funds 

extramural research through its STAR 

program. The program is managed by 

ORD’s National Center for Environmental Research (NCER). It provides 

grants and fellowships to scientists 

to address research goals in various 

environmental science and engineering disciplines through a competitive 

solicitation process and independent 

peer review. Many of these research 

goals are outlined in ORD’s Multi-year 

Plans (MYPs), which ORD develops to 

communicate research proposed for 

the next five to eight years. The STAR 

program engages the nation’s best 

scientists and engineers in targeted 

research that complements EPA’s own 

intramural research program and 

research of EPA’s partners in other 

federal agencies.

NCER issues RFAs for topics based on 

ORD’s Strategic Plan (EPA 2001) and, 

more specifically, on the research 

goals described in the MYPs. These 

RFAs are prepared in cooperation 

with other parts of the Agency and 

concentrate on areas of special 

significance to EPA’s mission. For 

example, ORD’s mercury research 

directly contributes to meeting one 

strategic goal “Goal 4: Healthy Communities and Ecosystems” in the EPA 

Strategic Plan (2003a) and supports 

achievement of the four other goals 

(see text box). 

1.3 Organization of this Report

This report describes the research 

performed under each of the nine 

grants awarded under this RFA as 

well as research performed under 

two additional, closely related 

STAR grants1. Section 2 provides 

a brief overview of mercury fate 

and transport, identifies the STAR 

research grants that addressed this 

topic, and describes the research 

themes covered by their results. Section 3 describes key findings from 

these grants according to research 

theme, emphasizing how they have 

advanced general understanding of 

mercury fate and transport. Section 

4 describes how this information can 

be practically applied to address EPA’s 

goals. Section 5 discusses outstanding research needs and data gaps 

identified by the grantees and other 

sources. Section 6 lists the cited 

references, and the Appendix to this 

report provides the lists of publications derived from the research 

conducted under these grants.

Alignment of ORD’s Mercury 

Research with EPA’s Strategic Goals

Directly contributes to:

Goal 4: Healthy Communities 

and Ecosystems

Supports achievement of:

Goal 1: Clean Air and Global 

Climate Change

Goal 2: Clean and Safe Water

Goal 3: Land Preservation 

and Restoration

Goal 5: Compliance and 

Environmental Stewardship

Source: EPA 2003a

1 Two research grants lead by Mae S. Gustin are presented here because they involve mercury fate and transport although they were not funded 

under the FY99 mercury RFA. These grants were instead funded under the Exploratory Research - Air Chemistry & Physics (1996) and Experimental 

Program to Stimulate Competitive Research (1998) STAR RFAs. 


This section first provides a brief 

overview of issues and information 

needs concerning mercury in the 

environment at the time ORD issued 

its RFA on this topic (Section 2.1). It 

then describes the 1999 STAR program mercury RFA and lists the nine 

STAR grants awarded under this RFA 

in 1999 and two other related STAR 

grants awarded under two other 

RFAs (Section 2.2). The relationship of those grants to ORD’s MYPs 

is highlighted next (Section 2.3). 

Finally, the overarching research 

themes covered by the results of the 

grants are described (Section 2.4).

2.1 Overview of Mercury in the 

Environment

As summarized in the introductory 

material included in the original solicitation for these 1999 STAR grants, 

the presence of mercury in the 

environment poses potential risks 

to human health and wildlife. Both 

natural and anthropogenic sources 

contribute to mercury in the environment, with a substantial increase 

in the contribution from anthropogenic sources since the beginning of 

the industrial age. Mercury can cycle 

between various environmental media, including air, land, water, and 

biota, through deposition, volatilization, and other fate and transport 

processes (see Exhibit 1). It is a 

metal with complicated chemistry 

and can transform between different 

chemical species, including elemental mercury liquid and vapor, inorganic salts, and organic forms (e.g., 

methylmercury), through a series 

of complex chemical and physical 

transformations. Methylmercury is 

toxic and can bioaccumulate in the 

tissues of animals. Consumption of 

contaminated fish is a predominant 

exposure pathway for human and 

wildlife populations. A number of 

adverse effects have been linked 

to mercury exposure, including 

behavioral abnormalities, impaired 

growth and development, reduced 

reproductive success, and death. Of 

particular concern are the potential 

neurotoxic effects of methylmercury 

exposure on the developing fetus. 

As required under section 

112(n)(1)(B) of the Clean Air Act, as 

amended in 1990, EPA has prepared the Mercury Study Report to 

Congress (EPA 1997a), which summarizes the magnitude of mercury 

emissions in the United States, the 

health and environmental effects 

of the emissions, and the cost and 

availability of control technologies. 

This and other major EPA reports 

(e.g., the Great Waters Second 

Report to Congress (EPA 1997b), the 

Utility Air Toxics Report to Congress (EPA 1998)) stress the adverse 

effects of mercury exposure on humans and wildlife. Those and other 

publications indicate that additional 

research is needed to better understand mercury fate and transport in 

the environment. 

EPA has taken and is planning different actions to address the health 

and environmental concerns about 

mercury, and the STAR research results should be useful to the Agency 

in these important and high-profile 

efforts. For example, most of the 

research results from these STAR 

grants have been published in time 

to be considered as the Agency 

finalized its regulations to reduce 

mercury emissions from coal-fired 

power plants. 

2.2 FY99 STAR Mercury Grants

The “Mercury: Transport and Fate 

through a Watershed” RFA was 

issued to fund fundamental research on the complex chemical 

and physical transformations and 

movement of mercury through the 

environment. The outcome of this 

research was intended to increase 

EPA’s ability to trace mercury from its 

entrance into an ecosystem through 

its biogeochemical cycling to the 

concentration of methylmercury in 

fish tissues in particular. The overall 

goal of the research solicited on 

mercury transport in a watershed 

context was summarized in the original RFA as follows: 

“The goal of this solicitation is to 

develop a better understanding of 

terrestrial and aquatic fate and 

transformation processes (especially microbial) that mediate 

ecological and human exposures 

to mercury. The development of 

improved models of the fate of 

mercury in aquatic and terrestrial systems in order to estimate 

ecosystem response to decreased 

anthropogenic inputs of mercury 

is also needed.”

The RFA specifically invited grant applications addressing the following 

critical research questions:

2. Overview of Star Mercury Grants


(1) For a given amount of mercury 

transported into a watershed, 

what is the predicted concentration of methylmercury in fish? 

How do mercury and methylmercury spatially distribute 

across the terrestrial and aquatic 

components of a watershed? 

What controls bioavailability of 

mercury in the food chain?

(2) What environmental and 

biochemical variables control 

transformation of mercury to 

methylmercury? What environmental variables control the 

reduction of divalent mercury 

to elemental mercury in soils, 

sediments, and surface waters?

(3) How does mercury cycling vary 

within different geographic regions of the United States (e.g., 

south Florida, Great Lakes, 

northeast, or west)? How might 

the variability be accounted for 

(e.g., resource types (wetlands), 

temperature regimes, microbial 

communities)? 

Finally, the RFA presented three 

specific research objectives:

(1) The performance of theoretical 

and laboratory investigations 

focused on understanding the 

behavior of mercury in the 

environment, including mercury cycling models; the role 

of biogeochemistry, especially 

mercury sulfide complexes; 

interactions among nutrients, 

carbon, and sulfur on methylation processes; the role of 

microorganisms; and the role of 

macrophytes, periphyton, and 

their interactions with hydrological processes.

(2) The development and evaluation of biogeochemical models 

of the microbial transformations 

of mercury in ecosystems in order to interpret the sources and 

distributions of total mercury 

and methylmercury in terrestrial 

and aquatic systems.

(3) Investigation of hypotheses 

about the regional behavior of 

mercury, extrapolating microbiological and biogeochemical 

process data from experimental 

scales to ecologically meaningful scales and time periods.

EPA awarded nine STAR grants 

under this solicitation. The grant 

titles and principal investigators are 

listed in Table 1 along with the information for two additional grants 

related to this topic. This document 

describes the results of these 11 

research projects. Throughout this 

report, we use the last name of the 

principal investigator to refer to a 

specific research project or research 

team and/or the grant reference 

number that is designated in the first 

column of Table 1.

2.3 Relationship of 1999 Mercury 

STAR Grants to ORD MYPs

As described in Section 1, one of 

the primary purposes of the STAR 

program is to provide grants and 

fellowships to scientists for research 

that addresses the goals outlined in 

the ORD MYPs. The mercury fate and 

transport STAR grants pertain primarily to the Mercury Research Multi-Year 

Plan (EPA 2003b). The research results 

obtained from these STAR grants 

relate most directly to Long-term Goal 

(LTG) number two of the Mercury 

Research MYP, which states: 

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Exhibit 1: Major Routes into the Environment


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Exhibit 2: Mercury MYP Research Assists Other MYPs

“to understand the transport and fate 

of mercury from release to the receptor and its effects on the receptor.” 

There are seven general Annual 

Performance Goals (APGs) defined 

by ORD under this LTG, with several 

specific Annual Performance Measures (APMs) for each APG. Specific 

laboratories and centers within ORD 

are designated as the lead for each 

APM. One of the APMs assigned to 

NCER under LTG number 2, “hold 

workshop/state-of-the science on 

mercury with emphasis on fate and 

transport in watersheds and ecosystem impacts,” was accomplished by 

NCER when it held the STAR Mercury Fate and Transport Final Progress Review Workshop in November 

21, 2003, which was attended by all 

FY99 mercury STAR grant recipients 

and other scientists. 

In addition, results from these STAR 

grants will support four key APGs 

and three key APMs from the Mercury MYP (as listed in Table 2). 

Research directed by the Mercury 

MYP will influence and advance the 

goals in other multi-year plans (see 

Exhibit 2). For example, the results 

of the mercury STAR research should 

also play a role in accomplishing the 

general goal of the Air Toxics MultiYear Plan (EPA 2003c) of reducing 

the uncertainty associated with risk 

assessment of air emissions of an important toxic air pollutant, mercury. 

2.4 Research Themes

Several overarching research themes 

were embodied by the specific research questions and objectives listed in the original RFA and presented 

in Section 2.2. Each investigator 

developed preliminary objectives for 

their proposed research that could 

be linked to one (or more) RFA 

research objective(s). The actual 

research results generated through 

the funded projects are more easily 

discussed according to their contributions to four somewhat more specific research areas or themes, which 

are described below. The first three 

of these relate to critical research 

question 2, while the last one relates 

to critical research question 1.

(1)	 Biogeochemical controls of

mercury transformations.

Research on factors that influence transformations among 

mercury species, including 

biologically and chemically mediated processes involving both 

organic (i.e., methylation and 

demethylation) and inorganic 

forms of mercury (i.e., reduction and oxidation of inorganic 

mercury).

(2)	 Biogeochemical controls of

mercury cycling in the environment. Research on factors 

that affect phase partitioning and 

inter-media transport of mercury 

in both aquatic and terrestrial 

systems, such as transfer between sediment and surface water, surface water and the atmosphere, soil and the atmosphere, 

and interchange between the 

atmosphere and plants through 

various processes.


Table 1. FY99 STAR Grants on Transport and Transformation 

of Mercury in Aquatic and Terrestrial Environments 

Grant 

Ref. No. 

in this 

Report

EPA STAR 

Grant No. Grant Titlea Principal Investigator Co-investigatorsa

[1] R827629 Watershed influences on transport, fate, and 

bioavailability of mercury in Lake Superior

Hurley, James P., 

Univ. of Wisconsin 

D. E. Armstrong, Richard C. 

Back, M. M. Shafer, Helen 

Manolopoulos

[2] R827630 Methylmercury sources to lakes in forested 

watersheds: has enhanced methylation increased 

mercury in fish relative to atmospheric deposition

Swain, Edward B., 

Minnesota Pollution 

Control Agency 

Jim Almendinger, Jim Cotner, 

Daniel Engstrom, Jeff Jeremiason, 

Edward Nater

[3] R827631 Responses of methylmercury production and 

accumulation to changes in mercury loading: a 

whole-ecosystem mercury loading study

Gilmour, Cynthia C., 

Academy of Natural 

Sciences and Univ. of 

Maryland

Andrew Heyes, Robert P. Mason, 

John M. Rudd

[4] R827632 Cycling of mercury in Saginaw Bay Watershed Nriagu, Jerome, 

Univ. of Michigan 

Gerald J. Keeler, John Lehrnan, 

Steve Lindberg, Xia-Qin Qang, 

Hong Zhang

[5] R827633 Chemical and biological control of mercury 

cycling in upland, wetland, and lake ecosystems 

in the northeastern United States

Driscoll, Charles T., 

Syracuse University

Ronald Munson, Robert Newton, 

Joseph Yavitt

[6] R827634 Processes controlling the chemical speciation and 

distribution of mercury from contaminated mine 

sites

Brown, Gordon Jr., 

Stanford University

Daniel Grolimund, Mae Sexauer 

Gustin, Trevor R. Ireland, 

Christopher S. Kim, James J. 

Rytuba, Greg Lowry, Samuel 

Shaw, Stephen B. Johnson, Aaron 

S. Slowey

[7] R827635 Mercury and methylmercury cycling in the coastal/

estuarine waters of Long Island Sound and its 

river-seawater mixing zones

Fitzgerald, William F., 

Univ. of Connecticut

Pieter T. Visscher, Prentiss 

H. Balcom, Chad R. 

Hammerschmidt, 

Carl H. Lamborg

[8] R827653 Understanding the role of sulfur in the 

production and fate of methylmercury in 

watersheds

Mason, Robert P., 

Univ. of Maryland

Cynthia C. Gilmour

[9] R827915 The redox cycle of mercury in natural waters Morel, Francois M., 

Princeton Univ.

N/A

[10]b R825249 Light induced mercury volatilization from 

substrate: Mechanisms responsible and in situ occurrence

Gustin, Mae S., 

Univ. of Nevada-Reno

N/A

[11]b R827622E02 Determining the role of plants and soils in 

the biogeochemical cycling of mercury on an 

ecosystem level

Gustin, Mae S., 

Univ. of Nevada-Reno

Ray Alden, James Coleman, 

Dale W. Johnson, Steve Lindberg

a The grant title and co-investigators listed in this table correspond to the information provided in the final summary report for each project and may differ from 

the title and staff listed on the original grant proposal.

b The two grants lead by Gustin were funded under different RFAs (Exploratory Research - Air Chemistry & Physics (1996) and Experimental Program to Stimulate 

Competitive Research (1998)).

(3)	 Sources and distribution of

mercury in terrestrial and

aquatic systems. Research on 

sources of inorganic mercury 

and methylmercury as well as 

factors influencing their distribution in the environment that 

are not already covered under 

the first or second themes (e.g., 

inputs via atmospheric deposition, inputs from runoff and 

erosion).

(4)	 Bioavailability and bioaccumulation of mercury in

aquatic systems. Research on 

the availability of mercury for 

uptake by aquatic organisms and 

the factors that influence this 

availability and accumulation.

