Prepared in cooperation with the National Park Service

Assessment of the Use of Sorbent Amendments 

for Reduction of Mercury Methylation in Wetland 

Sediment at Acadia National Park, Maine

Scientific Investigations Report 2014–5234

U.S. Department of the Interior

U.S. Geological Survey
Front cover. Photograph of Hodgdon Pond at Acadia National Park, on August 12, 2013, from a point near the outlet of the pond looking north.

Extensive wetlands occur on east and west sides of the outlet stream. Photograph courtesy of Ariel Lewis, University of Maine, Orono, Maine.
Assessment of the Use of Sorbent 

Amendments for Reduction of Mercury 

Methylation in Wetland Sediment at 

Acadia National Park, Maine

By Thomas G. Huntington, Ariel Lewis, Aria Amirbahman, 

Mark Marvin-DiPasquale, and Charles W. Culbertson

Prepared in cooperation with the National Park Service

Scientific Investigations Report 2014–5234

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior

SALLY JEWELL, Secretary

U.S. Geological Survey

Suzette M. Kimball, Acting Director

U.S. Geological Survey, Reston, Virginia: 2015

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Suggested citation:

Huntington, T.G., Lewis, Ariel, Amirbahman, Aria, Marvin-DiPasquale, Mark, and Culbertson, C.W., 2015, Assessment 

of the use of sorbent amendments for reduction of mercury methylation in wetland sediment at Acadia National 

Park, Maine: U.S. Geological Survey Scientific Investigations Report 2014–5234, 30 p., http://dx.doi.org/10.3133/

sir20145234.

ISSN 2328-0328 (online)
iii

Acknowledgments

The authors are grateful for the support they received from several people that contributed to 

the successful completion of this study. David Manski and Bill Gawley of Acadia National Park, 

National Park Service helped facilitate this report from the initial conceptualization of the study 

through completion. Clive Devoy at the University of Maine in Orono provided assistance with 

the chemical analysis of pore water and sediment at the Sawyer Environmental Laboratory.

Martha Nielsen of the U.S. Geological Survey provided the map used in figure 1. Marie-Noële 

Croteau of the U.S. Geological Survey provided adult snails for rearing juveniles and provided 

assistance with the methods used to rear and process snails for analysis in experiments involving metal uptake. Evangelos Kakouros, Michelle Arias, and Le Kieu of the U.S. Geological Survey 

processed and analyzed sediment iron and mercury species and snail tissue for mercury species.

v

Contents

Acknowledgments........................................................................................................................................iii

Abstract...........................................................................................................................................................1

Introduction and Background......................................................................................................................1

The Mercury Contamination Problem ...............................................................................................2

Potential Remediation..........................................................................................................................3

Purpose and Scope ..............................................................................................................................4

Site Description.....................................................................................................................................4

Methods...........................................................................................................................................................6

Experiment 1 (Littoral Zone Sediment) ..............................................................................................6

Experiment 2 (Profundal Zone Sediment) .........................................................................................7

Experiment 3 (Vegetated Microcosms With Great Pond Snails With No Pore-Water 

Sampling)..................................................................................................................................7

Experiment 4 (Field Mesocosms With Pore-Water Sampling)......................................................8

Experiment 5 (Vegetated Microcosms With Snails) .......................................................................9

Laboratory Analytical Procedures for Experiments 1 and 2...................................................................9

Total Mercury ........................................................................................................................................9

Reactive Mercury ...............................................................................................................................10

Methylmercury....................................................................................................................................10

Sediment Bulk Density, Percent Dry Weight, Porosity, and Organic Content..........................10

Total Reduced Sulfur..........................................................................................................................10

Iron Speciation....................................................................................................................................11

Laboratory Analytical Procedures for Pore Water Collected From Experiments 1, 2, 4, and 5 

and Sediment From Experiments 4 and 5 ...................................................................................11

Results of the Laboratory and Field Experiments...................................................................................11

Chemical Characteristics of Littoral Sediment and Sediment Pore Water...............................11

Effects of Zero-Valent Iron on Pore-Water Chemistry From Littoral Sediment 

(Experiment 1) ........................................................................................................................12

Effects of Zero-Valent Iron on Pore-Water Chemistry From Profundal Sediment 

(Experiment 2) ........................................................................................................................12

Effects of Zero-Valent Iron on Sediment Chemistry and Uptake of Mercury in Snails 

Using Vegetated Littoral Sediment (Experiment 3)..........................................................12

Effects of Zero-Valent Iron and Granular Activated Carbon Treatments on Pore-Water 

and Sediment Chemistry in Field Mesocosms (Experiment 4).......................................17

Effects of Zero-Valent Iron and Granular Activated Carbon on Sediment Chemistry and 

Uptake of Mercury in Snails Using Vegetated Littoral Sediment (Experiment 5).......19

Discussion of Experimental Results .........................................................................................................19

Limitations of This Study and Potential for Future Research ...............................................................24

Summary........................................................................................................................................................24

References Cited..........................................................................................................................................25
vi

Figures

1. Map showing the watershed area draining to Hodgdon Pond, Mount Desert Island, 

Maine, and the location, dimensions, and hydrologic connectivity of wetlands within 

the catchment ...............................................................................................................................5

2. Graphs showing pore water concentration of A, ferrous iron for 5 and 50 times iron 

treatments, B, pH for 5 and 50 times iron treatments, C, ferrous iron collected from 

2, 4, 6, and 8 centimeters depth below the sediment-water interface, and D, pH collected from 2, 4, 6, and 8 centimeters depth below the sediment-water interface.........13

3. Plots showing the relation between concentrations of A, sediment methylmercury 

and amorphous trivalent iron and B, sediment MeHg and acid-extractable divalent 

iron, Experiment 3 for Mount Desert Island, Maine, for zero-valent iron treatment 

and untreated (reference) samples.........................................................................................17

4. Graphs showing concentrations of methylmercury, sulfide, and divalent iron in 

pore water from field mesocosms following A, untreated (reference; Experiment 4), 

and treatments of B, zero-valent iron treatment and C, granular activated carbon, 

sampled at Mount Desert Island, Maine, September 6 and November 11, 2013, at 

four depths below the sediment-water interface .................................................................18

5. Graph showing concentrations of methylmercury in pore water collected from 

field mesocosms from Mount Desert Island, Maine, on September 6 and 

November 11, 2013, for Experiment 4 ......................................................................................19

6. Graphs showing concentrations of total mercury, methylmercury, and MeHg:THg 

for sediment sampled from field mesocosms (Experiment 4) from Mount Desert 

Island, Maine...............................................................................................................................20

7. Graphs showing log of the distribution coefficient for concentrations of methylmercury between sediment and pore water sampled from field mesocosms 

(Experiment 4) for A, 25 days posttreatment; and B, 91 days posttreatment from 

Mount Desert Island, Maine .....................................................................................................21

8. Graph showing a comparison of concentrations of methylmercury in sediment 

(four samples) and pore water (two samples) sampled 125 and 154 days 

posttreatment of zero-valent iron and granular activated carbon in laboratory 

mesocosms containing vegetated sediment to untreated (reference) samples, 

Experiment 5, for Mount Desert Island, Maine......................................................................23