In addition, STAR grant researchers developed several new analytical methods for measuring total or 

speciated mercury to facilitate their 

major research objectives. These 

new methods are discussed as a fifth 

research theme.

Table 3 presents a crosswalk of the 

11 STAR grant projects and these five 

research themes. Project results as they 

relate to the aforementioned themes 

are described in the next section.


Table 2. Crosswalk of STAR Grant Contributions to 

 ORD’s Mercury Research Goals and Measures 

Mercury Fate and Transport STAR Research Grants

Annual Performance Goals and Measures 

under LTG 2 of the Mercury MYP (EPA 2003b; 

target year, assigned lab/center for APMs)

Hurley [1]Swain [2]Gilmour [3]Nriagu [4]Driscoll [5]Brown [6]Fitzgerald [7]Mason [8]Morel [9]Gustin [10 and 11]

APG: Provide an assessment of key fate and transport 

issues for tracking the fate of mercury from sources to 

concentrations in fish tissue (2004)

l l l l l l l l l l

APM: Hold workshop/SOS on mercury with emphasis 

on Fate and Transport in watershed(s) and ecosystem 

impacts (2004)

l l l l l l l l l l

APM: Evaluate mercury cycling in complex ecosystems; 

including, air/water interface to accurately assess 

TMDLs for mercury and predict methylmercury 

concentrations in water and fish. Focus is on human 

exposure as the ecological endpoint (2004, NCER)

l l l l l l l l l l

APG: Develop a model for tracking mercury from 

deposition to concentrations in fish tissues (2006) l ❍ l ❍ l ❍ ❍ ❍ ❍

APG: Develop information on sources of mercury 

emissions including the regional/global atmospheric 

fate and transport of such emissions (2008)

l l l ❍

APG: Produce an integrated multimedia modeling 

framework for understanding mercury fate from source 

to fish concentrations (2010)

❍ ❍ ❍ ❍ ❍ ❍ ❍ ❍ ❍ ❍

APM: Develop an integrated multimedia modeling 

framework for the complete scientific understanding 

of mercury fate/transport and atmospheric chemistry/

processes (2010, NERL, NRMRL, NCER)a

❍ ❍ ❍ ❍ ❍ ❍ ❍ ❍ ❍ ❍

Key: l = significant contribution based on available results; ❍ = results provide background knowledge that may be useful in meeting goal/measure. a This APM pertains most directly to the 2001 RFA; although as stated the FY 99 RFA grants provide supporting materials.

Table 3. Crosswalk of Research Themes and Mercury STAR Grant Research Projects

Mercury Fate and Transport STAR Research Grants

Research Themes

Hurley [1] Swain [2]Gilmour [3]Nriagu [4]Driscoll [5]Brown [6]Fitzgerald [7]Mason [8]Morel [9]Gustin [10 and 11]

Biogeochemical controls of mercury transformations (see Section 3.1) l l l l l l l

Biogeochemical controls of mercury cycling 

in the environment (see Section 3.2) l l l l l l l

Sources and distribution of mercury in 

terrestrial and aquatic systems (see Section 3.3) l l l l l l l

Bioavailability and bioaccumulation of mercury 

in aquatic systems (see Section 3.4) l l l l l l l

New methods for mercury analysis (see Section 3.5) l l l l l

Key: l = Research results applicable to research theme.


The important results of the 11 STAR 

research projects on mercury fate 

and transport are described in this 

section. This discussion is organized 

according to the five research themes 

described in Section 2.4. In general, 

this section focuses on the study 

results, but details regarding the 

study approach are also included 

to facilitate interpretation of the 

results. Detailed descriptions of 

the research conducted under each 

individual grant can be found in the 

final summary reports prepared by 

the principal investigators and in the 

references listed in the Appendix. 

Throughout this report, the principal 

investigator’s name and/or grant reference number presented in Table 1 are 

used to identify research conducted 

under the different grants, even when 

co-investigators may have led the 

specific part of research that contributed to the result. This format is used 

to simplify the presentation of the 

grant findings. The primary sources 

used to compile this report were the 

final reports prepared by researchers 

and submitted to NCER. Presentation materials from the November 

21, 2003, STAR Mercury Fate and 

Transport Final Progress Review 

meeting were also used.2 In addition, 

Dr. Robert Mason of the University of 

Maryland, one of the grant principal 

investigators, provided an informal 

summary of what he believes to be the 

most important findings of the nine 

grants, and his input is reflected in this 

synthesis report as well.

3.1 Biogeochemical Controls of 

Mercury Transformations

STAR grant research included several 

studies that investigated the processes by which mercury transforms 

from one species to another. These 

ranged from ecosystem-level studies, in which the overall speciation 

of mercury was analyzed, to laboratory-scale experiments, in which the 

mechanisms of transformation were 

investigated. Results described here 

include new information on factors 

affecting the methylation/demethylation of mercury and inorganic 

transformations between elemental 

mercury (Hg(0)) and divalent mercury (Hg(II)).

Methylation/Demethylation

Net MeHg production in aquatic ecosystems is a complex process because 

there is rapid recycling between 

inorganic Hg (Hg(II)) and MeHg, 

as Hg is effectively methylated in 

low oxygen environments by sulfate 

reducing bacteria (SRB) in sediments, 

but MeHg is also efficiently demethylated in both anoxic and oxic environments. The recent STAR research 

suggest that the turnover time of 

MeHg in aquatic sediments is on the 

order of days to weeks and that Hg is 

likely recycled many times by these 

processes between the two forms 

before being bioaccumulated into 

fish as MeHg, or lost from the system 

as Hg(II), elemental Hg (Hg(0)) and 

MeHg by other processes. 

STAR grant research demonstrated 

that rates of mercury methylation 

and demethylation are affected by a 

variety of factors that influence both 

the availability of environmental 

mercury and the activity of SRB that 

are known to methylate inorganic 

mercury. Results from methylationrelated research performed under 

STAR grants include information 

on factors that increase methylmercury production rates, the 

pathway of reaction in SRB methylation, and demethylation rates of 

methylmercury. 

In general, there has been more 

study of methylation than of demethylation, and much of the STAR 

research conducted in this area also 

focused on factors that affect methylation of inorganic mercury. Study 

results showed that methylation rates 

increase with:

• Increased bacterial metabolic 

rates in summer [7];

• Higher concentrations of 

inorganic mercury in sediment 

pore water [7];

• Increased levels of bioturbation 

by benthic infauna [7];

• Increased mercury load to the 

ecosystem [3]; and

• Increased sulfate load to a 

wetland [2, 8].

3. Study Results

2 One exception is noted. Information on research conducted by Hines and Brezonik under Swain’s STAR grant that was not included in the final 

summary report was obtained from separate pre-publication drafts of journal articles. 


Additionally, research demonstrated 

that methylmercury production rates 

decrease with:

• Increased concentrations of organic matter in sediments (which 

binds inorganic mercury, limiting 

its bioavailability) [7];

• Higher sediment distribution coefficient (KD) values for inorganic 

mercury [7]; and

• Increased sulfide loads in the 

ecosystem [2, 7, 8].

STAR grant research also showed that 

methylmercury demethylation can 

occur via photolytically-induced reactions [2], and that demethylation is a 

major removal process of methylmercury, along with burial [1]. Additionally, researchers measured biological 

demethylation in sediments and 

related it to the net methylation rate 

[3, 8]. The research supporting these 

findings are described below.

Fitzgerald et al. [7] performed 

multiple experiments in Long Island 

Sound using mercury isotope tracers to identify factors that increase 

methylation in sediments. Methylation rates were enhanced in August 

relative to March and June, illustrating the importance of methylating 

bacteria in utilizing available substrate, 

Hg(II). In areas where organic matter 

was lower, there were higher Hg(II) 

concentrations in the porewater 

and correspondingly higher rates of 

methylation. Conversely, the presence of organic matter (i.e., dissolved 

organic carbon, or DOC) resulted in 

decreased methylation rates by binding available inorganic mercury and 

making it less available to the bacteria. 

Sediment-water partition coefficients 

(KDs) for inorganic mercury also were 

measured, and higher coefficients 

were correlated with lower methylation rates. Additionally, increased levels of bioturbation by benthic infauna 

appear to extend the zone of active 

mercury methylation in sediment and 

might move buried mercury to the 

active methylating zone. 

Other STAR research demonstrated 

that increasing mercury or sulfate 

loads to an ecosystem can increase 

methylmercury formation. Gilmour 

et al. [3] report their study to be 

the first direct set of whole ecosystem-level experiments to show the 

rapid and linear increase in methylmercury production in an aquatic 

ecosystem in response to increasing 

surface loading of inorganic mercury. 

Response time (i.e., time between 

increase in mercury load and a measured change in methylmercury production) varied with environmental 

conditions. For shallow, warm sites 

in the Florida Everglades, response 

time was days to weeks, depending on temperature. Response time 

for the Experimental Lakes Area in 

Canada ranged from weeks for small, 

shallow enclosures, to about a year 

for the whole lake. Time to reach a 

new equilibrium for the increased 

mercury loading rate (about four 

times higher than normal mercury 

loading) in Experimental Lakes Area 

is expected to be at least three years.

Gilmour et al. [3] noted that the 

magnitude of the change in methylmercury production due to mercury 

loading depended on when and 

where the loading occurred. Mercury added to the lake surface was 

readily methylated in lake sediments. 

Mercury loaded to uplands and 

wetlands was slow to move to sites 

of methylation (i.e., subsurface areas 

near the water table). Mercury mass 

budgets calculated from isotope 

measurements suggest that most of 

the methylmercury in the lake was 

formed in the lake rather than being 

transported to the lake from other 

parts of the watershed.

Sulfate concentrations influence 

environmental methylation rates 

because SRB require sulfate to break 

down organic matter to yield sulfide, 

carbon dioxide, and energy, with 

concomitant methylation of Hg(II). 

Swain et al. [2] applied sulfate to 

the surface of a naturally sulfatepoor wetland and to lake sediment 

samples. Sulfate sprayed onto half of 

a 2-ha peatland at the Marcell Experimental Forest in northeastern Minnesota increased the annual sulfate load 

by approximately four times relative 

to the control half of the wetland. 

Measurements of sulfate and methylmercury in the wetland two weeks 

after the first application indicated 

that methylmercury porewater concentrations had increased three-fold. 

Three subsequent sulfate additions 

that year did not further increase 

methylmercury porewater concentrations; methylmercury concentrations 

in the treated portion of the wetland 

remained elevated relative to the 

control. Methylmercury concentrations in the outflow from the wetland 

increased following each addition. 

Based on these results, Swain et al. 

[2] concluded that methylmercury 

production in sulfate-poor wetlands 

is as much a function of atmospheric 

deposition of sulfate as of mercury.

The addition of sulfate to lake 

sediment samples in the laboratory, 

however, reduced mercury methylation rates [2]. Addition of glucose 

and ammonium to those sediment 

samples also reduced net mercury 

methylation rates. Swain et al. [2] 

suggested that these additions probably reduced the bioavailability of mercury through the binding of sulfides, 

reduced pH, increased formation of 

mercury-Cl species, or a combination 

of these factors. Pyruvate treatments 

were the only additions to the lake 

sediment samples that increased net 

mercury methylation rates. Swain et 

al. [2] suggested that this occurred 

because the SRB were fermenting pyruvate instead of sulfate and 

were not producing the sulfides that 

reduce mercury bioavailability. In a 

set of sediment cores taken across a 

transect in the lake, methylmercury 

concentrations were negatively cor-


related with total inorganic sulfides. 

Swain et al. [2] proposed that 

bioavailability of mercury, which can 

be reduced by binding with sulfides 

(see Section 3.2), may control the 

distribution of methylmercury in the 

lake sediments.

Mason et al. [8] designed their experiments to evaluate how different 

types of mercury-sulfide complexes 

affect the rate of mercury methylation 

by SRB. In the presence of dissolved 

sulfide, mercury forms both neutral 

complexes (e.g., HgS0, Hg(SH)2

0) and 

charged complexes (e.g., Hg(SH)+, 

HgS2

2-). Mason et al. [8] hypothesized 

that SRB bioconcentrate neutral mercury-sulfide complexes via passive 

diffusion, but do not bioconcentrate 

charged mercury-sulfide complexes. 

Thus, the neutral mercury-sulfide 

complexes are believed to be the 

bioavailable form by which inorganic 

mercury can enter these bacteria cells 

and undergo methylation. Mason 

et al. [8] measured a relatively high 

octanol-water partition coefficient 

(KOW) for two neutral mercury sulfide 

complexes. Based on estimated cell 

membrane permeability for compounds with similar KOW values, 

Mason et al. [8] estimated that SRB 

could uptake these complexes at 

rates that are more than sufficient 

to achieve observed methylation 

rates in cultures, both pure and field 

cultures. 

The fraction of mercury-sulfide 

complexes in sediments that are 

neutral is a function of the sulfide 

concentration in the medium. 

Chemical complexation modeling by 

others has shown that the speciation of mercury tends to shift toward 

charged complexes as sulfide levels 

increase. Mason et al. [8] conducted 

mesocosm experiments in various 

locations in the Florida Everglades 

and observed that as sulfide concentrations increased, methylation rates 

decreased. The investigators also 

compared modeled uptake rates, 

assuming passive diffusion of neutral 

mercury-sulfide complexes (HgS0 and 

HOHgSH0), and compared them to 

observed methylation rates in pure 

cultures of the SRB Desulfobulbus 

propionicus. Uptake and methylation 

rates both increased with increasing HgS0 concentration with similar 

slopes. Mason et al. [8] believe that 

these findings confirm that speciation 

of mercury in the medium surrounding methylating bacteria is a key 

factor affecting mercury methylation 

rates in sediments. 

Mason et al. [8] also observed that 

sulfate levels affect mercury methylation rates in sediments. Mason et al. 

studied sediment cores taken from 

various regions of the Florida Everglades that differ in sulfate concentrations. When the investigators added 

sulfate to cores with low sulfate 

levels, they observed an induction of 

SRB activity and increased mercury 

methylation, even though that activity 

increased sulfide levels. Confirming previous findings, they found 

that adding sulfide alone inhibited 

methylation. Overall, the results of 

both the laboratory and field studies 

indicated that the balance between 

sulfate levels, which affect SRB 

activity, and sulfide production and 

accumulation, which reduce mercury bioavailability (see Section 3.2), 

affects net mercury methylation rates.

One STAR grant indicated that bacterial community structure also can 

affect mercury methylation rates. 

Previous research had shown that the 

acetyl-coenzyme A pathway is key for 

producing methylmercury in several 

SRB strains. Morel et al. [9] hypothesized that some methylation can 

also occur via different pathways in 

other strains. In a series of laboratory 

experiments on seven strains of SRB, 

Morel et al. used carbon monoxide 

dehydrogenase (CODH) activity to 

indicate activity of the acetyl-CoA 

pathway and chloroform to inhibit 

the acetyl-CoA pathway. They found 

four incomplete-oxidizing SRB strains 

that clearly do not utilize the acetylCoA pathway for mercury methylation and, therefore, might not require 

vitamin B12 as a coenzyme.

Demethylation of methylmercury is 

important in determining the net formation of methylmercury in an ecosystem. In Spring Lake in northern 

Minnesota, Hines and Brezonik [2] 

measured methylmercury photodegradation rate constants of 0.0025 hr-1

in the field. They estimated that photolytically-induced demethylation 

accounts for twice as much removal 

of methylmercury from surface 

waters as does methylmercury burial 

in sediments. The depth profile of 

methylmercury photodegradation 

showed loss of methylmercury up 

to 50 to 60 cm, which corresponded 

to the depth of 90 percent attenuation of PAR (photosynthetic active 

radiation). Using a mass-balance 

approach to estimate the distribution and cycling of mercury species 

in Lake Superior, Hurley et al. [1] 

concluded that both sediment burial 

and photo-demethylation are major 

removal processes for methylmercury 

in surface waters. 

In addition to photodemethylation, 

biological demethylation is also an 

important process, as demonstrated 

by Mason et al. [8] and Gilmour et al. 

[3]. Results showed that methylmercury speciation and bacterial activity 

influence the rate of demethylation 

in sediments and thus also affect the 

overall net production of methylmercury in the system. Therefore, 

demethylation rates vary across ecosystems. The researchers concluded 

that studies should measure both 

methylation and demethylation rates 

to properly understand the rate of 

net methylmercury production.

Inorganic Transformations

STAR research also focused on transformations between inorganic species of mercury via redox reactions, 
10

with a particular focus on photoinduced reactions as an important 

pathway. These redox transformations can affect the fate and transport 

of mercury by converting mercury 

to chemical species that are more or 

less mobile in the environment. This 

section presents STAR grant findings 

related to rates of transformation, 

chemical mechanisms, and factors 

affecting these transformations. 

Research focused more on the overall 

transfer of mercury between media 

(including transfers that might occur 

due to redox reactions involving inorganic mercury transformation) are 

covered in the Section 3.2.

STAR grants showed that oxidation of 

Hg(0) to Hg(II): 

• Occurs via photo-oxidation, 

perhaps indirectly via hydroxy 

radicals and other reactive chemical intermediates [2, 9];

• Can at times account for the majority of Hg(0) lost from surface 

waters [9]; and

STAR grants also demonstrated that 

reduction of Hg(II) to Hg(0):

• Can occur via photoreduction [2];

• May require the Hg(II) to be 

bound to organic matter [7]; and

• Can occur at a higher rate than 

photooxidation [2].

In some experiments, however, 

both oxidation and reduction were 

observed under similar conditions 

[4]; additional research is needed to 

understand such reactions.

Both Morel et al. [9] and Hines and 

Brezonik [2] found photooxidation 

of Hg(0) to be a significant means of 

Hg(0) removal from surface waters. 

In laboratory and field studies on 

the St. Lawrence River, Morel et al. 

[9] found the rate of loss of Hg(0) 

to photooxidation to be about two 

orders of magnitude greater than 

its loss via volatilization to the air. 

Through treatment with heat, chloroform, and filtration to eliminate SRB 

activity, Morel et al. [9] confirmed 

that the oxidation is chemically 

mediated chiefly by UV radiation 

rather than biologically mediated. 

Their data suggested that factors 

other than photon flux, however, 

are rate-limiting. They proposed that 

hydroxy radicals and reactive halogen 

and oxygen intermediates may favor 

Hg(0) oxidation. 

In water samples from Spring Lake 

in northern Minnesota, Hines and 

Brezonik [2] calculated a pseudofirst order rate constant of 0.58 h-1

for the photooxidative loss of Hg(0). 

Measurements of air and water Hg(0) 

concentrations indicated that lake 

water was supersaturated with Hg(0). 

Fluxes of Hg(0) from the lake to the 

air were greatest in warmer, sunnier months, but correlation with 

solar radiation was weak. The loss of 

Hg(0) by photooxidation may have 

exceeded loss by volatilization to air 

at times. Other losses of Hg(0) from 

the lake water were reaction with 

ozone, hydroxyl radical, and possibly singlet oxygen. Burial was also a 

major loss of total mercury.

In studies of the photoreduction of 

Hg(II), Fitzgerald et al. [7] concluded 

that it must be bound to organic 

material to be reduced by abiotic 

processes in natural waters. Their 

research demonstrated that the addition of humic acid to synthetic seawater increases the rate of Hg(II) reduction to Hg(0), as has been reported 

by others. However, Fitzgerald et al. 

also observed that adding ethylenediaminetetraacetic acid (EDTA) and 

chlorine (Cl-) as competitive ligands 

at concentrations that are high 

enough to outcompete the humic 

acid as a ligand resulted in reduction 

rates falling to almost zero. Fizgerald 

et al. [7] concluded that higher Hg(0) 

concentrations in surface water in the 

summer are supported by the higher 

summer concentrations of DOC and 

other ligands that enhance reduction of Hg(II). However, they also 

concluded that DOC may sometimes 

be high enough to prevent reduction 

to Hg(0) and that mercury-organic 

associations can explain the reduction that was measured below the 

photic zone (i.e., dark reductions 

of mercury may be abiological and 

organically mediated). Their data 

suggest that organic ligands and DOC 

affect the speciation of dissolved inorganic mercury complexes in anoxic 

low-sulfide sediments such as those 

in Long Island Sound. That conclusion appears to contrast with existing 

chemical speciation models that indicate sulfide to be the major ligand of 

both dissolved methylmercury and 

Hg(II) in natural porewaters.

Data collected by Nriagu et al. [4] 

also suggest that photoreduction 

of Hg(II) to Hg(0) is important 

in surface waters. They measured 

diurnal cycles of dissolved gaseous 

mercury (DGM; dissolved gaseous 

Hg(0)) concentrations in Great Lakes 

water to correlate with solar radiation 

both temporally and spatially. Their 

calculations indicated that the Hg(0) 

was supersaturated at most locations, 

a finding consistent with those of 

Hines and Brezonik [2]. In addition, 

Nriagu et al. [4] investigated the photoreduction of Hg(II) by examining 

the role of amino acids in the redox 

cycle of mercury. Preliminary experiments demonstrated that Hg(II) was 

photo-reduced in the presence of 

eight amino acids, but not photoreduced in the presence of at least 

four other amino acids. Hines and 

Brezonik [2] also measured photoreduction of Hg(II) in the lake and 

found the rate constant to be 1.1 h-1, 

which was twice the rate constant 

for photooxidation. They concluded 

that redox reactions between Hg(0) 

and Hg(II) readily occur, providing a 

source of ionic mercury to the system 

that can be methylated.
11

These STAR results indicate that 

both oxidation and reduction of 

inorganic mercury species appear 

to occur readily in natural waters 

under the right conditions. Additionally, in surface water samples 

spiked with Fe(III) and exposed to 

sunlight, Nriagu et al. [4] found that 

DGM increased in some samples 

and decreased in others. Nriagu et 

al. estimated a rate constant for the 

reduction and proposed chemical 

mechanisms for both reduction and 

oxidation. It is not entirely clear why 

both oxidation and reduction were 

observed under similar conditions. 

Experimental conditions introduced 

by the investigators were proposed 

as one explanation; however, further 

research is necessary.

3.2 Biogeochemical Controls 

of Mercury Cycling in the 

Environment

Mercury cycling (i.e., phase partitioning and inter-media transport) 

is key to the distribution of mercury 

in the environment. Most of the 

STAR research in this area focused 

on processes controlling mercury 

transfer between surface water and 

the atmosphere or surface water 

and sediment as summarized in the 

subsections below.

Sediment—Surface Water Transfers

As MeHg formation appears to occur 

primarily in sediments, the processes 

responsible for its transport from the 

sediment into the water column are 

very important in determining its rate 

of bioaccumulation into fish. Partitioning of mercury between dissolved 

and solid phases in surface water 

determines the mechanisms by which 

it can be transferred to other media 

(e.g., mercury bound to particles in 

the water column will be subject to 

deposition and to resuspension from 

the benthic sediment, whereas dissolved mercury complexes will not; 

some dissolved complexes may be 

bioavailable whereas other are not). 

Fitzgerald et al. [7] found that most 

of the ionic inorganic mercury dissolved in the Long Island Sound 

water column is complexed with 

dissolved organic matter with a high 

affinity, given that the conditional stability constants were very high (log K’ 

of about 21 to 24). Principal sources 

of organic ligands in Long Island 

Sound appear to be influx from river 

water as terrestrial organic matter 

(47 percent) and phytoplankton 

exudation (31 percent). The primary 

loss of organic ligands is via tidal 

exchange with low DOC/low ligand 

waters of the continental shelf. The 

Connecticut River is the primary 

source of organic matter to Long 

Island Sound, which is reflected in 

the seasonal variation in the ligand 

abundance (highest in summer/

spring, lowest in winter).

Fitzgerald et al. [7] studied factors 

affecting the fraction of mercury in 

more labile forms which can potentially increase the rates of methylmercury formation. In the mixing 

zone of the Connecticut River and 

the Long Island Sound, Fitzgerald et 

al. found that the fraction of mercury 

present as reactive mercury increased 

from levels present in river water 

upstream of the mixing zone. The 

reactive fraction was primarily in 

the particulate phase. On the basis 

of field measurements and a simple 

competitive ligand model, Fitzgerald 

et al. concluded that the enhanced 

reactive mercury in the estuary is a 

result of dilution of the dominant 

organic ligand class (characteristic of 

fresh water mercury) and competition with chloride (from Long Island 

Sound) in the mixing zone, followed 

by coagulation/adsorption onto suspended particles.

Fitzgerald et al. [7] found both sedimentary organic matter and acid-volatile sulfide to affect partitioning of 

methylmercury and Hg(II) between 

particulate and dissolved phases 

in waters of Long Island Sound. As 

expected on the basis of bioavailability of Hg(II) to SRB, Fitzgerald et al. 

[7] found mercury methylation rates 

to vary inversely with the KD of Hg(II) 

and positively with the concentration 

of Hg(II) in porewater (which exists 

mostly as HgS0). 

Mason et al. [8] also explored how 

mercury partitioning to sediment 

particles impacts mercury concentration and bioavailability in sediment 

pore waters. They developed a model 

accounting for adsorption and sulfide 

concentration to predict Hg(II) concentrations in sediment pore water 

and sorbed to particles. The model 

was applied to two ecosystems, the 

Patuxent River and the Florida Everglades, and the results were validated 

with measurements. The model’s 

predictions of neutral mercury-sulfide complexes correlated well with 

the methylmercury concentrations 

for the samples.

Binding of mercury to dissolved 

organic carbon (DOC) in surface 

waters has been found to be important, but the findings of Mason et al. 

[8] suggest that the influence of DOC 

on mercury partitioning in sediments 

is more complex. Their research 

suggested that binding of mercury to 

DOC is less important in sediment 

pore waters under typical DOC levels 

and when there is greater than 0.01 

micromolar sulfide. Mercury binding with DOC in the solid phase is 

important, and these studies [8] 

suggest that in aerobic sediments, 

binding with the organic fraction of 

sediments is more important than 

binding with methyl oxide phases. 

Surface Water—Atmospheric Transfers

Deposition of mercury from the 

atmosphere to surface waters (and 

soils and plants) occurs via both wet 

and dry deposition. Dry deposition 

is less well studied than wet, and 

includes deposition of atmospheric 

particulate Hg and gas phase Hg. Gas 

phase Hg consists of both Hg(II), 

often called reactive gaseous mercury 

(RGHg), and Hg(0). Dry deposition 
12

of RGHg depends on its chemical 

composition (e.g., HgBr2, HgCl2), 

wind speed, and other factors. Less 

well studied is the loss of mercury 

from surface waters via volatilization 

of Hg(0) to the atmosphere. Several 

STAR grants investigated the relative 

importance of Hg(0) volatilization 

from surface water in the total mercury mass balance.

Estimating net dry deposition of 

Hg(0) to surface water (and soils) is 

complex because Hg(0) is constantly 

degassing (i.e., volatilizing) from the 

surface. As described above, Hines 

and Brezonik [2] and Nriagu et al. [4] 

found Hg(0) concentrations in lakes 

typically to be above saturation with 

respect to the atmosphere, indicating 

that dry deposition of Hg(0) will not 

occur; rather, Hg(0) will escape from 

the surface water into the atmosphere. Both Hurley et al. [1] and 

Fitzgerald et al. [7] found volatilization of Hg(0) from surface water to 

be a major process by which Hg(0) 

leaves surface water. Based on a mass 

balance of mercury derived for Lake 

Superior, Hurley et al. suggested that 

volatilization of Hg(0) is one of the 

main removal processes for total mercury from that ecosystem, followed 

by sediment burial. Fitzgerald et al. 

found volatilization of Hg(0) to be 

important in removing Hg(0) from 

Long Island Sound as well, and noted 

that this transfer is a potentially 

important source of atmospheric 

Hg(0). Fitzgerald et al. estimated 

annual emissions of Hg(0) from Long 

Island Sound to be 80 kg in total. 

This corresponds to 35 percent of 

the annual input of mercury to Long 

Island Sound being remobilized into 

the atmosphere as Hg(0). Fitzgerald 

et al. concluded that volatilization 

competes with sediment methylation of mercury as a mechanism of 

removal of Hg(0) from surface water. 

Results from a short-term study of a 

Tobico marsh at the coast of Saginaw 

Bay suggested to Nriagu et al. [4] 

that loss of Hg(0) from that marsh 

resulted primarily from volatilization 

of Hg(0). They estimated a “lifetime” 

of about one hour for DGM in the 

marshlands. Volatilization of Hg(0) 

may be more or less important than 

other processes that remove Hg(0) 

from surface waters in other types 

of ecosystems. As stated earlier, 

results from Nriagu et al. indicate 

that mercury cycling between surface 

water and air varies diurnally and 

seasonally, with rather large daily 

variations in DGM in surface water of 

Saginaw Bay. Bay waters were generally supersaturated with Hg(0) relative to the partial pressure of Hg(0) 

in the atmosphere. Distinct diel 

variations in DGM concentrations 

were observed that indicate a high 

dependence on the intensity of solar 

radiation. In their research, Nriagu 

observed a maximum DGM in surface 

water samples around noon, with 

DGM levels generally highest in June. 

These variations seem to indicate that 

the dissolved mercury pool is being 

actively recycled in the water column.

Soil—Atmospheric Transfers

Deposition of mercury from the 

atmosphere to surface soils, which 

occurs via both wet and dry deposition, has been extensively studied. 

Less well studied is the transfer of 

mercury from soils back into the 

atmosphere. Several STAR grants 

investigated factors affecting the 

emission of mercury from soils to the 

atmosphere. 