Tables

1. Chemical characteristics of littoral sediment and pore water extracted from 

sediment from wetlands fringing Hodgdon Pond, Mount Desert Island, Maine..............12

2. Pore-water chemistry from laboratory microcosms with blended littoral 

sediment (Experiment 1) from Mount Desert Island, Maine................................................14

3. Pore-water chemistry from laboratory microcosms with profundal sediment 

(Experiment 2) from Mount Desert Island, Maine .................................................................15

4. Summary of sediment and snail chemistry from Experiment 3 vegetated 

mesocosms from Mount Desert Island, Maine, with great pond snails............................16

5. Summary of sediment and great pond snail chemistry from Experiment 5 for 

Mount Desert Island, Maine .....................................................................................................22
vii

Conversion Factors

SI to Inch/Pound

Multiply By To obtain

Length

millimeter (mm) 0.03937 inch (in.)

centimeter (cm) 0.3937 inch (in.)

meter (m) 3.281 foot (ft)

kilometer (km) 0.6214 mile (mi)

meter (m) 1.094 yard (yd)

Area

square meter (m2) 0.0002471 acre

square meter (m2) 10.76 square foot (ft2)

hectare (ha) 2.471 acre

square centimeter (cm2) 0.1550 square inch (ft2)

Volume

cubic centimeter (cm3) 0.06102 cubic inch (in3)

liter (L) 1.057 quart (qt)

Mass

gram (g) 0.03527 ounce, avoirdupois (oz)

kilogram (kg) 2.205 pound avoirdupois (lb)

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as 

°F = (1.8 × °C) + 32.

Datum

Vertical coordinate information is referenced to the North American Vertical Datum of 1988 

(NAVD 88).

Horizontal coordinate information is referenced to North American Datum of 1983 (NAD 83).

Altitude, as used in this report, refers to distance above the vertical datum.

Supplemental Information

Specific conductance is given in microsiemens per centimeter at 25 degrees Celsius (μS/cm at 

25 °C).

Concentrations of chemical constituents in water are given either in milligrams per liter (mg/L) 

or micrograms per liter (μg/L).
viii

Acronyms, Abbreviations, and Chemical Elements

Acronyms

%LOI percent loss on ignition

ANP Acadia National Park

CVAFS cold vapor atomic fluorescence spectroscopy

DMA Direct Mercury Analyzer

DOC dissolved organic carbon

DOM dissolved organic matter

EPA U.S. Environmental Protection Agency

GAC granular activated carbon

HDPE high density polyethylene

KD sediment-pore water distribution coefficient

LDPE low density polyethylene

MOD moderately hard water

PTFE polytetrafluoroethylene, Teflon®

PVC polyvinyl chloride

RPD relative percent deviation

THg total mercury

TRS total reduced sulfur

USGS U.S. Geological Survey

ZVI zero-valent iron

Abbreviations

mg milligram

mg/kg milligram per kilogram

mL milliliter

ng nanogram

ng/g nanogram per gram (part per billion)

ng/g/d nanogram per gram (of dry sediment) per day; also ng/g dry sed/d

ng/L nanogram per liter

μg microgram

μg/g microgram per gram
ix

μg/kg microgram per kilogram

μL microliter

μm microns

μmol/L micromoles per liter

μS/cm microsiemens per centimeter

Chemical Elements and Compounds

BrCl bromium chloride

CaSO

4

•2H

2

O calcium sulfate dihydrate

Fe iron

Fe(II) divalent iron, ferrous iron, Fe2+

Fe(II)AE acid-extractable ferrous iron

Fe(III) trivalent iron, Fe3+

HCl hydrochloric acid

Hg mercury

Hg(0) elemental mercury

Hg(II) divalent mercury

Hg(II)

R

 reactive mercury

HgS mercury sulfide H2S hydrogen sulfide

KCl potassium chloride

MeHg methylmercury

MgSO

4

 magnesium sulfate

MOD moderately hard water

NaHCO

3

 sodium bicarbonate

NaOH sodium hydroxide

NH

2

OH–HCl hydroxylamine hydrochloride

ZnAc zinc acetate

ZnS zinc sulfide

Assessment of the Use of Sorbent Amendments for 

Reduction of Mercury Methylation in Wetland Sediment at 

Acadia National Park, Maine

By Thomas G. Huntington,1 Ariel Lewis,2 Aria Amirbahman,2 Mark Marvin-DiPasquale,1 and 

Charles W. Culbertson1

Abstract

Mercury is a contaminant of ecological concern 

because of its ubiquity and toxicity to fish and wildlife, and 

is considered a severe and ongoing threat to biota at Acadia 

National Park in Maine. The formation and biomagnification 

of methylmercury is the primary concern of resource 

managers at Acadia, and information is needed to develop 

strategies for remediation or mitigation of this contaminant. 

The U.S. Geological Survey in cooperation with Acadia 

National Park, National Park Service carried out a series of 

laboratory and field experiments to evaluate the potential 

of zero-valent iron and granular activated carbon to reduce 

the rate of the bacterially mediated process of mercury 

methylation and subsequent biological uptake by the great 

pond snail Lymnaea stagnalis. The addition of zero-valent iron 

resulted in an increase in ferrous iron that was then further 

oxidized to poorly crystalline amorphous ferric iron, as was 

anticipated. Our original hypothesis was that these reactions 

would reduce methylation by decreasing the concentrations 

of substrates for bacterial methylation (sulfide and divalent 

mercury) through sorption to ferric iron surfaces, formation 

of iron sulfide compounds, or conversion of mercury to 

gaseous forms and subsequent evasion. The results of our 

experiments did not consistently support this hypothesis. In 

one experiment the application of zero-valent iron increased 

the amount of methylmercury associated with surficial 

sediment. In another experiment zero-valent iron decreased the 

amount of methylmercury associated with surficial sediment. 

The addition of zero-valent iron may have stimulated 

mercury methylation by iron reducing bacteria and if that 

effect outweighed the processes that could have decreased 

methylation then methylation would not be decreased.

The results of field mesocosm experiments indicated that 

there was a decreasing trend in pore-water methylmercury 

concentration after application of granular activated carbon 

but methylation was not affected because there was no corresponding decrease in sediment methylmercury concentration. The application of granular activated carbon resulted in 

the sorption of methylmercury. The application of granular 

activated carbon resulted in an increase in the distribution 

coefficient for methylmercury indicating that this amendment 

caused a higher proportion of methylmercury to be associated 

with the sediment than the pore water in comparison to the 

reference (untreated) condition. Experiments to test whether 

zero-valent iron or granular activated carbon would reduce 

the biouptake of methylmercury in snails were inconsistent; 

zero-valent iron had no effect on uptake in one experiment but 

resulted in a significant decrease in uptake in a second experiment. Granular activated carbon did not affect biouptake in 

either experiment.

Introduction and Background

The introduction and background is divided into four 

sections. The first section describes the problem of mercury 

(Hg) contamination, the concept of biomagnification, previous 

studies documenting mercury contamination in biota at Acadia 

National Park. The first section also describes the process 

of microbial methylation of mercury and the purpose of this 

report. The second section describes mechanisms of potential 

remediation of mercury contamination using chemical amendments. The third section is a brief statement of the purpose and 

scope of the report. The final section is a brief site description. The purpose of this report is to provide Acadia National 

Park resource managers with information that could be used 

to develop strategies to reduce the methylation and biological 

uptake of methylmercury (MeHg).