Gustin et al. [10] performed laboratory, field, and mesocosm experiments to investigate light-enhanced 

mercury volatilization. In laboratory 

experiments, a well-mixed single pass 

gas exchange chamber was used to 

test various hypotheses about mercury volatilization from soil. All mercury species and mercury-containing 

substrates constantly emitted mercury to the air. However, of the pure 

phase mercury species, only mercury 

sulfide exhibited light-enhanced 

emissions. In a test of whether 

mercury adsorbed to iron oxides 

and organic matter would exhibit 

light-enhanced emissions, iron oxide 

amended with mercury chloride, and 

organic material amended with Hg(0) 

and mercury chloride exhibited lightenhanced emissions. 

At various mining, Superfund, and 

geothermal sites in Nevada and 

California, Gustin et al. [10] used 

chamber and micrometeorological 

flux measurements to assess lightenhanced mercury emissions. At all 

sites, light-enhanced emissions of 

mercury were observed with little 

increase in soil temperature. A rapid 

increase in emissions was observed 

in the morning (e.g., at New Idria, 

CA, emissions from soils doubled as 

morning light irradiated the soil), 

and on cloudy days the flux would 

mimic the shadowing and lighting 

of the sites. A 100-fold increase in 

emissions was observed by artificially 

lighting tailings at night at the Carson 

River Superfund site. Gustin et al. 

[10] also calculated the activation 

energy for the mercury flux in dark 

and light conditions. The activation 

energy was lower for the nighttime 

flux, indicating that the light energy 

in the day is being used for a process 

in addition to volatilization, perhaps 

to generate elemental mercury by 

photoreduction.

In follow-up field work, Gustin et al. 

(under the Brown et al. [6] grant) 

compiled data on mercury emissions from mine wastes and surrounding mercury-enriched terrains 

by measuring fluxes at 18 areas of 

mining activity. They studied the different emissions under light versus 

dark conditions and determined the 

speciation of the mercury sources 

at each site. Overall, the researchers 

found that mercury-contaminated 

mining sites with higher concentrations of soluble mercury-containing 

species emit more mercury to the 

atmosphere than those containing 

less soluble mercury-containing species. Specifically: (1) as particle size 

decreases, mercury concentration, 
13

mercury emissions, and light-to-dark 

emission ratio increase; (2) metacinnabar-bearing samples exhibit higher 

light-to-dark emission ratios than 

those containing cinnabar; and (3) 

the presence of more soluble mercury 

species also correlates positively with 

higher light-to-dark emission ratios. 

Gustin et al. [11] used large mesocosms (Ecologically Controlled 

Lysimeter Laboratories (EcoCELLs)), 

smaller plant-exposure chambers 

(ecopods), and a single-plant gasexchange system to further study 

the role of soils (and plants, see next 

section) in controlling the fate and 

transport of mercury in the environment at the ecosystem level. Measurements of mercury emissions from 

the soil were made in the EcoCELLs. 

Dominant factors driving mercury 

emissions from soil were incident 

light, precipitation, and the presence 

of vegetation. Soil temperature and 

moisture also influenced flux. Data 

indicated that soil gas efflux is not 

a diffusion-driven process, except 

during certain periods of the day 

(particularly midday). Therefore, soil 

gas concentrations are not an effective predictor of soil mercury flux.

Plant Transfers

In the EcoCELL experiments, Gustin 

et al. [11] found that nearly all the 

mercury in plant foliage was derived 

from the atmosphere, as opposed 

to the soil. Mercury concentrations 

in the foliage of trees grown in 

mercury-enriched soil (concentration of 12.3 μg/g) was similar to the 

foliage of trees grown in soil with a 

background mercury concentration 

(0.03 μg/g). These results indicate 

that foliage may act as a significant 

sink for atmospheric mercury, and 

that the mercury content in litterfall would represent a significant 

mercury input to terrestrial ecosystems, consistent with the findings of 

Driscoll et al. [5] described below. 

Additionally, experiments (using 

leaf washing) showed that less than 

three percent of mercury would be 

removed with precipitation. Gustin 

et al. concluded that litterfall may be 

an effective way to estimate mercury 

deposition in deciduous forests. 

Ecopod and gas exchange chamber 

experiments supported these results 

by showing that mercury concentration in foliage increased with increasing air concentrations of mercury, 

and that mercury soil concentrations 

either did not significantly influence 

foliar concentration or influenced 

it much less than the air concentrations. Experiments in the gas 

exchange chamber indicated that 

some mercury is transported from 

the soil via the transpiration stream, 

because there was a pulse in water 

vapor and mercury emitted when the 

chamber was first illuminated.

In their study of two upland forest 

sites (one deciduous, one coniferous) in the Sunday Lake watershed, 

Driscoll et al. [5] also found forest 

vegetation to be very important in 

mediating the inputs of mercury 

to the forest floor. They found that 

the flux of mercury to the forest 

ecosystem was dominated by dry 

deposition (70 percent of total 

deposition, including both litter fall 

and throughfall). Ten-fold higher 

levels of total mercury were reported 

for the throughfall (i.e., deposition 

passing through tree canopy) at the 

coniferous plot, however, than at the 

deciduous plot. The major pathway 

for dry deposition in the coniferous 

plot was throughfall, while the major 

pathway in the deciduous plot was 

litterfall, which is consistent with the 

conclusions from Gustin et al. [11] 

regarding litterfall being an effective 

measure of deposition in deciduous 

forests.

Studies by Nriagu et al. [4] indicate 

that plants are contributing in a 

minor way to the release of mercury from soils to the air. Mercury 

flux measurements over soybean 

plants (3 ng/m2-hr) were higher 

than mercury flux measurements 

over soil with no plants that had 

previously been sown with corn 

seed dressed with mercury (from 

< 0.1 to 1.7 ng/m2-hr). Results of 

these measurements indicated that 

the mercury in the soil from the 

seeds was a minor source of mercury 

to the atmosphere, but that vegetation could play a role in transferring 

mercury from soil to the atmosphere. 

3.3 Sources and Distribution of 

Mercury in Terrestrial and 

Aquatic Systems

Several STAR grants identified the 

major sources and means of distribution of mercury in a given study area. 

Researchers identified and investigated important processes influencing movement of mercury throughout a watershed (e.g., production of 

methylmercury, predominant mercury fate and transport processes). In 

addition, several researchers performed general mass balance studies 

to quantify sources and inputs within 

a given watershed. Relevant results 

are summarized here.

Mass Balance Studies

Using data from open water cruises 

and GIS-based studies, Hurley et al. 

[2] developed a mass balance model 

of mercury in Lake Superior. They 

reported that total mercury inputs 

were dominated by atmospheric 

deposition (58 percent), followed 

by tributary inputs (21 percent) 

and contributions from particulate 

remineralization (19 percent), while 

sources of methylmercury were 

somewhat evenly divided between 

the atmosphere, tributaries, groundwater, remineralization, and in-situ

methylation. Elevated concentrations of mercury in nearshore waters 

appeared to be the result of tributary 

inputs. Hurley et al. [2] estimated 

volatilization of Hg(0) to be the 

principal removal process of total 

mercury from the surface water, 

followed by sedimentation, while 

they estimated methylmercury to be 

removed primarily by sedimentation 

and photo-demethylation. Based 

on trace amounts of methylmercury 

detected in open surface waters 
14

(about 1 percent of total mercury in 

the lake), Hurley et al. [2] concluded 

that wet deposition may be a small 

but measurable source of methylmercury to surface water in the lake. 

They found the proportion of total 

mercury as methylmercury to be 

maximum at the thermocline in open 

surface waters of the lake, probably 

because plankton, with their accumulated methylmercury, aggregate at the 

thermocline. Using their data, Hurley 

et al. [2] developed a simple model 

to relate annual tributary loading 

from a watershed to the lake based 

on land use/type in the watershed. 

Driscoll et al. [5] performed mass 

balance calculations on mercury 

cycling in the Sunday Lake watershed 

of the Adirondacks. Results from 

this study indicate that overall, the 

watershed and lake are sinks for total 

mercury, while the wetlands and the 

lake are sources of methylmercury. 

At both coniferous and deciduous 

upland plots, mass balance calculations indicated that the soil acts as a 

net sink for inputs of total mercury 

and methylmercury. Discrepancies 

in these calculations suggest that 

mercury is either accumulating in 

the forest floor or is lost by volatilization or both. As a part of this 

study, total mercury concentrations 

were measured in soil horizons. 

The concentrations were highest in 

the surface layer, lowest in the next 

layer, peaked slightly in the third 

layer, and then decreased in the 

lowest layer analyzed. In measurements taken in soil water, fluxes of 

total mercury decreased with depth, 

which coincided with decreases in 

DOC. Driscoll et al. [5] concluded 

that mercury is immobilized in the 

mineral soil by the deposition of 

organic matter. 

Swain et al. [2] performed a mass balance study of mercury in the Spring 

Lake watershed in northern Minnesota. They concluded that atmospheric deposition is the main source 

of total mercury to the lake, methylation in the lake is the main source of 

methylmercury, demethylation and 

photodegradation consume most of 

the methylmercury in the lake, and 

Hg(0) is lost from the lake mostly by 

photooxidation and volatilization.

In Situ Methylmercury Production 

As studies described thus far indicate, 

although atmospheric deposition 

and watershed inputs are important 

sources of methylmercury (MeHg), 

in situ production within the aquatic 

system is the dominant source of 

MeHg, at least for freshwater lakes 

(except for those that have a large 

watershed or extensive abridging 

wetlands) and for estuarine systems (e.g., Long Island Sound and 

the Chesapeake Bay). Hines and 

Brezonik [2] performed mercury 

mass balance calculations for Spring 

Lake in northern Minnesota (a lake 

with little wetland coverage in the 

watershed) and found that in situ

methylation is the main source of 

methylmercury, accounting for 76 

percent of input). In Long Island 

Sound, Fitzgerald et al. [7] concluded 

that in situ production is also the 

major source of methylmercury. The 

calculated diffusive sediment-water 

flux of methylmercury in Long Island 

Sound (11 kg/yr) is higher than 

methylmercury inputs to the Sound 

from external sources (estimated at 

5.2 kg/yr). Results from both a model 

developed by Driscoll et al. [5] (i.e., 

the Mercury Cycling Model for Headwater Drainage Lake Systems) and 

the mercury mass balance calculations for the Sunday Lake watershed 

in the Adirondacks indicate that 

both the surrounding wetlands and 

Sunday Lake itself are the dominant 

sources of methylmercury to the lake. 

Gilmour et al. [3] studied methylmercury sources in both the Everglades 

and the Experimental Lake Area in 

Canada. Calculations of mercury 

mass balance based on mercury 

isotope measurements in the systems 

suggested that most of the methylmercury found in both systems is 

formed in situ. Mercury added to surface water was readily methylated in 

the sediments, while mercury loaded 

to uplands and wetlands was slow 

to move to sites of methylation. The 

major locations for new methylation 

in both the shallow, warm sites in the 

Everglades and the deeper, cooler 

Experimental Lakes Area were saturated surface peats and sediments, 

and for the Experimental Lakes Area, 

also anoxic bottom waters. The newly 

deposited mercury was more readily 

methylated than existing mercury 

at those sites (i.e., sediments and 

soils), suggesting that older mercury 

deposits are less bioavailable for 

methylation. 

In Lake Superior, Hurley et al. [1] 

measured sediment-water partition 

coefficients (log KD) for total mercury 

and methylmercury in sediment that 

were relatively low compared to 

other lakes, suggesting that mercury 

is more mobile in Lake Superior sediments than in other lakes. The highest concentrations in sediment were 

observed in spring and near tributary 

inputs. They proposed several factors 

as likely to control methylmercury 

concentrations in sediments, but 

did not acquire quantitative data on 

those factors during this study.

Morel et al. [9] conducted a statistical 

analysis of methylmercury concentrations in marine fish tissues and a 

simple modeling exercise of the Pacific 

Ocean. Statistical analyses of methylmercury concentrations in yellowfin 

tuna caught off of Hawaii in 1971 

and 1998 indicated no statistically 

significant change in methylmercury 

concentrations over this time period. 

By contrast, a simple, three-layer box 

model of the ocean over this time span 

indicated an increase in methylmercury concentrations in surface waters 

if methylmercury were formed in the 

thermocline or the mixed layer (and 
15

therefore were susceptible to changes 

in mercury deposition from the atmosphere). Based on this contrast, Morel 

et al. [9] suggested that methylation is 

occurring in deep waters, sediments, 

or hydrothermal vents where mercury 

concentrations are not greatly affected 

by human activities. 

Inputs via Atmospheric Deposition

As mass balance studies indicate 

(e.g., Hurley et al. [1] and Swain 

et al. [2] results described above), 

atmospheric deposition can be a 

major source of total mercury to 

an ecosystem and can have a direct 

effect on surface water mercury 

concentrations. Based on studies in 

Spring Lake in northern Minnesota, 

Hines, and Brezonik [2] reported 

that atmospheric deposition is the 

primary input of total mercury to the 

lake. Likewise, Hurley et al. [1] found 

that atmospheric deposition accounts 

for 58 percent of the total mercury 

input to Lake Superior. In a direct 

ecosystem-level experiment (see 

Section 3.1 and previous section), 

Gilmour et al. [3] observed a rapid 

and linear increase in methylmercury production within the aquatic 

ecosystem in response to changes 

in mercury load by surface application designed to simulate elevated 

atmospheric deposition. Findings 

from that study also indicate that the 

newly deposited mercury was more 

bioavailable than existing mercury at 

these sites and contributes substantially to methylmercury production 

and bioaccumulation (Section 3.4). 

As described in Section 3.2 under 

Plant Transfers, Driscoll et al. [5] 

analyzed mercury deposition in two 

upland forest sites in the Sunday 

Lake watershed and found 70 percent of the total mercury deposition 

to be via dry deposition. 

Driscoll et al. [5] also sampled sections of sediment cores from eight 

lakes in the Adirondacks for total 

mercury to quantify the increase in 

mercury deposition in that area over 

the last 200 years. Sediment samples 

were dated using 210Pb to measure 

changes in sediment deposition 

throughout this period. On average, 

sites showed a 5.8-fold increase in 

sediment mercury deposition from 

background values (before 1900) to 

peak values (which occurred from 

1973 to 1995). Current mercury sediment deposition has decreased from 

the peak to a level approximately 3.5 

times the background values. Driscoll 

et al. [5] also estimated rates for current and pre-anthropogenic mercury 

deposition in a perched seepage 

lake. Using the sediment deposition 

data, Driscoll et al. concluded that 

retention of mercury in Adirondack 

lakes has decreased over the past 200 

years, but the mechanism for this 

decrease is unclear.

Inputs from Runoff/Erosion

As described in Section 3.3, mass 

balance calculations by Hurley et al. 

[1] indicate that tributary loading is 

a significant input of total mercury to 

Lake Superior, suggesting that runoff 

and erosion can be major sources 

of mercury to a water body. In their 

watershed study of Lake Superior 

involving primarily groundwater 

sampling in the basin, Hurley et al. 

[1] found (among other results) that 

forested areas in a watershed basin 

can be a significant source of methylmercury, apparently generated by in 

situ methylation. Measurements of 

methylmercury in groundwater and 

surface water samples of the Sunday 

Lake watershed indicated to Driscoll 

et al. [5] that the riparian wetland 

was a net source of methylmercury 

to the lake and presumably a site for 

methylation. Furthermore, Driscoll 

et al.’s mass balance calculations 

indicated that fluxes of methylmercury were elevated in waters draining 

riparian wetlands, and that the entire 

watershed and the lake were net 

sources of methylmercury to downstream surface waters. However, the 

mass balance calculations did suggest 

that the upland forest floor may be 

an active zone of demethylation.

On the other hand, mass balance 

calculations by Gilmour et al. [3] 

using measured mercury isotope 

concentrations from their wholeecosystem study in the Experimental 

Lakes Area of Canada show that most 

of the mercury loaded to uplands 

has been slow to move to sites of 

methylation in the lake and wetlands. 

Only a small fraction of the mercury 

loaded to upland areas had migrated 

to the lake after two years; much of 

the deposited mercury remained on 

surface vegetation. Mercury loaded 

to the wetland surface has been slow 

to move to subsurface areas near 

the water table where methylation 

occurs. The differences between 

these observations and those of 

Hurley et al. [1] and Driscoll et al. [5] 

may indicate the site-specific nature 

of inorganic mercury mobility in the 

environment.

Subsurface Transport

As noted above, Hurley et al. [1] 

sampled ground water along a 

stream passing through wetlands and 

forested area in the Lake Superior 

basin. A strong link between methylmercury concentrations in subsurface 

samples and the hydrologic cycle was 

observed, indicating that subsurface 

transport and/or production of methylmercury can be an important source 

of methylmercury to surface waters. 

Hurley et al. [1] proposed that the 

source of the methylmercury from the 

forested areas may be the hyporheic 

zone of streams (i.e., the subsurface 

interface occurring between groundwater and stream water). 

Brown et al. [6] studied how mercury 

from mining activities has entered 

watersheds. The researchers performed column experiments to evaluate the fate of Hg(0) introduced via 

dredge tailings during gold mining 

and subsequent restorations (such as 

the Clear Creek tributary to the Sacramento River in Redding, CA). Varying 

influent ionic strengths and low 

molecular weight organic acids were 

applied to columns filled with tailings 
16

containing mercury in an attempt to 

simulate infiltration through the rooting zone of revegetated mine waste. 

Significant concentrations of mercury 

leached from the tailings in dissolved 

and particle-associated forms. Chemical extractions indicated that Hg(0) 

was transformed into readily soluble 

species such as mercury oxides and 

chlorides (3-4 percent), intermediately extractable phases that likely 

include sorption complexes and 

amalgams (75-87 percent), and fractions of highly insoluble forms such 

as mercury sulfides (6-20 percent; 

e.g., cinnabar and metacinnabar). 

These results imply that Hg(0) is 

transformed into other forms of mercury (soluble mercury and insoluble 

mercury sulfides) during its transport 

from placer gold mining regions 

to major wetlands such as the San 

Francisco Bay Delta. Also the results 

indicate that colloid-associated 

mercury from revegetated mercury 

mine tailings sites may be of potential 

environmental concern, and that the 

gradual infiltration of even very low 

concentrations of organic acids into 

mine tailings will eventually instigate 

conditions where colloid mobilization can occur.

Brown et al [6] also attempted to 

measure the variation of different 

mercury isotopes from a variety of 

natural mining samples to determine 

if any natural variations in the ratios 

may be useful as a mercury source 

indicator. No significant variation 

was found (above the experimental 

error), but further studies were suggested.

3.4 Bioavailability and 

Bioaccumulation of Mercury 

in Aquatic Systems

Bioavailability

Several STAR grant findings related to 

factors that affect the bioavailability of 

mercury to SRB have been discussed 

in the preceding sections on mercury 

transformations and mercury cycling 

in the environment. Additional findings on the topic are described here.

STAR grant researchers found that 

newly deposited mercury is more 

bioavailable than “older” mercury 

and that DOC levels in a water body 

appear to affect bioavailability. In 

both the Florida Everglades and the 

Experimental Lakes Area in Canada, 

Gilmour et al. [3] found that newly 

deposited mercury was more readily 

methylated than existing mercury. In 

the Everglades, Gilmour et al. demonstrated that sulfate and DOC have 

a much larger effect on the methylation of newly deposited mercury 

than on the methylation of mercury 

already present in soils. Additionally, 

new methylmercury production from 

mercury isotope spikes added during 

the experiments was a good predictor of total methylmercury concentrations in surface sediments. After two 

years of mercury loading, Gilmour et 

al. [3] found that about 25 percent of 

the mercury in perch in the Experimental Lakes Area lake originated 

from the isotopically-labeled spike. 

Partitioning of methylmercury to 

biota appeared to decrease with time 

in Hurley’s study. Hurley recorded 

the uptake of methylmercury by 

Selenastrum over time; a partition 

factor of 105 was measured through 

24 hours. After this time, the factor 

decreased to 104. The reason for this 

change is unclear, but may reflect the 

increased algal growth rate.

Hurley et al. [1] found that mercury 

concentrations in biota in Lake Superior are generally higher in the spring 

than summer, a trend that might result 

from bio-dilution by growing phytoand zooplankton during the summer. 

Their results also showed higher concentrations of methylmercury in zooplankton from the epilimnion. These 

observations suggest that mercury 

in surface waters deposited from the 

atmosphere may be more bioavailable 

than mercury in deeper waters. 

Hurley et al. [1] also reported that 

samples of water from Lake Superior 

with higher DOC had lower algal 

concentrations of methylmercury, suggesting that DOC may be complexing 

mercury and making it unavailable for 

uptake. That observation agrees with 

those of Fitzgerald et al. [7] (discussed 

in Section 3.1) that increasing DOC 

reduces methylation rates. 

Driscoll et al. [5] found that bioconcentration factor (BCF) values of fish 

in Sunday Lake in the Adirondacks 

increased with each trophic level. 

Total mercury concentrations in 

yellow perch were also higher than 

values measured in many remote 

regions (locations not specified). 

However, BCF values were generally 

lower for yellow perch in Sunday 

Lake than values for this species in 

other Adirondack lakes. Driscoll et 

al. suggested that this may occur 

because methylmercury in Sunday 

Lake is less bioavailable due to binding with the high concentrations of 

DOC.

Mason’s team [8] evaluated the role 

of pH in mercury-sulfide complexation, given that a number of studies 

have shown an inverse relationship 

between lake water pH and mercury contamination of fish tissues. 

Mason et al. modeled mercury-sulfide 

speciation with changing pH and 

found that at low pH, as sulfide levels 

increase, the tendency for speciation 

of mercury to shift toward charged 

complexes is not as strong as it is 

at higher pH. In other words, when 

sulfide levels are high, neutral mercury-sulfide complexes, which are the 

bioavailable forms (see Section 3.1), 

are more likely to form if the pH is 

low than if the pH is high. This relationship may help explain why fish 

tend to have higher tissue mercury 

residues in waters of lower pH. 

The bioavailability of mercury at 

mining sites could vary based on the 

mercury speciation at the site and 

presence of ions such as chloride and 

sulfate. Brown et al. [6] took samples 

from multiple mining locations in 

California Coast Range mercury 

mineral belt, the placer gold mining 
17

belt in the Sierra Nevada foothills, 

and Nevada and found significant 

variations in mercury speciation from 

site to site. Highly soluble forms 

of mercury, such as the minerals 

corderoite, montroydite, schuetteite, 

egglestonite, and terlinguite, found at 

some sites are potentially much more 

bioavailable than highly insoluble 

forms of mercury, such as cinnabar 

and metacinnabar. Therefore, bioavailability may vary from site-to-site 

with mine waste contamination. 

However, Brown et al. [6] also 

found that sorption of Hg(II) onto 

mineral particles may effectively 

sequester mercury in mine tailing 

and aquatic environments if the sorption complexes are strongly bonded 

to the particle surfaces. An assessment of Hg(II) sorption complexes 

that form on mineral particles in 

mercury-mine waste environments 

showed that Hg(II) forms dominantly 

inner-sphere sorption complexes 

on these common mineral particles. 

An important result of this study is 

the finding that iron and aluminum 

(hydr)oxides can play a significant 

role in the uptake of Hg(II) through 

direct inner-sphere sorption processes. Additionally, the sorption of 

Hg(II) on these particles was examined in the presence of chloride and 

sulfate ions as a function of pH. The 

presence of chloride reduced Hg(II) 

sorption (through the formation of 

HgCl2 complexes), and the presence 

of sulfate enhanced Hg(II) sorption 

(through the reduction in electrostatic repulsion at positively-charged 

particle surfaces).

Bioaccumulation

Hurley et al. [1] and Fitzgerald et al. 

[7] studied the river/lake interfaces 

(i.e., the riverine mixing zones) 

of Lake Superior and Long Island 

Sound, respectively, and observed 

that the concentration of mercury 

bound to the particulate phase 

increased in the mixing zones relative 

to the river water, which can affect 

bioavailability and bioaccumulation. 

Hurley et al. [1] measured mercury 

and methylmercury concentrations at the riverine mixing zone in 

Lake Superior. The methylmercury 

concentration on particles increased 

within the mixing plume with distance from the interface; in other 

words, the methylmercury content 

of particles is enriched at the outer 

edge of the plume. Larger particles 

(and the methylmercury associated 

with these particles) settle out closer 

to the mouth of the river; smaller 

particles remain in solution and carry 

their load further offshore. Therefore, 

methylmercury in the particulate 

phase enters the lake water column 

from the river associated predominantly with small particles, which 

can be ingested by zooplankton. The 

result is enhanced bioaccumulation 

of methylmercury in plankton in the 

mixing zone. 

3.5 New Methods for 

Mercury Analysis

Several new analytic methods for 

measuring total and speciated mercury in the environment were developed under the STAR grants:

• A batch culture bioassay using Selenastrum capricorutum was developed to measure bioaccumulation 

of methylmercury by algae [1].

• Refined methods to measure 

methylation rates in sediments at 

near ambient mercury levels were 

developed using individual stable 

mercury isotopes and ICP-MS 

analysis [1, 8]. These methods 

also allow tracking of both the in 

situ mercury methylation and the 

fate of the added mercury. Using 

these techniques, methylation and 

demethylation were measured in 

the same sample concurrently [8].

• Synchrotron-based x-ray absorption fine structure (XAFS) spectroscopy was used to determine 

the molecular-level speciation 

of Hg in many different mining 

localities in California [6]. This 

method was compared to sequential extraction methods and solid 

phase mercury thermodesorption 

in a collaborative study [10].

• An in vitro reductible-mercury 

titration approach was developed 

to determine the concentration 

and conditional stability constants 

of dissolved organic matter towards mercury [7].

• A semi-automated dissolved elemental mercury analyzer (DEMA) 

was developed for measuring 

Hg(0) concentrations in laboratory samples of surface water [7].

• A method was refined to estimate 

the bioavailability of mercury at 

environmental levels through 

measurement of the octanol-water 

partitioning of mercury [8]. Previous measures were effective only 

for high concentrations of mercury. Clean techniques and other 

modifications allowed detection of 

mercury at picomolar levels.

• A new technique for measuring 

low levels of sulfide in natural 

waters was developed to detect at 

nanomolar to micromolar levels of 

mercury [8].

• EcoCELLs were used for investigating the biogeochemical cycling of an 

environmental contaminant [11].
18

In this section, the findings presented 

by research theme in Section 3 are 

summarized in three ways according 

to their potential application in future 

mercury research and modeling. First, 

results are grouped according to how 

they may be used in the development or improvement of mercury fate 

and transport models. Second, key 

variables studied here that affect the 

fate and transport of mercury in the 

environment are cross-walked with 

mercury cycling processes that are 

influenced by these variables. Finally, 

information about how these research 

results broaden our understanding 

of variation across ecosystems and 

regions is presented. 

4.1 Potential Improvements to 

Mercury Fate and Transport 

Models

This section summarizes new models developed as a part of the STAR 

research and groups key findings that 

will be useful in the refinement of 

existing mercury fate and transport 

models.

New Models

Work under two of the STAR grants 

included developing models that 

incorporated the observations made 

by the investigators in their ecosystem studies. Methods and algorithms 

from these models could potentially 

be used in the development of other 

models as well. Driscoll et al. [5] 

constructed the Mercury Cycling 

Model for Headwater Drainage Lake 

Systems (MCM-HD) to simulate 

mercury interactions in headwater 

drainage lakes and the adjacent 

watersheds, including wetlands, and 

found that results from the model 

were generally in agreement with 

mass balance calculations. Hurley 

et al. [1] developed a simple model 

that related annual tributary loading 

from a watershed to the lake based 

on watershed composition (i.e., land 

use/type). The model was applied to 

the entire lake watershed. Data from 

that study [1] also are being incorporated into the Lake Superior dynamic 

Mercury Cycling Model (D-MCM) to 

elucidate the effects of mercury loads 

in specific nearshore regions.

New Data 

STAR grant investigators gained new 

understanding of various processes 

that affect mercury fate and cycling in 

the environment that can be incorporated into existing or new mercury 

models. Findings are summarized 

according to their relevance to methylation, inorganic transformations, 

and mercury transport.

Factors Affecting Methylation

Several investigators studied mercury methylation and the factors 

that enhance or inhibit it. Fitzgerald 

et al. [7] demonstrated that methylation rates increase with elevated 

bacterial metabolic rates and higher 

concentrations of inorganic mercury. 

Factors affecting methylation studied 

by other researchers can be grouped 

into these same two categories.

Bacteria methylation rates influence 

the total amount of methylmercury 

that is formed. Three important factors that influence bacteria methylation rates were studied, including:

•	Time of year/temperature. Methylation rates are higher during the 

warmer summer months [7].

•	Sulfate levels. Increasing sulfate 

levels can increase methylation in 

sulfate-poor systems [2, 8]. 

•	 Availability of Hg(II). Increasing 

inorganic mercury loading to a watershed increases methylation rates [3].

The STAR grant investigators demonstrated the effect of several factors on 

the availability of inorganic mercury 

for methylation. Many of these factors 

would be important in mercury fate 

models.

•	Organic Matter. Increasing 

organic matter concentrations 

decreases the concentration of 

methylmercury, probably because inorganic mercury forms 

complexes with DOC, making it 

unavailable for uptake by methylating bacteria [1,7]. 