1

U.S. Geological Survey.

2

University of Maine.
2 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

The Mercury Contamination Problem

Mercury is a widespread contaminant of ecological concern because of its ubiquity and toxicity to fish and wildlife 

in environmental settings where it accumulates in biota and 

is biomagnified. The adverse effects of mercury exposure are 

commonly associated with its neurotoxicity; however, mercury exposure also has been linked to immune suppression, 

endocrine disruption, physical malformations, and mortality 

in organisms (Eisler, 2006; Scheuhammer and others, 2007; 

Tan and others, 2009; Wada and others, 2009; Weiner and 

Spry, 1996; Wolfe and others, 1998). The threat of this toxin 

to human health has resulted in the establishment of consumption advisories for marine and freshwater fish in 44 States in 

the United States, including much of the east coast (including 

Maine) and all eastern Canadian Provinces.

Based on indirect evidence, such as mercury found 

in lake sediment, atmospheric deposition of mercury from 

nonpoint sources increased in the continental United States 

from the late 1800s to the 1990s by 2.5 to 15 fold (Driscoll 

and others, 2007) resulting in an increase in the extent of 

mercury contamination in fish in freshwater and coastal 

environments (Boesch and others, 2001). Since the 1990s 

measurements indicate decreases in wet deposition (from 

precipitation samples) of mercury over the continental United 

States averaging -4.1 plus or minus 0.49 percent per year for 

the period from 2004 to 2010 (Zhang and Jaeglé, 2013).

Biomagnification is the process by which the 

concentration of a contaminant increases from the 

environment at large through successively higher trophic 

levels within the food chain because of the contaminant being 

very slowly metabolized or excreted. Biomagnification of 

Hg is a severe and ongoing threat to biota at Acadia National 

Park (ANP; Acadia National Park, 2010), as elevated 

concentrations have been documented in amphibian, reptiles, 

birds, fish, and mammals (Bank and others, 2005, 2007a, b; 

Longcore and others, 2007). Average mercury concentrations 

in fish fillets from predator fish from 11 lakes at Acadia 

National Park exceeded health thresholds established for the 

safe consumption of fish by humans and wildlife (Acadia 

National Park, 2010). Biomagnification of mercury in the great 

pond snail Lymnaea stagnalis has been studied in microcosm 

experiments (Tessier and others, 2007).

In one study at Acadia National Park bullfrog and 

green frog tadpoles collected from nine ponds had median 

total mercury (THg) concentrations of 0.075 milligrams per 

kilogram (mg/kg) wet weight and the median percent MeHg 

was 22 percent (Bank and others, 2007b). Among these 

ponds, the dimictic and mesotrophic Hodgdon Pond has an 

average sediment mercury concen-tration of 0.22 mg/kg, 

and one of the highest water and fish mercury concentrations 

among Maine lakes. In summer 2010, Hodgdon Pond had 

average surface water THg and MeHg concentrations of 

2.63 nanograms per liter (ng/L) and 1.8 ng/L, respectively, 

and an average mercury body burden in the predatory 

smallmouth bass Micropterus dolomieu greater than (>) 

2 mg/kg wet weight (Linda Bacon and Aria Amirbahman, 

unpub. data, [February 15, 2014]), well in excess of the 

U.S. Environmental Protection Agency (EPA) 0.3 mg/kg 

fish consumption advisory for MeHg (U.S. Environmental 

Protection Agency, 2001). Mercury accumulation rates in bog 

and pond sediment at two locations at Acadia National Park 

during the 1980s were reported to be in the range of 100 to 

200 micrograms per square meter per year (μg/square meter/

year) (Norton and others, 1997). Amirbahman and others 

(2004) reported THg and MeHg concentrations of 180 grams 

per hectare (g/ha) and 0.16 g/ha, respectively, in undisturbed 

Acadia National Park forest soil. At Acadia National Park, 

wetland soil has significantly higher MeHg concentrations 

(about 3.0 micrograms per kilogram (μg/kg) dry weight) 

compared with the well-drained upland soil (about 0.2 μg/kg 

dry weight; Ruck, 2002).

The principal nonpoint source of Hg in the natural 

environment is atmospheric deposition, a major part of which 

originates from waste incineration and the combustion of 

fossil fuels. The annual rate of atmospheric wet deposition of 

THg in the northeastern United States is 6 to 8 micrograms 

per meter square meter (μg/square meter) (Prestbo and Gay, 

2009). Divalent mercury (Hg[II]) is the dominant form 

of mercury in atmospheric deposition. Once deposited in 

terrestrial or aquatic environments mercury can enter into 

complex biogeochemical processes (abiotic and biotic) in 

which it can be converted into other forms of mercury. From 

the standpoint of environmental contamination and biological 

toxicity, it is the formation of MeHg that is of most concern 

because this form bioaccumulates to as much as a factor of 

107 in the food chain to concentrations that are harmful or 

lethal (for example, Driscoll and others, 2007). It has been 

proposed that methylation is brought about by bacteria that 

have taken up small, uncharged mercury complexes (for 

example, neutral aqueous mercury sulfide (HgS0[aq]) and 

mercury disulfide (HgHS2

0

) capable of passive diffusion across 

bacterial cell membranes (Benoit and others, 1999, 2001). 

The produced MeHg may be taken up by phytoplankton or 

other organisms, which are then consumed by zooplankton. 

The factors controlling the rates of Hg(II) methylation and 

MeHg demethylation (Oremland and others, 1991) determine 

environments likely to lead to toxicity in biota.

Microbial methylation of Hg(II) to MeHg is affected 

by environmental factors, including temperature, dissolved 

oxygen concentrations/redox status, dissolved organic carbon 

(DOC) concentration and composition, sulfate, pH, and the 

activity of the sulfate reducing bacteria (Compeau and Bartha, 

1985; Gilmour and others, 1992; Benoit and others, 2003; 

Marvin-DiPasquale and others, 2009a,b) and trivalent iron 

(Fe[III]) reducing bacteria (Warner and others, 2003; Schafer 

and Morel, 2009). In the northeastern United States, there is 

a positive relation between concentrations of total organic 

carbon and total elemental mercury (Hg(0)) and MeHg in 

surface water (Dennis and others, 2005). Organic carbon in 

surface water that is in equilibrium with sediment pore water 

may enhance methylation in sediment. Macrophytes also may 
Introduction and Background 3

affect methylation in wetlands through effects on pH, redox 

potential (Eh) and by the exudation of organic ligands (Cosio 

and others, 2014).

Mercury methylation is reported to exist largely in the 

surficial sediment zone (Han and others, 2007; Sunderland and 

others, 2004). Mercury accumulation rates in ponds, however, 

are higher than in their surrounding watersheds because of 

accumulation of mercury from surrounding watershed runoff 

(Perry and others, 2005). Dissolved organic carbon can form 

complexes with Hg(II) and sulfide, thus facilitating transport of these precursors for methylation from terrestrial to 

aquatic environments.