•	Sulfide Concentration. The addition of sulfide can decrease the 

rate of methylation [8]. The effect 

of sulfide on methylation depends 

on whether neutral or charged 

sulfide mercury complexes form. 

Neutral complexes enhance methylation, while charged complexes 

inhibit it [8].

4. Applications of Findings from Star Grants
19

•	Particulate Binding. Inorganic 

mercury binding to sediment and 

particulate matter will decrease 

methylation. Higher sedimentwater distribution coefficients for 

inorganic mercury are associated 

with lower methylation rates [7]. 

•	Mixing Zones. The fraction of 

reactive mercury increases in the 

mixing zones of rivers with lakes 

or sounds, thereby increasing the 

amount available for methylation 

and bioaccumulation [1,7].

•	Bioturbation. Increased bioturbation of benthic infauna in the sediment can increase available mercury 

and increase methylation [7].

•	 “Newness” of Mercury. In some 

environments, newly deposited 

mercury contributes more substantially to production of methylmercury than “older” mercury [3]. Sulfate and DOC have a much larger 

effect on the methylation of newly 

deposited mercury than of mercury 

already present in soils [3].

Factors Affecting Inorganic 

Mercury Transformations

Transformations between Hg(II) and 

Hg(0) also were studied by STAR 

grant investigators, and their results 

include key information that could be 

incorporated into mercury environmental fate models. Reduction of 

Hg(II), photooxidation of Hg(0), and 

volatilization of Hg(0) are all important processes in the mercury cycle 

that can ultimately affect the availability of mercury for methylation and 

subsequent accumulation in the food 

chain, as indicated below.

•	Photochemical Reduction.

Hg(II) is readily reduced to Hg(0) 

in aquatic surface waters by photochemical reduction; as a result, 

Hg(0) formation and subsequent 

volatilization display a diurnal 

cycle [4].

•	Organic Matter. Some data indicate that Hg(II) must be bound 

to organic material to be reduced 

by abiotic processes in natural 

waters. However, excess ligands 

can result in significantly lower 

reduction rates of Hg(II) [7]. 

•	Photooxidation. Photooxidation 

is responsible removing much 

of the Hg(0) in surface waters. 

In some cases, photooxidative 

flux can exceed volatilization as 

a loss mechanism for Hg(0) in an 

aquatic system [2, 9].

Factors Affecting Transport

Some data points collected on the 

major transport processes for mercury through a watershed could be 

incorporated into fate and transport 

models.

•	Volatilization. Volatilization is an 

important process for removing 

Hg(0) from surface waters and a 

potentially significant source of 

Hg(0) to the atmosphere [1, 2, 4, 

6, 7]. For example, Fitzgerald et 

al. [7] estimated annual emissions 

of Hg(0) from Long Island Sound 

to be 35 percent of the input of 

mercury to Long Island Sound. 

Volatilization of mercury from 

soils, which has been shown to 

vary based on light, could also be 

an important source of mercury to 

the atmosphere [10, 11].

•	Deposition and Plant Transfers.

Several studies confirmed that 

atmospheric deposition is the key 

pathway for emitted mercury to enter watersheds [1, 2, 5], and some 

provided quantitative data on this 

relationship (e.g., dry deposition 

was 70 percent of the total deposition of mercury to the forest floor 

[5]). Additionally, plant foliage may 

be a significant sink of mercury 

from the atmosphere, and consequentially may be a large source of 

mercury to watershed via litterfall 

and deposition [11].

•	Runoff/Erosion. Results were 

mixed regarding the extent to 

which uplands and wetlands function as a source of methylmercury 

to downstream water bodies. The 

rate of transport and methylation 

in a watershed depends on ecosystem-specific properties [1, 3, 5] 

as well as sorption onto mineral 

particles and availability of complexing ligands [6].

4.2 Understanding Key Variables 

Affecting Mercury Fate and 

Transport 

Several specific variables that affect 

mercury fate and transport in the 

environment were identified and 

studied in the STAR mercury research 

projects. Key variables are summarized in Table 4. The applicability of 

each variable to mercury transport, 

availability, or speciation (as determined by these studies) is indicated. 

4.3 Understanding Variation 

in Key Variables by Ecoregion 

and Ecotype

The STAR grant research on mercury 

was performed in a variety of ecosystems across North America, including an Adirondack lake, Long Island 

Sound, the Experimental Lakes Area 

in northwestern Ontario, Canada, the 

Florida Everglades, Lake Superior, 

the Patuxent River in Maryland, the 

St. Lawrence River, the Saginaw Bay 

watershed in Michigan, the Marcell 

Experimental Forest in northeastern 

Minnesota, and mining and other 

mercury source locations in California and Nevada. Some regional 

differences were noted by researchers; however, no study attempted 

to systematically identify regional 

characteristics affecting mercury fate 

and transport in the environment.

Regional differences are expected 

for processes that are affected by 

temperature and solar radiation. 

For example, Gilmour et al. [3] 

performed similar experiments 

in the Florida Everglades and the 

Experimental Lakes Area in Canada 

and observed differences in methylmercury response to mercury surface 
20

T

Table 4. Variables Affecting Mercury Fate and Transport

Variable

Processes Affected by Variable

Transport 

between media

Availability to 

methylating bacteria

Transformation 

between Hg speciesa

Deposition rate / mercury load 4 4 4

Sulfate load 4b

Sulfide concentrations and charge of 

mercury-sulfide complexes 4

Dissolved organic carbon (DOC) 4 4 4

Suspended particulate matter 4 4

Sediment burial 4 4

Temperature 4 4b

Bacterial activity 4b

“Newness” of mercury 4 4

Phase partitioning of mercury between 

solid and liquid in sediment and surface 

water

4 4 4

Bioturbation of sediments 4

pH 4

Photo-demethylation of methylmercury 4

Oxidation of Hg(0) 4

Reduction of Hg(II) 4 4

Volatilization of Hg(0) (phase 

partitioning between water and air) 4

Vegetation 4

a For reasons other than availability to methylating bacteria, which is covered in the previous column.

b Principally due to effects on bacteria methylation rates. 
21

application at these two locations, 

with shorter response times measured in the shallow, warm sites of 

the Everglades. Both photooxidation of Hg(0) and photoreduction of 

Hg(II) will be more prevalent in areas 

with less cloud cover and, during the 

summer, at lower latitudes with more 

intense solar radiation. The balance 

between those two processes is likely 

to be affected by more localized factors, however.

Ecosystem differences are expected 

for processes that are affected by 

DOC and particulate matter in surface waters, which can vary substantially depending on the type of water 

body and characteristics of the watershed. The effects of these factors on 

mercury bioavailability, methylation, 

and bioaccumulation are complex, 

and will need to be considered in 

conjunction with other factors. In 

freshwater systems, high levels of 

suspended sediment particles tend 

to result in lower dissolved Hg(II) 

concentrations as the mercury sorbs 

to the particles and may become 

buried in the benthic sediments. 

Where rivers meet large lakes or the 

ocean, on the other hand, the smaller 

suspended sediment particles might 

enhance transfer of mercury through 

aquatic food chains as the mercurycontaining particles are consumed 

by zooplankton in the mixing zone. 

DOC can affect mercury cycling in 

aquatic systems by binding dissolved 

inorganic and organic mercury in 

complexes limiting its bioavailability to methylating bacteria and to 

other aquatic organisms, by limiting 

UV light penetration and photolytic 

processes, and by generating reactive molecules that assist in Hg(II) 

reduction. The affinity of DOC for 

mercury varies across ecosystems 

as well. Given the importance of 

these localized factors, differences 

in mercury distribution can occur 

within relatively small geographic 

regions. For example, Driscoll et al. 

[5] compared measured BCF values 

in Sunday Lake in the Adirondacks to 

values measured in other Adirondack 

lakes and observed that BCF values 

were lower in Sunday Lake. He suggested that the methylmercury in 

Sunday Lake may be less bioavailable 

due to binding associated with high 

concentrations of DOC.
22

5. Future Research Needs

Several overarching future research 

needs related to mercury fate and 

transport were identified by comparing the STAR research results for these 

eight projects to the goals outlined 

in the ORD Research Plan and the 

original STAR Mercury RFA (see Section 2.2). Additionally, areas of further 

research that were specifically identified by investigators as follow-up to 

the work completed under these STAR 

grants are summarized below.

Topics Included in ORD 

Research Plan

The research summarized here 

contributes new data and analyses 

that improve general understanding 

of processes and factors that affect 

the fate and transport of mercury 

in the environment. Thus, these 

research projects have contributed 

significantly to the first two ORD 

goals/measures listed in Table 2. 

Specific ORD research goals and performance measures that are related 

to the grants described here (see 

Section 2.3), but may require further 

work to fully accomplish, include 

development of a model for mercury 

in fish, identification of sources of 

mercury emissions, and the eventual 

creation of an integrated multimedia 

modeling framework for mercury 

in the environment. The research 

covered by the eight STAR grants 

described here will help EPA accomplish these broader, long-term goals; 

however, it is expected that other 

ORD-related research will also play a 

role in ultimately meeting those goals 

(especially the long-term goal of an 

integrated multimedia framework).

Topics of STAR RFAs

The specific research questions 

posed in the original RFA are listed 

in Section 2.2. The majority of 

the STAR grant research projects 

described in this report focused on 

Question 2 (i.e., variables controlling 

mercury transformations). Specific 

variables were identified that control 

the transformation of mercury to 

methylmercury as well as transformations between inorganic species of 

mercury. 

Question 1 (i.e., bioavailability and 

fish concentrations of mercury) was 

addressed by some investigators 

(e.g., Hurley et al. [1] and Driscoll 

et al. [5]). Further research into the 

relationship between mercury in 

the watershed to mercury in biota, 

however, may be necessary to close 

information gaps regarding uptake by 

fish and other organisms. 

The STAR research described here 

was performed in a range of locations; however, follow-up research 

systematically analyzing the differences between results for these different ecosystems would be required 

to comprehensively answer Question 

3 (i.e., how mercury cycling varies 

in different geographic regions). In 

addition to further analyzing the 

existing results across regions, studies that conduct research in parallel 

in more than one region (similar 

to the Gilmour et al. [3] study in 

Canada and Florida) may be useful 

in clearly identifying how mercury 

cycling varies by region. The focus 

for most of the research presented 

here was on freshwater ecosystems; 

therefore, further investigation into 

the differences between freshwater 

and estuarine or coastal ecosystems 

would be beneficial, especially 

because research presented here 

(Fitzgerald [7]) indicates that processes may be different in the estuary 

environment than in freshwater.

Topics Specified by Researchers

Many of the STAR grant final reports 

list areas where further research 

is warranted. Investigators have 

reported both general subjects worthy of future consideration and more 

specific research topics that follow 

directly from work completed to 

date. Both types of recommendations 

are summarized here.

General areas for continued or future 

research identified by the STAR grant 

investigators include the following.

•	Effects of sulfate on methylation. Driscoll et al. [5] suggested 

developing a new formulation for 

the response of methylation to 

changes in sulfate concentrations 

in a wetland environment. Other 

research mentioned by Driscoll et 

al. has shown a greater-than-linear 

response in methylation when 

sulfate is added; however, the 

current model, based on data for 

lakes, does not predict that type of 

response.

•	Intracellular sequestration of

mercury. Mason et al. [8] noted 

that not all of the mercury entering a sulfate-reducing bacterium 
23

in neutral sulfide complexes is 

methylated. Methylating bacteria 

carry out both methylation processes and competing processes 

that sequester mercury and render 

it unavailable for methylation. It 

has been hypothesized that differences in cell physiology (e.g., size 

and membrane composition) and 

in mercury partitioning within the 

cell cause the variation in methylation rates observed among bacterial strains. Further work is needed 

to understand the mechanisms 

and kinetics of these intracellular 

sequestration pathways. 

•	Demethylation. Mason et al. [8] 

also observed that the factors influencing demethylation processes 

are not well understood and 

deserve further study.

•	 “Newness” of mercury. Mason et 

al. [8] further suggested that more 

work is needed to elucidate the 

differences in methylation rates of 

newly added mercury compared 

to in situ mercury. It is difficult 

to assess the pool of mercury 

available for methylation, and this 

information is also necessary to 

accurately estimate methylation 

rates.

•	Photooxidation. Morel et al. [9] 

recommended further research 

in different types of water bodies 

to determine if more of the Hg(0) 

in surface waters is converted to 

Hg(II) via photooxidation or is 

lost to the atmosphere via volatilization of Hg(0).

•	SRB studies. Morel et al. [9] 

suggested additional research 

on bacterial methylation by SRB, 

including: the role of vitamin B12

and other methyl-transferring 

coenzymes in methylmercury production by SRB; the role internal 

mercury speciation plays in methylation; the pathways responsible 

for methylation; and the biomechanical mechanism of bacterial 

methylation.

•	Mercury in seafood. Morel et al. 

[9] and Nriagu et al. [4] both noted that more research is needed 

to determine the ultimate sources 

of mercury in seafood.

Specific research needs noted by 

investigators that relate to sites or 

topics studied in STAR grants include 

the following.

• Gilmour et al. [3] suggested 

follow-up research in the Experimental Lakes Area in Canada to 

determine time to equilibrium 

for increased mercury loading at 

Experimental Lakes Area; to determine the time course for mercury 

movement through uplands and 

wetlands; to investigate bioavailability of mercury from uplands 

and wetlands for methylation 

and bioaccumulation; to investigate biogeochemical parameters 

that affect bioavailability; and to 

more fully investigate “aging” of 

dosed mercury (including ligand 

exchange of mercury).

• Hurley et al. [1] listed several 

specific next steps for research in 

Lake Superior, including the investigation of atmospheric deposition 

(i.e., effects of local urban sources 

on nearshore regions and direct 

deposition of methylmercury in 

the basin), mixing zones/watershed influences (i.e., mechanisms 

for algal and zooplankton uptake 

and fate of colloidal methylmercury), and in-lake processes (i.e., 

fate of changing anoxia on methylmercury dynamics in Lake Erie, 

mechanisms of bioaccumulation 

in offshore zones, and methylation in deep sediments).

• Nriagu et al. [4] observed associations between photosynthetically 

active radiation and indicators of 

biological activity (e.g., chlorophyll-a). They plan to use those 

data to further explore the interaction of biological processes with 

the mercury cycle in the Saginaw 

Bay. 

• Brown et al [6] saw in laboratory 

experiments that organic acids 

may affect mercury transport from 

mining sites, and suggested further field studies into the potential impact of revegetation on the 

mobilization of colloidal materials 

from mine tailings. Additionally, Brown et al. [6] suggested 

further studies into the question 

of whether ratios of mercury isotopes could be used to determine 

the source of mercury. Specifically, 

measurements of the isotope 

ratios need to be made with more 

precision (using more sensitive 

multi-collector, inductively coupled plasma mass spectrometer).