Potential Remediation

Potential remediation of mercury contamination in ponds 

with the ultimate goal of decreasing MeHg uptake in higher 

trophic level biota could be accomplished through different 

mechanisms. One approach is to apply a chemical amendment 

that has the potential to reduce the concentration of Hg(II) 

available for Hg(II)-methylation by reduction to volatile 

elemental Hg(0) (Vernon and Bonzongo, 2014). Another 

approach is to add a complexation agent that has the potential 

to reduce the concentration of sulfides that are reported to 

facilitate Hg(II)-methylation by the formation of neutral mercury sulfide (HgS) complexes (Mehrotra and Sedlak, 2005). 

One of the more common approaches to remediation is to add 

a sorbent to adsorb Hg(II) and MeHg, thus decreasing the concentrations of the MeHg precursor and product (Weisener and 

others, 2005; Vernon and Bonzongo, 2014. Zero-valent iron 

(ZVI) has the potential to reduce Hg(II)-methylation through 

all of these mechanisms. Some amendments like biochar or 

granular activated carbon (GAC) can act as sorbents for Hg(II) 

and MeHg and may serve as reactive barriers to the movement 

of mercury in soil and sediment (Gilmour and others, 2013).

According to the model proposed by Benoit and coworkers (Benoit and others, 1999, 2001), one of the factors limiting 

Hg(II)-methylation is the concentration of the dissolved 

neutral HgS complexes, because only these complexes are 

capable of passive diffusion across bacterial cell membranes. 

The concentration of these complexes, in turn, is dependent on 

the concentration of dissolved Hg(II) and the concentration of 

dissolved sulfide. Anything that lowers the concentrations of 

either constituent or actively competes with either constituent 

to form other complexes or compounds that are not capable 

of diffusion across microbial membranes will slow the rate of 

Hg(II)-methylation (Benoit and others, 1999, 2001). Historical and ongoing mercury contamination from atmospheric 

deposition, as well as in place biogeochemical processes, 

result in a steady supply of mercury species that are available 

for Hg(II)-methylation (Ulrich and others, 2001). Natural and 

anthropogenic sources (acidic deposition) of sulfate provide 

a steady supply of sulfate for bacterial conversion to sulfide 

([sulfate reducers that methylate Hg] in anaerobic sediment 

Gilmour and Henry, 1991). In ponds that stratify during the 

summer, like Hodgdon Pond at Acadia National Park, surface 

sediment likely is suboxic or anoxic throughout the summer 

season, thus favoring anaerobic microbial processes, such as 

iron and sulfate reduction (Benoit and others, 2003; Kerin and 

others, 2006; Schafer and Morel, 2009). Therefore, to reduce 

Hg(II)-methylation one of the more promising remediation 

options would be to reduce the availability of Hg(II) otherwise 

available for Hg(II)-methylation and inhibit the formation of 

the neutral Hg-S complexes.

The addition of Fe(II) has been reported to inhibit 

Hg(II)-methylation in controlled laboratory studies with 

sediment slurries and mesocosms (Mehrotra and others, 

2003; Mehrotra and Sedlak, 2005, Liu and others, 2009). 

There are at least three mechanisms by which iron addition 

could reduce the net rate of Hg(II)-methylation. In the first 

mechanism, Fe(II) reduces the dissolved sulfide concentration 

through the formation of solid iron sulfide (FeS[s]) (Mehrota 

and others, 2003) to the extent that the formation of neutral 

HgS is inhibited (Ulrich and Sedlak, 2010), or charged 

Hg(II)-polysulfides are formed (Liu and others, 2009). A 

second mechanism is that the addition of iron results in the 

formation of amorphous iron hydroxide (FeOH3(s)), the surface 

of which is highly reactive and has a high affinity for Hg(II) 

(Tiffreau and others, 1995), thus decreasing the concentration 

of dissolved Hg(II) available for Hg(II)-methylation. A third 

mechanism holds that iron additions may cause Hg(II) to 

be reduced to elemental Hg(0) (Charlet and others, 2002; 

Peretyazhko and others, 2006; Amirbahman and others, 2013), 

decreasing the amount of Hg(II) available for methylation.

The addition of ZVI to suboxic aquatic sediment will 

cause a series of reactions that include production of hydrogen 

(H2) from water and co-reactions on iron surfaces whereby Fe0

is oxidized to ferrous, Fe(II), and ferric trivalent iron Fe(III). 

These reactions will elevate Fe(II) and Fe(III) concentrations 

that, in theory, will constitute a protective reactive barrier (an 

iron curtain), by the mechanisms mentioned above, to reduce 

the concentration of Hg(II) species available for methylation 

(that is neutral Hg-S complexes). ZVI has been reported to 

have potential as a protective reactive barrier for mercury in 

field column studies and is capable of accumulating mercury 

as a Hg-sulfide complex with a stoichiometry similar to 

the mercury sulfide minerals cinnabar and metacinnabar in 

contaminated groundwater (Weisener and others, 2005). The 

fate of the added ZVI may depend on the unique sediment 

redox conditions; in oxic environments, the presence of 

ZVI leads to the formation of Fe(III) species and FeOH3(s), 

whereas in anoxic environments, the presence of ZVI leads 

preferentially to the formation of Fe(II) species, including 

Fe-sulfide complexes, and FeS(s). ZVI should persist in 

surficial sediment, as the iron corrodes and fresh reactive 

surfaces continually form. Some of the conversion of ZVI to 

Fe(II) may be microbially mediated (Roden and Wetzel, 2002).

It was once considered that in suboxic wetland sediment 

that sulfate reducing bacteria were exclusively responsible 

for Hg(II)-methylation but it is now recognized that other 

bacteria including Fe-reducers and methanogens also are 

capable of methylating mercury (Kerin and others, 2006; 
4 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

Fleming and others, 2006, Hamelin and others, 2011). It also 

was reported that transport across the cell membrane was 

passive but it has recently been reported that it can be an 

active process (Schaefer and others, 2011). Which group of 

bacteria is responsible for Hg(II)-methylation in any given 

environment is probably dependent on several factors and is 

not easy to predict. Bacteria also demethylate mercury and 

it is the net rate of MeHg production that will determine the 

potential for biological uptake (Marvin-DiPasquale and others, 

2000). Because methylation and demethylation are complex 

interacting processes dependent on many environmental 

variables it is difficult to predict the effect of a given 

amendment therefore experimentation is necessary.

The application of sorbent amendments in the field 

is emerging as a potential approach for the remediation or 

management of mercury-contaminated sediment (Ghosh 

and others, 2011). Sorbents bind contaminants and thereby 

reduce bioavailability. Sorbents can be applied as a cap where 

they can act as both sorbent and as a reactive barrier to the 

transport of the contaminant or they can be incorporated into 

contaminated sediment to maximize contaminant binding. 

There are potential pitfalls with both strategies, such as 

adverse effects on benthic fauna and flora. Mercury often 

is complexed with dissolved organic matter (Hg-DOM) 

in natural systems and DOM has been determined to have 

a strong affinity for activated carbon surfaces (Schreiber 

and others, 2005; Gerbig and others, 2011). Activated 

carbonaceous sorbents have been determined to reduce porewater contaminant concentrations (Gomez-Eyles and others, 

2013) and reduce biouptake of mercury and MeHg (Ghosh and 

others, 2008, Gilmour and others, 2013).

There have been a few studies evaluating the potential 

for surface application of GAC to reduce pore-water mercury 

and MeHg and biouptake of these species (Ghosh and others, 

2008, Gilmour and others, 2013), but no studies that we know 

of that have evaluated the potential for ZVI to reduce biouptake of MeHg. GAC studies usually have involved laboratory 

microcosms and not field studies, typically using homogenized 

and unvegetated sediment or sediment slurries. In place studies are much less common. This study was carried out as a 

preliminary evaluation of the potential for GAC and ZVI to 

reduce pore-water mercury and MeHg concentrations in place 

in freshwater wetland sediment. An additional purpose for this 

study was to evaluate the potential for GAC and ZVI to reduce 

mercury and MeHg concentrations in pore water and surficial 

bulk sediment in microcosms containing intact vegetated 

sediment in the laboratory. Finally, this study evaluated the 

potential for GAC and ZVI to reduce mercury and MeHg 

biouptake in the great pond snail L. stagnalis in microcosms 

containing intact vegetated sediment in the laboratory.

Surface applications of ZVI or GAC may be cost 

effective for treating surface sediment that are contaminated 

with atmospherically deposited mercury compared with 

treating large volumes of heavily contaminated soil or 

sediment. ZVI can be purchased for about $1,100 per metric 

ton (Amy Marchefka, ConnellyGPM, Inc., written commun., 

July 28, 2014) and GAC could be purchased for about 

$2,100 per metric ton (Ghosh and others, 2011). To treat 

1 hectare (ha) with ZVI at the rate used in the field experiment 

in this study would require 864 kg. To treat 1 ha with GAC 

at the rate used in the field experiment in this study would 

require 802 kg. Actual costs would include costs associated 

with shipping and application.

Purpose and Scope

The purpose of this report is to provide Acadia National 

Park resource managers with information that could be used 

to develop strategies to reduce the methylation and biological uptake of MeHg. This report describes a series of laboratory and field experiments using wetland sediment from 

areas fringing Hodgdon Pond on Mount Desert Island to test 

if amendments including ZVI and GAC had the potential to 

reduce the methylation of mercury and biological uptake of 

MeHg. This report includes descriptions of the effects of ZVI 

and GAC additions on pore-water chemistry at various depths 

below the sediment water interface, sediment chemistry, and 

mercury uptake in the great pond snail L. stagnalis.

Site Description

Hodgdon Pond in Tremont Township (not shown) on 

Mount Desert Island, Maine, was the subject of this study. 

The pond has a surface area of 18.2 ha and a perimeter of 

3,300 meters (fig. 1) and fringing wetlands near the outlet 

at the southern end, on the western shore near the midpoint, 

and, at the northern end. Most of the Hodgdon Pond watershed lies within Acadia National Park, but the western shore 

is privately owned. Several wetlands have been identified 

within the Hodgdon Pond watershed that include wetlands 

that fringe the pond perimeter (lacustrine wetlands) as well as 

riverine, depressional, and slope wetlands in the larger pond 

catchment ( Nielsen, 2006, fig. 1). The nonlacustrine wetlands 

are located in first-, second-, or third-order catchments nested 

within the larger pond catchments and are hydrologically connected through ephemeral and perennial channels that drain 

into the pond (Nielsen, 2006, plate 1). These wetlands are 

likely important to the levels of MeHg observed in Hodgdon 

Pond because wetlands are estimated to contribute as much as 

80 times more MeHg to receiving water bodies than upland 

areas (St. Louis and others, 1994).

Hodgdon Pond is 6.7 meters (m) deep at the deepest point 

and has an average depth 3.35 m (Maine.gov, [2001]). The 

pond drains from north to south into the substantially larger 

Seal Cove Pond. The water level is variable depending on the 

season and antecedent precipitation conditions. Water quality 

data indicate a pH in the range of 6.2 to 6.5 and conductivity of 41 to 53 microsiemens per centimeter (μS/cm) at 25 °C 

(Bacon, 2013).
Introduction and Background 5

384

396

364-b

366

360

381-b

381-a

Long

Pond

Base from U.S. Geological Survey digital line graph

Bartlett Island, 1981, 1:24,000

EXPLANATION

Wetland and identifying

number (Nielsen, 2006)

Wetland catchment

First order

Second order

Third order

Largest

Atlantic

Ocean

Hodgdon

Pond

Seal

Cove

Pond

44°19'

44°19'30''

44°20'

68°24'30'' 68°24' 68°23'30'' 68°23'

366

Mt. Desert

Island

Location

map

Topographic contour—

Contour interval 6 meters

120

90

60

60

60

60

30

30

30

30

30

Location of sediment associated

with the indicated experiment number

Modified from Nielsen (2006)

362

376

3,5

4

2

1

2

0.25 0.5 MILE

0 0.25 0.5 KILOMETER

0

Figure 1. The watershed area draining to Hodgdon Pond, Mount Desert Island, Maine, and the location, 

dimensions, and hydrologic connectivity of wetlands within the catchment. The locations where the sediment 

was collected or instrumented is also shown with numbers corresponding to the experiments. This figure was 

modified from Nielsen (2006). The wetland identifier numbers are described in Nielsen (2006, table 2–1).
6 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

Methods

Five experiments were carried out for this project. In the 

first experiment, sediment from the littoral zone of wetlands 

fringing Hodgdon Pond were maintained in cylindrical 

microcosms and treated with two levels of ZVI addition. In 

the second experiment, sediment from the profundal zone of 

Hodgdon Pond were maintained in cylindrical microcosms 

and treated with one level of ZVI addition. In the third experiment, vegetated intact sediment clumps from the littoral zone 

of wetlands fringing Hodgdon Pond were maintained in cylindrical microcosms and treated with one level of ZVI addition. 

Great pond snails (L. stagnalis) were used in this experiment 

to evaluate biouptake of Hg. Great pond snails were selected 

as test organisms because they are relatively easy to raise and 

have been used successfully in short-term exposure metal 

uptake studies (Tessier and others, 2007, Croteau and others, 

2009, 2011, Sychra and others, 2011). The microbial communities in periphyton (biofilms) can accumulate and methylate 

Hg (Desrosiers and others, 2006; Hamelin and others, 2011). 

Great pond snails consume periphyton (Farag and others, 

1998) so that these snails are useful for studying bioaccumulation of Hg and MeHg in animals. In the fourth experiment, 

mesocosms were established in the littoral zone of wetlands 

fringing Hodgdon Pond and maintained from August through 

November 2013. In this experiment, one treatment involved 

the application of ZVI and another treatment involved the 

application of GAC. In the fifth experiment, vegetated intact 

sediment clumps from the littoral zone of wetlands fringing 

Hodgdon Pond were maintained in cylindrical microcosms 

and treated with one level of ZVI addition and one level of 

GAC addition. Great pond snails (L. stagnalis) were used in 

this experiment to evaluate biouptake of Hg.