• Gustin et al. [11] follow-up 

research has already begun and 

includes using the EcoCELLs to 

study mercury cycling associated 

with tall grass prairie vegetation 

and soils from Oklahoma.

In addition, these STAR grant results 

have provided a wealth of data that 

merits further analysis and comparative review. For example, the research 

of Nriagu et al. [4] could be analyzed 

for additional patterns and reasons 

for reduction of mercury and other 

transformations. Measured mercury 

concentrations in environmental 

media at the different locations studied by the STAR grant investigators 

could be tabulated, and ratios of total 

mercury, Hg(0), Hg(II), and methylmercury, could be compared to 

identify regional and temporal trends 

or patterns of mercury distribution in 

the environment.
24

6. References Cited

Environmental Protection Agency 

(EPA). (1997a). Mercury Study 

Report to Congress. EPA-452/R97-003 to -010. U.S. EPA, Office 

of Air Quality Planning and Standards, and Office of Research 

and Development, Washington, 

DC.

EPA. (1997b). Great Waters Second 

Report to Congress. EPA-453/R97-011. U.S. EPA, Office of Air 

Quality Planning and Standards, 

Research Triangle Park, NC.

EPA. (1998). Utility Air Toxics Report 

to Congress. EPA-453/R-98-004. 

U.S. EPA, Office of Air Quality Planning and Standards, 

Research Triangle Park, NC.

EPA. (2001). Office of Research and 

Development Strategic Plan.

EPA/600/R-01/003. U.S. EPA, 

Office of Research and Development, Washington, DC. January.

EPA. (2003a). 2003-2008 EPA Strategic Plan: Direction for the 

Future. EPA-190-R-03-003. U.S. 

EPA, Office of the Chief Financial 

Officer, Office of Planning, Analysis, and Accountability, Washington, DC. September.

EPA. (2003b). Mercury Research 

Multi-Year Plan. FY 2005 Planning - Final Version, May 9, 2003. 

U.S. EPA, Office of Research and 

Development, Washington, DC. 

Available at: http://www.epa.

gov/osp/myp/mercury.pdf.

EPA. (2003c). Air Toxics MultiYear Plan. U.S. EPA, Office of 

Research and Development, 

Washington, DC. April 2003 

Update.

Mason RP, Abbott ML, Bodaly RA, 

Bullock Jr. OR, Driscoll CT, Evers 

D, Lindberg SE, Murray M, Swain 

EB. (2005). Monitoring the 

response to changing mercury 

deposition. Environmental 

Science and Technology A-Pages

39(1):14A-22A.
A–1

Appendix: Grant Publication Lists

[1] Hurley, EPA Grant Number: 

R827629

In Print

Babiarz CL, Hurley JP, Hoffmann SR, 

Andren AW, Shafer MM, Armstrong 

DE. (2001). Partitioning of total 

mercury and methylmercury to the 

colloidal phase in fresh waters. Environmental Science and Technology, 

35(24):4773-4782.

Babiarz CL, Hoffmann SR, Shafer 

MM, Hurley JP, Andren AW, Armstrong 

DE. (2000). A critical evaluation of 

tangential-flow ultrafiltration for trace 

metal investigations in fresh water 

systems: Part II Total and Methylmercury. Environmental Science and 

Technology, 34(16):3428-3434.

Back RC, Hurley JP, Rolfhus KR. 

(2002). Watershed influences on the 

transport, fate and bioavailability of 

mercury in Lake Superior: Field measurements and modeling approaches. 

Lakes and Reservoirs: Research and 

Management, 7:201-206.

Back RC, Gorski PR, Cleckner LB, 

Hurley JP. (2003). Mercury content 

and speciation in the plankton and 

benthos of Lake Superior. Science of 

the Total Environment, 304:327-348. 

Cleckner LB, Back RC, Gorski PR, 

Hurley JP, Byler S. (2003). Seasonal 

and size-specific distribution of methylmercury in seston and zooplankton of two contrasting Great Lakes 

embayments. Journal of Great Lakes 

Research, 29(1):134-144.

Hoffmann SR, Shafer MM, Babiarz 

CL, Armstrong DE. (2000). A critical 

evaluation of tangental-flow ultrafiltration for trace metal investigations 

in fresh water systems: Part I Organic 

Carbon. Environmental Science and 

Technology, 34(16):3420-3427. 

Hurley JP, Cleckner LB, Shafer MM. 

(2001). Watershed influences on 

mercury transport to Lake Superior. 

Verhundlungen Internationale Vereinigung Limnologie, 27:3425-3428.

Hurley JP, Manolopoulos H, Babiarz 

CL, Sakamoto H, Rolfhus KR, Back 

RC, Shafer MM, Armstrong DE, Harris 

RC. (2003). Methyl mercury in Lake 

Superior: Offshore processes and bioaccumulation. Journal de Physique 

IV, 107:641-644.

Manolopoulos H, Hurley JP, Rolfhus 

KR, Sakamoto H, Back RC. (2002). 

Nearshore accumulation of methyl 

mercury in Lake Superior. Abstract 

Papers of the American Chemical 

Society, 223:101 ENVR Part1, April 7.

Manolopoulos H, Hurley JP, Babiarz 

CL, Back RC, Rolfhus KR. (2003). 

Riverine mixing zones as regions of 

enhanced methylmercury bioaccumulation in Lake Superior. Journal 

de Physique IV, 107:83-86.

Rolfhus KR, Sakamoto HE, Cleckner 

LB, Stoor RW, Babiarz CL, Back RC, 

Manolopoulos H, Hurley JP. (2003). 

The distribution of mercury in Lake 

Superior. Environmental Science and 

Technology, 37(5):865-872.

Stoor RW. (2002). Groundwater 

contributions of methyl mercury to a 

Lake Superior watershed. M.S. Thesis, University of Wisconsin-Madison, 

Madison, WI.

In Preparation

Babiarz CL, Hurley JP, Shafer MM, 

Wieben A, Robertson D, Hoffmann 

SR, Andren AW, Armstrong DE. Watershed and hydrologic influences on 

colloidal phase loading of mercury 

from U.S. tributaries into Lake Superior. To be submitted to the Journal 

of Great Lakes Research.

Gorski P. Factors influencing the 

bioaccumulation of mercury in aquatic 

systems. Ph.D. Dissertation, University 

of Wisconsin-Madison, Madison, WI.

Manolopoulos H, Babiarz CL, Hurley 

JP, Back RC. Mechanisms of methyl 

mercury assimilation and bioaccumulation in Lake Superior riverine mixing 

zones. To be submitted to Environmental Science and Technology.

[2] Swain, EPA Grant Number: 

R827630

In Print

Hines NA. (2004). Mercury and 

methylmercury in Spring Lake, Minnesota: a mass balance approach 

comparing redox transformations, 

sediment loading, lake dynamics, and 

watershed processes. Ph.D. Thesis, 

University of Minnesota, Minneapolis, 

Minnesota. 
A–2

Hines NA, Brezonik PL. (2004). 

Mercury dynamics in a small northern Minnesota lake: Water to air 

exchange and photoreactions of 

mercury. Marine Chemistry, 90(1-

4):137-139.

Johnson BM. (2004). Sulfate reducing 

bacteria and the role of nutrients in 

mercury methylation in sediments of 

Spring Lake, Minnesota. M.S. Thesis, 

University of Minnesota, Minneapolis, 

Minnesota.

In Preparation

Hines NA, Brezonik PA, Engstrom DR. 

Sediment and porewater profiles and 

fluxes of mercury and methylmercury 

in a small seepage lake in northern 

Minnesota. (Submitted).

[3] Gilmour, EPA Grant Number: 

R827631

In Print

Babiarz CL, Hurley JP, Krabbenhoft 

DP, Gilmour CC, Branfireun B. 

(2003). Application of ultrafiltration 

and stable isotopic amendments to 

field studies of mercury partitioning 

to filterable carbon in lake water and 

overland runoff. Science of the Total 

Environment, 304:295-303.

Benoit JM, Gilmour CC, Mason RP. 

(2001). Aspects of the bioavailability 

of mercury for methylation in pure 

cultures of Desulfobulbous propionicus (1pr3). Applied and Environmental Microbiology, 67:51-58.

Benoit JM, Gilmour CC, Mason RP. 

(2001). The influence of sulfide on 

solid-phase mercury bioavailability 

for methylation by pure cultures of 

Desulfobulbous propionicus (1pr3). 

Environmental Science and Technology, 35:127-132. 

Benoit JM, Mason RP, Gilmour CC, 

Aiken GR. (2001). Mercury binding constants for dissolved organic 

carbon isolates from the Florida Everglades. Geochimica et Cosmochimica 

Acta, 65:4445-4451. 

Benoit J, Gilmour C, Heyes A, Mason 

RP, Miller C. (2003). Geochemical and 

biological controls over methylmercury production and degradation in 

aquatic ecosystems. In: Biogeochemistry of Environmentally Important 

Trace Elements, ACS Symposium 

Series #835, Y. Chai and O.C. Braids, 

Eds. American Chemical Society, 

Washington, DC. pp. 262-297. 

Jay JA, Murray KJ, Gilmour CC, Mason 

RP, Morel FMM, Roberts AL, Hemond 

HF. (2002). Mercury methylation by 

Desulfovibrio desulfuricans ND132 

in the presence of polysulfides. 

Applied and Environmental Microbiology, 68:5741-5745.

[4] Nriagu, EPA Grant Number: 

R82763201

In Print

Zhang H, Lindberg S. (2001). Sunlight and iron(III)-induced photochemical production of dissolved 

gaseous mercury in freshwater. 

Environmental Science and Technology, 35:928-935.

Zhang H, Lindberg S, Gustin M, Xu 

X. Towards a better understanding of 

mercury emissions from soils. In: Cai 

Y, Braids O [eds.] Biogeochemistry of 

Environmentally Important Elements.

American Chemical Society Symposium Series Book, American Chemical 

Society, Oxford University Press. 

[5] Driscoll, EPA Grant Number: 

R827633

In Preparation

Driscoll CT, Munson RK, Yavitt J, 

Newton RM, Demers J, Kalicin M, 

McLaughlin E. Chemical and biological control of mercury cycling in 

upland, wetland and lake ecosystems 

in the Adirondack Region of New 

York, USA. Environmental Pollution.

Kalicin M, Driscoll C, Yavitt J, Newton 

R, Munson R. The dynamics of mercury in upland forests of the Adirondack region of New York. Environmental Pollution.

McLaughlin E, Driscoll C, Yavitt J, 

Newton R, Munson R. Mercury in 

upland and riparian wetland vegetation. Environmental Pollution.

McLaughlin E, Driscoll C, Sutherland J, Yavitt J, Newton R, Munson 

R. Trophic transfer of mercury in an 

Adirondack lake ecosystem. Environmental Pollution.

Munson RK, Harris RC, Driscoll CT, 

Yavitt J, Newton RM. The mercury 

cycling model for headwater drainage 

lakes (MCM-HD): Model theory and 

processes. Environmental Pollution.

Perry ER, Norton SA, Kamman KC, 

Lorey PM, Haines T, Driscoll CT. Mercury accumulation in lake sediments 

in the northeastern United States 

during the last 150 years. Ecotoxicology (In press).

Yavitt JB, Demers J, Driscoll CT, Kalicin M, Newton R, Munson R. Wetland 

characteristics and mercury behavior 

within an Adirondack (New York 

State) watershed. Environmental 

Pollution.

[6] Brown, EPA Grant Number: 

R827634

In Print

Coolbaugh MF, Gustin MS, Rytuba JJ. 

(2002). Annual emissions of mercury to the atmosphere from three 

natural source areas in Nevada and 

California, Environmental Geology,

42:338-349.

Engle MA, Gustin MS, Zhang H. 

(2001). Quantifying natural source 

mercury emissions from the Ivanhoe 

Mining District, north-central Nevada, 

USA. Atmospheric Environment, 

35:3987-3997.

Engle MA, Gustin MS. (2002). Scaling 

of atmospheric mercury emissions 

from three naturally enriched areas: 

Flowery Peak, Nevada, Peavine Peak, 

Nevada and Long Valley Caldera, California. Science of the Total Environment, 290(1-3):91-104.
A–3

Giglini T. (2003). Measurement of 

Total and Reactive Mercury above a 

Naturally Enriched, an Anthropogenically Contaminated and a Pristine 

Site, Nevada. M.S. Thesis, University 

of Nevada, Reno.

Gustin MS. (2003). Are mercury 

emissions from geologic sources 

significant? A status report submitted 

to Science of the Total Environment,

304:153-167.

Gustin MS, Lindberg SE, Austin K, 

Coolbaugh M, Vette A, Zhang H. 

(2000). Assessing the contribution of 

natural sources to regional atmospheric mercury budgets. Science of 

the Total Environment 259(1):61-72.v

Gustin MS, Biester H, Kim CS. (2002). 

Investigation of light enhanced 

emissions of mercury from naturally 

enriched substrate. Atmospheric Environment, 36(20):3241-3254.

Gustin MS, Coolbaugh M, Engle M, 

Fitzgerald B, Keislar R, Lindberg 

S, Nacht D, Quashnick J, Rytuba 

J, Sladek C, Zhang H, Zehner RE. 

(2003). Atmospheric mercury 

emissions from mine wastes and 

surrounding geologically enriched 

terrains. Environmental Geology,

43:339-351.

Johnson SB, Yoon TH, Slowey AJ, 

Brown GE Jr. (2004). Adsorption 

of organic matter at mineral/water 

interfaces: 3. Implications of surface 

dissolution for adsorption of oxalate. 

Langmuir, 20(26):11480-11492.

Kim CS. (2002). Mercury Speciation 

and Sorption Processes in Mining 

Environments. Ph.D. Thesis, Department of Geological & Environmental 

Sciences, Stanford University, Stanford, CA.

Kim CS, Rytuba JJ, Brown GE Jr. 

(1999). Utility of EXAFS in characterization and speciation of mercurybearing mine wastes. Journal of 

Synchrotron Radiation, 6:648-650.

Kim CS, Brown GE Jr., Rytuba JJ. 

(2000). Characterization and speciation of mercury-bearing mine wastes 

using X-ray absorption spectroscopy 

(XAS). Science of the Total Environment, 261:157-168.

Kim CS, Catalano JG, Grolimund D, 

Warner JA, Morin G, Juillot F, Calas 

GC, Ildefonse P, Rytuba JJ, Parks GA, 

Brown GE Jr. (2000). EXAFS determination of the chemical speciation and 

sorption processes of Hg(II), Sr(II), 

and Zn(II) in natural and model systems. 1999 Activity Report, Stanford 

Synchrotron Radiation Laboratory 

Report, Stanford, CA.

Kim CS, Bloom NS, Rytuba JJ, Brown 

GE Jr. (2003). Mercury speciation 

by extended x-ray absorption fine 

structure (EXAFS) spectroscopy and 

sequential chemical extractions: A 

comparison of speciation methods. 