Experiment 1 (Littoral Zone Sediment)

Sediment was collected from fringing wetlands 

in the littoral zone midpoint eastern edge of Hodgdon 

Pond on November 18, 2011 (latitude 44° 19′ 18.9′′, 

longitude –68° 23′ 46.8′′). Intact columns 20-cm deep and 

10-cm diameter were collected in clear plastic cylinders and 

were brought to the laboratory. Core material appeared to 

be extremely heterogeneous containing many coarse (1 to 

1.5 cm diameter) roots and large voids (some filled with 

air and some with water). Because of the heterogeneity we 

concluded that Rhizon samplers (Rhizosphere Research 

Products) would be ineffective. All sediment was aggregated 

and the large roots were removed and then the sediment was 

blended to create a uniform and well-mixed material and was 

stored refrigerated at 3 degrees Celsius (°C). The blended 

sediment was then repacked in the original 10-cm diameter 

columns on February 2, 2012.

A subsample of the blended sediment was maintained 

refrigerated at 3 °C until March 21, 2012, when it was shipped 

on ice to the U.S. Geological Survey laboratory in Menlo 

Park, California, for geochemical analysis. Upon arrival, the 

sample was subsampled in an anoxic glove bag for the following suite of whole sediment constituents: THg, MeHg, reactive 

mercury (Hg(II)R), percent dry weight, organic content (as percent loss on ignition [percent LOI]), grain size, total reduced 

sulfur (TRS) and iron speciation. The subsamples for THg, 

MeHg, Hg(II)R, TRS and iron-speciation were all subsequently 

preserved frozen. The subsamples for dry weight, percent LOI 

and grain size were all subsequently preserved refrigerated. In 

addition, approximately 100 cubic centimeters (cm3) of sediment was centrifuged and the supernatant subsequently filtered 

(0.45 microns [μm]) in an anoxic glove bag for pore-water 

THg and MeHg. Both of these fractions were then preserved 

with 0.5 percent trace metal clean hydrochloric acid (HCl). 

Total iron as used in this report refers to the sum of the three 

fractions that were determined by speciation, acid-extractable 

divalent ferrous iron (Fe(II)AE), amorphous trivalent ferric iron 

(Fe(III)a), and crystalline trivalent ferric (Fe(III)c).

Columns were instrumented with Rhizon pore-water 

samplers inserted through ports machined into the sidewalls of 

the 10-cm diameter cylinders. The Rhizons were connected to 

plastic tubing leading to a luer-lock syringe port. The pore size 

of the porous part of the Rhizons is 0.12 to 0.18 μm, hence 

filtering the samples before analyzing was unnecessary. The 

sediment was covered with approximately 3 cm of Hodgdon 

Pond water that was maintained at this depth with additions of 

Hodgdon Pond water as needed. The overlying water was aerated with aquarium pumps. Two columns were left untreated 

as reference treatments, two columns received 0.384 g of ZVI 

(5 times native iron) and two columns received 3.84 g ZVI 

(50 times native iron). The ZVI was incorporated into the top 

3 plus or minus (±) 1 cm of sediment using a glass stirring rod 

to gently mix the sediment. Rhizon samplers were installed 

through ports in column walls at depths of approximately 2, 

4, 6, and 8 cm below the sediment water interface for collection of pore water. Pore water was sampled 12, 42, 53, 75, and 

95 days following the addition of the treatments using syringes 

to apply suction to the Rhizon samplers.

Rhizon-collected pore-water samples were collected in 

two separate syringe pulls. The first pull of approximately 

3 milliliters (mL) was allocated to ancillary chemistry: Fe(II), 

TOC, dissolved hydrogen sulfide, and pH; 4-mL plastic bottles 

were prepared for preservation of each sample. For Fe(II), low 

density polyethylene (LDPE) bottles containing 100 microliters (μL) of 1 Molar nitric acid (1M HNO3

) samples were 

frozen immediately after collection; for TOC, high density 

polyethylene (HDPE) vials were used and frozen immediately after taking a pH measurement; for dissolved hydrogen 

sulfide, low density polyethylene (LDPE) vials that contained 

100 μL of 1M zinc acetate (ZnAc) in 2 Molar sodium hydroxide (2M NaOH) were frozen immediately after collection. The 

second pore-water sample was collected for THg measurement. Samples were collected in pre-weighed 15 mL polypropylene falcon tubes and preserved with 10-percent bromine 

chloride (BrCl). Samples were refrigerated until analysis.
Methods 7

Experiment 2 (Profundal Zone Sediment)

Sediment was collected from the profundal zone of 

Hodgdon Pond on May 30, 2012, in about 6 m of water (latitude 44° 19′ 24.35′′, longitude –68° 23′ 50.91′′). A box-corer 

was used to collect the surface sediment (nominally from the 

top 5 cm) from 10 locations within an area of about 60 meters 

squared (m2) and the individual samples were composited and 

well mixed to obtain a representative sample. The sediment 

was packed in six 15-cm diameter cylinders (microcosms) to 

a total depth of 16 cm. These microcosms were constructed 

using plastic water jugs.

After a 31-day equilibration period two columns were 

left untreated as reference treatments, two columns received 

0.878 g of ZVI (5 times native iron), and two columns 

received 8.73 g ZVI (50 times native iron). The ZVI was 

incorporated into the top 3 cm of sediment. All columns had 

2 cm sediment above the topmost port that was set in the 

middle of the layer that the ZVI. Aquarium pumps were used 

to keep the overlying water aerated from July 2 through 12, 

2012, and after this initial period the aeration was removed.

Each microcosm was fitted with an equilibrium dialysis 

frame with draw ports in the back (known as a draw-peeper) 

for pore-water sampling. Every other cell (beginning with 

the bottom-most cell) of the draw-peeper was modified by 

drilling a hole, threading it, and installing a connector to adapt 

this opening to a 1/16-inch polytetrafluoroethylene, Teflon®

(PTFE) tube that reached the surface, and connected to a 

luer-lock on the for pulling in-place samples. The holes were 

staggered to prevent stress on the draw-peeper’s frame. This 

resulted in a total of four draw-cells per draw-peeper.

Draw-peepers were acid-leached (10-percent HCl) then 

rinsed by soaking for 2 weeks in a deionized water (D.I. H2O) 

bath in which the water was changed every 3 days. Frames 

were assembled by submerging the base plate in D.I. H2O, 

placing a precut polysulfone dialysis membrane (0.45 μm) 

across the filled cells and then securing the face plate with 

nylon screws. The frames were immersed in portable tank 

columns and bubbled with nitrogen N2 for 2 weeks before 

deployment (Carignan and others, 1994). Immediately 

following deoxygenation draw-peepers were inserted into 

the sediment of each of the microcosms on May 30, 2012, 

so that the uppermost draw-cell was centered about 2 cm 

below the sediment water interface. The other draw cells 

were at approximately 4, 6, and 8 cm below the sediment 

water interface. Pore water was sampled before application of 

treatments, and at 10 and 133 days following the addition of 

the treatments using syringes to apply suction.

Draw-peeper pore-water samples were collected in 

three separate syringe pulls. The time between syringe pulls 

varied depending on the length of time it took for pore water 

to be extracted from each draw-peeper, but was usually less 

than 5 minutes. The first pull of approximately 3 mL was 

allocated to ancillary chemistry and preserved as described for 

Experiment 1. The second pore-water sample was allocated 

for MeHg. Samples were collected in pre-weighed 15 mL 

polypropylene falcon tubes and preserved with 0.5-percent 

HCl. Samples were frozen until analysis. The one-third porewater pull was collected for THg measurement. Samples 

were collected in pre-weighed 15 mL polypropylene falcon 

tubes and preserved with 10-percent BrCl. Samples were 

refrigerated until analysis.

Experiment 3 (Vegetated Microcosms With 

Great Pond Snails With No Pore-Water 

Sampling)

Hodgdon Pond vegetated intact sediment clumps or 

“plugs” were collected from fringing wetlands near the 

outlet of Hodgdon Pond (44° 18′ 57′′, -68° 23′ 50′′) on 

September 10, 2012. The sediment clumps were bowl-shaped 

and approximately 15 to 20 cm in width and 15 cm in height. 

The sediment was fully saturated and was collected under 2 to 

3 cm of overlying water. The vegetation in the clumps was 

American bur-reed (Sparganium americanum) and common 

rush (Juncus effusus). The clumps were brought to the USGS 

Augusta, Maine, laboratory where the sediment clumps were 

trimmed to fit and placed in 8 1-gallon acid-washed glass 

aquaria and covered with approximately 1 L of Hodgdon Pond 

water so that the sediment was covered with 5 cm of water. 

The microcosms were maintained on a 16 hour light and 

8 hour dark schedule. The vegetation initially was clipped to 

a height of 30 cm. The weight of saturated sediment in each 

aquaria was between 3.17 and 3.47 kilograms (kg) (average 

3.31 kg). Water was added regularly to maintain the water 

level to replace water lost to evaporation and transpiration. 

The specific conductance of the overlying water never 

exceeded 170 microsiemens per centimeter.

In Experiment 3, run 1, the microcosms were allowed to 

equilibrate for 35 days. Following this equilibration period 

1.36 g of ZVI was added to 4 of the microcosms (5 times 

the estimated amount of total Fe in the sediment to a depth 

of 3 cm) and incorporated into the surface 3 ±1 cm sediment 

using a glass stirring rod to gently incorporate the ZVI 

in the sediment around the existing vegetation stems and 

roots. The other 4 microcosms were maintained as reference 

treatments and the surface sediment also was stirred in 

these reference treatments. The microcosms were allowed 

to equilibrate with the ZVI treatments for 121 days. During 

this period the water levels were maintained using Hodgdon 

Pond water or moderately hard water (MOD) composed of 

sodium bicarbonate (NaHCO3), calcium sulfate dihydrate 

(CaSO4•2H2O), magnesium sulfate (MgSO4), and potassium 

chloride (KCl) (U.S. Environmental Protection Agency, 

2002b). The specific conductance of the overlying water never 

exceeded 170 μS/cm at 25 °C during this treatment period 

and the specific conductance of the MOD water was within 

20 percent of the Hodgdon Pond water.

Great pond snails (L. stagnalis) were selected for 

mercury biouptake experiments because they have been 

used successfully as test organisms to assess uptake of trace 
8 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

metals including mercury in laboratory microcosms (Croteau 

and Luoma, 2009; Croteau and others, 2011; Tessier and 

others, 2007). Adult snails were provided by Marie-Noële 

Croteau (U.S. Geological Survey, Menlo Park, California). 

Snails were reared in the laboratory in MOD described 

above and fed a diet of romaine lettuce. On February 13, 

2013, a cohort of juvenile snails of a restricted size range 

(shell lengths 10 to 15 millimeters (mm) (mean soft tissue 

dry weight of 7.8 ±0.5 milligrams [mg]) was assumed based 

on size and weight relations determined by M.-N. Croteau 

(U.S. Geological Survey, oral commun., [January 12, 2012]) 

was transferred to a glass tray and sorted by size to obtain 

12 snails of approximately equal sizes for use in each 

microcosm. The vegetation in the microcosm was clipped to 

5 cm above the aquarium surface. Twelve snails were then 

placed in each of the 8 microcosms and the microcosms were 

covered with plastic wrap to prevent snails from escaping. 

During the period when the snails were maintained in the 

microcosms no food was added so the snails had a diet of 

periphyton and detritus. The snails were maintained in the 

microcosms for a 28-day exposure period after which they 

were collected and placed in 1-quart glass jars filled with 

MOD and allowed to depurate for 48 hours. After depuration, 

the snails were frozen. Partially thawed L. stagnalis were 

dissected to remove the soft tissue that was refrozen and later 

freeze dried. At the end of the exposure period the surface 

3 ±1 cm of sediment was sampled from each microcosm and 

frozen until it was analyzed. Pore water was not sampled 

from the mesocosms before or during the exposure period in 

this experiment.

In Experiment 3, run 2, this experiment was repeated 

using the same vegetated mesocosms but with doubling the 

amount of ZVI addition. The reference mesocosms were 

maintained as in run 1 of this experiment. On March 13, 2013, 

an additional 1.36 g of ZVI was added to the previously ZVItreated mesocosms so that at the start of this experiment these 

mesocosms had received 10 times the estimated amount of 

total Fe in the sediment to a depth of 3 cm. The microcosms 

were maintained on a 16 hour light and 8 hour dark schedule. 

In Experiment 3, run 2, the equilibration period following the 

addition of ZVI was reduced to a single day. On March 14, 

2013, 10 juvenile snails (shell lengths 10 to 15 mm) were 

placed in each of the 8 microcosms, and the microcosms were 

covered with plastic wrap to prevent snails from escaping. 

The snails were maintained in the microcosms for a 25-day 

exposure period after which they were allowed to depurate for 

48 hours and then frozen. Partially thawed L. stagnalis were 

dissected to remove the soft tissue that was refrozen and later 

freeze dried. At the end of the exposure period, the surface 

3 ±1 cm of sediment was sampled from each microcosm and 

frozen until it was analyzed.

The sediment samples collected at the end of the snail 

exposure periods from this experiment were shipped on ice 

to the USGS Menlo Park laboratory on July 10, 2013. At the 

USGS Menlo Park laboratory all sediment samples initially 

were freeze dried and stored in a desiccator before analysis for 

sediment MeHg, Hg(II)R, and iron-speciation. The frozen snail 

samples from this mesocosm experiment were freeze dried and 

shipped to the USGS Menlo Park laboratory where they were 

subsequently stored in a desiccator before analysis for MeHg.

Experiment 4 (Field Mesocosms With PoreWater Sampling)

On August 12, 2013, 12 mesocosms were established in 

the fringing marsh near the outlet of Hodgdon Pond in a line 

adjacent to and parallel with the outlet channel centered on 

latitude 44° 19′ and longitude –68° 23′ 50.6′′. The treatments 

were randomly assigned along this line to points that were 

2 m apart. The mesocosms were established in vegetated sites. 

Polyvinyl chloride cylinders of approximately 38-cm diameter 

and 39-cm length were driven into the sediment where the 

water depth was a few centimeters above the sediment surface. 