Environmental Science and Technology, 37:5102-5108.

Kim CS, Rytuba JJ, Brown GE Jr. 

(2004). Geological and anthropogenic factors influencing mercury 

speciation in mine wastes. Applied 

Geochemistry, 19:379-393.

Kim CS, Rytuba JJ, Brown GE Jr. 

(2004). Mercury(II) sorption to 

Fe- and Al-(hydr)oxides: II. Effects of 

chloride and sulfate. Journal of Colloid and Interface Science, 270:9-20.

Kim CS, Rytuba JJ, Brown GE Jr. 

(2004). Mercury(II) sorption to 

Fe- and Al-(hydr)oxides: I. Effects of 

pH. Journal of Colloid and Interface 

Science, 271:1-15.

Lowry GV, Shaw S, Kim CS, Rytuba 

JJ, Brown GE Jr. (2004). Particlefacilitated mercury transport from 

New Idria and Sulphur Bank mercury 

mine tailings. 1. Column experiments 

and macroscopic analysis. Environmental Science and Technology, 

38(19):5101-5111.

Nacht DM. (2002). Measurement 

of Reactive Gaseous Mercury and 

Mercury Flux from Substrates in 

California and Nevada. M.S. Thesis, 

University of Nevada, Reno.

Nacht DM, Gustin MS. (2004). Mercury emissions from background and 

altered geologic units throughout 

Nevada. Water, Air, and Soil Pollution, 151:179-193.

Sladek C, Gustin MS, Kim CS, Biester 

H. (2002). Application of three methods for determining mercury speciation in mine waste. Geochemistry: 

Exploration, Environment, Analysis, 

2:369-376.

Slowey AJ, Rytuba JJ, Brown GE Jr. 

(2005). Speciation of mercury and 

mode of transport from placer gold 

mine tailings. Environmental Science 

and Technology, 39(6):1547-1554.

Zehner RE, Gustin MS. (2002). Estimation of mercury vapor flux from 

natural substrate in Nevada. Environmental Science and Technology, 

36(19):4039-4045.

Zhang H, Lindberg SE, Gustin MS, Xu 

X. (2003). Toward a better understanding of mercury emissions from 

soils. In: Biogeochemistry of Environmentally Important Trace Elements, 

ACS Symposium Series #835, Y. 

Chai and O.C. Braids, Eds. American 

Chemical Society, Washington, DC. 

pp. 246-261. 

In Preparation

Kim CS, Gustin MS, Rytuba JJ, Brown 

GE Jr. Associations between Hg 

speciation and Hg vapor flux rates in 

mine wastes. Environmental Science 

and Technology.

Nacht DM, Gustin MS, Engle MA, 

Zehner RE, Giglini AD. Quantifying 

total and reactive gaseous mercury 

at the Sulphur Bank Mercury Mine 

Superfund Site, Northern California. 

Environmental Science and Technology. (Submitted).
A–4

Slowey AJ, Johnson SB, Rytuba JJ, 

Brown GE Jr. Role of Organic Acids 

in Promoting Colloid Transport of 

Mercury from Mine Tailings. Environmental Science and Technology.

(Submitted).

[7] Fitzgerald, EPA Grant Number: 

R827635

In Print

Fitzgerald WF, Vandal GM, Rolfhus 

KR, Lamborg CH, Langer CS. (2000). 

Mercury emissions and cycling in the 

coastal zone. Journal of Environmental Science, 12(1):92-101. 

Fitzgerald WF, Lamborg CH. (2003). 

Geochemistry of mercury in the 

environment. In: Sherwood-Lollar B, 

ed. Treatise on Geochemistry, Vol. 9: 

Environmental Geochemistry. Elsevier: St. Louis, MO.

Hammerschmidt CR, Fitzgerald WF. 

(2001). Formation of artifact methyl 

mercury during extraction from a 

sediment reference material. Analytical Chemistry, 73(24):5930-5936. 

Hammerschmidt CR, Fitzgerald WF. 

(2004). Geochemical controls of 

the production and distribution of 

methylmercury in near-shore marine 

sediments. Environmental Science 

and Technology, 38(5):1480-1486.

Lamborg CH, Tseng CM, Fitzgerald 

WF, Balcom PH, Hammerschmidt CR. 

(2003). Determination of mercury 

complexation characteristics of 

dissolved organic matter in natural 

waters through “reducible Hg” titrations. Environmental Science and 

Technology, 37(15):3315-3322.

Langer CS, Fitzgerald WF, Visscher 

PT, Vandal GM. (2001). Biogeochemical cycling of methylmercury at Barn 

Island Salt Marsh, Stonington, CT, 

USA. Wetlands Ecology and Management, 9(4):295-310.

Rolfhus KR, Lamborg CH, Fitzgerald 

WF, Balcom PH. (2003). Evidence 

for enhanced mercury reactivity in 

response to estuarine mixing. Journal of Geophysical Research-Oceans, 

108(C11):3353.

Rolfhus KR, Fitzgerald WF. (2001). 

The evasion and spatial/temporal distribution of mercury species in Long 

Island Sound, CT-NY. Geochimica et 

Cosmochimica Acta, 65(3):407-417.

Tseng CM, Balcom PH, Lamborg CH, 

Fitzgerald WF. (2003). Dissolved 

elemental mercury investigations in 

Long Island Sound using on-line Au 

amalgamation-flow injection analysis. 

Environmental Science and Technology, 37(6):1183-1188.

Vandal GM, Fitzgerald WF, Rolfhus 

KR, Lamborg CH, Langer CS, Balcom 

PH. (2002). Sources and cycling of 

mercury and methylmercury in Long 

Island Sound (Project No. CWF-326-

R). Final Report to the Connecticut 

Department of Environmental Protection, Long Island Sound Program.

In Preparation

Balcom PH, Fitzgerald WF, Vandal 

GM, Lamborg CH, Rolfhus KR, Langer 

CS, Hammerschmidt CH. Mercury 

sources and cycling in the Connecticut River and Long Island Sound. 

Marine Chemistry (Submitted).

Hammerschmidt CR, Fitzgerald WF, 

Lamborg CH, Balcom PH, Visscher PT. 

Biogeochemistry of methylmercury 

in sediments of Long Island Sound. 

Marine Chemistry (Submitted).

Lamborg CH, Fitzgerald WF, Skoog 

A, Visscher PT. The abundance and 

source of mercury-binding organic 

ligands in Long Island Sound. Marine 

Chemistry (Submitted).

[8] Mason, EPA Grant Number: 

R827653-01-0

In Print

Benoit JM, Mason RP, Gilmour CC. 

(2001). Aspects of the bioavailability 

of mercury for methylation in pure 

cultures of Desulfobulbous proprionicus (1PR3). Applied and Environmental Microbiology, 67:51-58. 

Benoit JM, Gilmour CC, Mason RP. 

(2001). The influence of sulfide on 

solid-phase mercury bioavailability 

for methylation by pure cultures of 

Desulfobulbous proprionicus (1PR3). 

Environmental Science and Technology, 35:127-132. 

Benoit JM, Mason RP, Gilmour CC, 

Aiken GR. (2001). Constants for 

mercury binding by dissolved organic 

matter isolates from the Florida Everglades. Geochimica et Cosmochimica 

Acta, 65:4445-4451.

Benoit JM, Gilmour CC, Heyes A, 

Mason RP, Miller CL. (2003). Geochemical and biological controls over 

mercury production and degradation 

in aquatic systems. pp. 262-297. In: 

Y Cai and OC Brouds [eds.], Biogeochemistry of Environmentally Important Trace Elements, ACS Symposium 

Series 835, ACS, Washington, DC. 

Heyes A, Miller C, Mason RP. (2004). 

Mercury and methylmercury in the 

Hudson River sediment: impact of 

resuspension on partitioning and 

methylation. Marine Chemistry, 

90:75-89.

Mason RP, Benoit JM. (2003). Organomercury compounds in the environment. pp. 57-99. In: P Craig [ed.], 

Organometallics in the Environment, John Wiley & Sons, New York.

[9] Morel, EPA Grant Number: 

R827615

In Print

Amyot M, AuClair JC, Poissant L. 

(2001). In situ high temporal resolution analysis of elemental mercury 

in natural water. Analytica Chimica 

Acta, 447:153-159. 

Ekstrom EB, Morel FMM, Benoit JM. 

(2003). Mercury Methylation Independent of the Acetyl-Coenzyme A 

Pathway in Sulfate-Reducing Bacteria. 

Applied and Environmental Microbiology, 69(9):5414-5422.
A–5

Kraepiel AL, Keller K, Chin HB, 

Malcolm EG, Morel FMM. (2003). 

Sources and Variations of Mercury in 

Tuna. Environmental Science and 

Technology, 37:5551-5558.

LaLonde JD, Amyot M, Kraepiel AML, 

and Morel FMM. (2001). Photooxidation of Hg(0) in Artificial and Natural 

Waters. Environmental Science and 

Technology, 35:1367-1372. 

LaLonde JD, Amyot M, Orvoine J, 

Morel FMM, AuClair JC, Ariya PA. 

(2004). Photoinduced Oxidation of 

Hg0(aq) in the Waters from the St. 

Lawrence Estuary. Environmental 

Science and Technology, 38:508-514.

[10] Gustin, EPA Grant Number: 

R825249

In Print

Engle ME. (2003). The mobility of 

mercury in epithermal mercury 

deposits in an arid environment. 

Masters Thesis, University of Nevada, 

Reno, 155 p.

Engle MA, Gustin MS, Zhang H. 

(2001). Quantifying natural source 

mercury emissions from the Ivanhoe 

Mining District, north-central Nevada, 

USA. Atmospheric Environment, 

35:3987-3997.

Engle MA, Gustin MS. (2002). Scaling 

up atmospheric mercury emissions 

from three naturally enriched areas: 

Flowery Peak, Nevada, Peavine Peak, 

Nevada and Long Valley Caldera, California. Science of the Total Environment, 290(1-3):91-104.

Gustin MS. (2003). Are mercury emissions from geologic sources significant?: A status report. Science of the 

Total Environment, 304:153-167.

Gustin MS, Rasmussen P, Edwards 

G, Schroeder W, Kemp J. (1999). 

Application of a laboratory gas 

exchange chamber for assessment of 

in situ mercury emissions. Journal of 

Geophysical Research - Atmospheres,

104(D17):21, 873-78.

Gustin MS, Lindberg SE, Austin K, 

Coolbaugh M, Vette A, Zhang H. 

(2000). Assessing the contribution of 

natural sources to regional atmospheric mercury budgets. The Science 

of the Total Environment, 259:61-72. 

Gustin MS, Biester H, Kim C. (2002). 

Investigation of light enhanced 

emission of mercury from naturally 

enriched substrate. Atmospheric 

Environment, 36:3241-3254. 

Gustin MS, Coolbaugh M, Engle M, 

Fitzgerald B, Keislar R, Lindberg 

S, Nacht D, Quashnick J, Rytuba J, 

Sladek C, Zhang H, Zehner R. (2003). 

Atmospheric mercury emissions from 

mine wastes and surrounding geologically enriched terranes. Environmental Geology, 43:339-351. 

Lindberg SE, Zhang H, Vette AF, 

Gustin MS, Barnette MO, Kuiken 

T. (2002). Dynamic flux chamber 

measurement of gaseous mercury 

emission fluxes over soils: Effect of 

flushing flow rate and verification of 

a two-resistance exchange interface 

simulation model. Atmospheric Environment, 36:847-859.

Sladek C. (2001). Investigation of 

methods for determining mercury 

speciation and mobility in substrate. 

Masters Thesis, University of Nevada, 

Reno.

Sladek C, Gustin MS, Biester H, Kim 

C. (2002). Application of three methods for determining mercury speciation in mine waste. Geochemistry: 

Exploration, Environment, Analysis, 

2(4):369-375. 

Sladek C, Gustin MS. (2003). Evaluation of sequential extraction methods 

for determination of mercury speciation and mobility in mine waste. 

Applied Geochemistry, 18(4):567-576.

Zehner RE, Gustin MS. (2002). Estimation of mercury vapor flux from 

natural substrate in Nevada. Environmental Science and Technology 

36:4039-4045.

Zhang H, Lindberg SE, Barnette MO, 

Vette AF, Gustin MS. (2002). Simulation of gaseous mercury emissions 

from soils measured with a dynamic 

flux chambers using a two-resistance 

model. Atmospheric Environment, 

36:835-846.

Zhang H, Lindberg SE, Gustin MS, Xu 

X. (2002). Towards a better understanding of mercury emissions from 

soils, in Biogeochemistry of Environmentally Important Trace Elements. 

Cai Y and Braids OC, eds. American 

Chemical Society, Washington DC, 

246-261.

[11] Gustin, EPA Grant Number: 

R827622E02

In Print

Benesch, JA. (2002). Assessing the 

role of deciduous forests in the 

biogeochemical cycling of mercury. 

Masters Thesis, University of Nevada, 

Reno, 90p.

Ericksen JA, Gustin MS, Schorran DE, Johnson DW, Lindberg SE, 

Coleman JS. (2003). Accumulation 

of atmospheric mercury in forest 

foliage. Atmospheric Environment, 

37(12):1613-1622. 

Ericksen JA, Gustin MS. (2004). Foliar 

exchange of mercury as function 

of soil and air concentration. The 

Science of the Total Environment,

324:271-279.

Frescholtz TF. (2002). Assessing the 

role of vegetation as sources and 

sinks of atmospheric mercury using 

Quaking Aspen. Masters Thesis, University of Nevada, Reno, 67p.

Frescholtz TF, Gustin MS, Schorran 

DE, Fernandez GC. (2003). Assessing the source of mercury in foliar 

tissue of quaking aspen. Environmental Toxicology and Chemistry, 

22(9):2114-2119.

Frescholtz TF Gustin MS. (2004). 

Soil and foliar mercury emission as a 

function of soil concentration. Water, 

Air, Soil Pollution,155:223-237.
A–6

Gustin MS, Ericksen JA, Schorran DE, 

Johnson DW, Lindberg SE, Coleman 

JS. (2004). Application of controlled 

mesocosm for understanding mercury plant-soil-air exchange. Environmental Science and Technology, 38: 

6044-6050.

Gustin MS, Stamenkovic J. (2005). 

Effect of watering and soil moisture 

on mercury emissions from soils. 

Biogeochemistry, 76(2):215-232.

Johnson D, Benesch JA, Gustin MS, 

Schorran DS, Lindberg SE, Coleman 

JS. (2003). Experimental evidence 

against diffusion control of Hg evasion from soils. Science of the Total 

Environment, 304(1-3):175-184. 

Recycled/Recyclable

Printed with Vegetable Oil Based Inks on Recycled Paper

(Minimum 50% Postconsumer) Process Chlorine Free

Office of Research and Development (8101R)

Washington, DC 20460

EPA/600/S-06/013

January 2006

www.epa.gov