To drive the cylinders into the sediment, rubber mallets were 

used and keyhole saws were used to cut around the outside of 

the cylinders so that they could penetrate the thick roots and 

woody detrital materials. The cylinders were driven into sediment to a depth of about 15 cm deep. Sediment samples were 

collected from each mesocosm before the application of treatments. GAC was added to four mesocosms, ZVI was added to 

four mesocosms, and four mesocosms were left untreated as 

reference treatments. Four mesocosms received 9.8 g ZVI to 

attain an estimated 8.9 times the amount of total iron present 

in the top 3 cm of the sediment. Four mesocosms received 

9.1 g GAC to attain an estimated concentration of 1.7 percent 

GAC by dry weight basis in the top 3 cm of sediment. Four 

mesocosms were left untreated as reference treatments. The 

ZVI and GAC were mixed into the sediment to a depth of 

about 3 cm and the surface 3 cm of the reference treatments 

was similarly mixed but with no additions.

Draw-peepers, as described above, for pore-water 

sampling were installed vertically in 2 reference, 2 ZVItreated, and 2 GAC-treated mesocosms so that the sampling 

depths were approximately 1.5, 3.5, 5.5 and 7.5 cm below the 

sediment water interface. Pore-water samples were collected 

in two separate syringe pulls. The first pull of approximately 

3 mL was allocated to ancillary chemistry and preserved as 

described for Experiment 1. However, pH measurements were 

not collected. The second pore-water sample was allocated 

for MeHg and THg. Samples were collected in pre-weighed 

15 mL polypropylene falcon tubes and were flash frozen and 

preserved with 0.5-percent HCl within 6 hours of return to 

the laboratory. After preservation, all samples were frozen 

until analysis. After an adequate sample was used for MeHg 

analysis, the remaining sample was preserved with 0.5 mL of 

BrCl, and refrigerated for THg analysis.

Pore water was sampled at various depths 25 and 91 days 

following the addition of the treatments. The surface 3 cm 

of sediment was sampled on August 12, 2013, prior the 

application of ZVI and GAC and 25 and 91 days following the 

addition of the treatments. In some cases the sample volumes 
Laboratory Analytical Procedures for Experiments 1 and 2 9

were too low to permit analysis. The sediment-pore water 

distribution coefficient for MeHg (KD [L/kg]) was calculated 

as the ratio of the dry weight sediment MeHg concentration to 

the pore-water MeHg concentration:

K Sediment MeHg concentration

Dissolved pore-water D

L

kg





 



 = ( )MeHg concentration

. (1)

The MeHg KD was calculated for the mesocosms that 

were sampled for pore-water MeHg concentration at various 

depths 25 and 91 days following the addition of the treatments.

Sediment samples were flash frozen in the field in 50-mL 

centrifuge tubes, delivered to the laboratory, semithawed in the 

refrigerator, homogenized in whirl packs, and then refrozen in 

their respective 50-mL tubes. Subsamples of these were collected by thawing the sediment in the refrigerator and pouring 

the slurry into a 15-mL polypropylene falcon tube. These subsamples were freeze dried for >50 hours using a Lyph-Lock 

4.5 freeze drier.

Experiment 5 (Vegetated Microcosms With 

Snails)

On October 14, 2013, Hodgdon Pond vegetated intact 

sediment clumps were collected from fringing wetlands 

near the pond outlet and were placed in 12 1-gallon glass 

aquaria and returned to the laboratory. Hodgdon Pond water 

was added to each aquarium (microcosm) to a depth of 

approximately 10 cm. The microcosms were maintained on 

a 16 hour light and 8 hour dark schedule. Treatments were 

applied and draw-peepers were installed on October 18, 2013. 

Four mesocosms received 1.4 g ZVI to attain an estimated 

5 times the amount of total Fe present in the top 3 cm of the 

sediment. Four mesocosms received 1.3 g GAC to attain an 

estimated concentration of 0.96 percent GAC by dry weight 

basis in the top 3 cm of sediment. Four mesocosms were left 

untreated as reference treatments. The ZVI and GAC were 

mixed into the sediment to a depth of about 3 cm and the 

surface 3 cm of the reference treatments was similarly mixed 

but with no additions. The draw-peepers were installed in two 

mesocosms that were treated with ZVI, two that were treated 

with GAC, and in two of the reference treatments.

Great pond snails (L. stagnalis) were reared in the 

laboratory in MOD and fed a diet of romaine lettuce. On 

February 7, 2014 (116 days following establishment of the 

microcosms and 112 days following application of treatments) 

10 juvenile snails of a restricted size range (shell lengths 

10 to 15 mm) were placed in each microcosm. The tops of the 

aquaria were wrapped with plastic to ensure that the snails 

did not escape. The snails were removed from the mesocosms 

on March 20, 2014, after an exposure period of 41 days. The 

snails were allowed to depurate for 24 hours and were then 

frozen, before dissecting to remove tissue and the tissue was 

freeze dried before analysis. At the end of the exposure period, 

3 cm of surface sediment was sampled from each microcosm 

and frozen until it was analyzed. Pore water was sampled 

at 2 cm below the sediment/water interface and sediment 

(top 3 cm) was sampled 125, 154, and 174 days following 

the addition of the treatments using the same procedures as 

described for Experiment 4.

Laboratory Analytical Procedures for 

Experiments 1 and 2

Total Mercury

Sediment THg associated with the single field sample 

collected during November 2011 was assayed in duplicate 

using a standard USGS approved method (Olund and others, 

2004). Approximately 0.2 g of sediment was initially digested 

with aqua regia (2 mL of concentrated HNO3 and 6 mL of 

concentrated HCl) in Teflon bombs overnight at room temperature. Subsequently, 22 mL of 5-percent BrCl was added 

to each sample, and the bombs were heated to 50 °C in an 

oven overnight. Once cooled, a 5-mL subsample was transferred into a precombusted glass container. The digestate was 

analyzed on an Automated Mercury Analyzer (Tekran Model 

2600, Tekran, Inc., Canada), according to EPA method 1631, 

revision E (U.S. Environmental Protection Agency, 2002). 

This method is based on the stannous (tin) chloride (SnCl2) 

reduction of Hg(II) to gaseous Hg(0), trapping Hg(0) on goldplated sand, thermal desorption, and quantification of Hg(0) 

by way of cold vapor atomic fluorescence spectrometry.

The method detection limit at the level of the Tekran analyzer was 0.10 nanograms per liter (ng/L) of aqueous digest, 

giving a sediment dry weight detection limit of approximately 

0.5 nanograms per gram (ng/g) (as run), depending on the 

wet sediment mass digested and its water content. This single 

sample was assayed with a larger batch of sediment samples 

from another project. The quality assurance metrics for that 

batch run included the following: (a) recovery of the certified 

reference material (PACS 2, marine sediment, THg certified 

value of 3.04 microgram per gram (μg/g) was 89.9 percent 

(n=1), (b) matrix spike recovery was 114±17 percent (n=2), 

and (c) the relative percent deviation (RPD) for duplicate 

analyses was 2.9±2.6 percent (n=3).

Pore-water THg was assayed on the November 2011 

sample after heated digestion with BrCl oxidation followed 

by SnCl2 reduction, purge and trap quantification of gaseous 

elemental Hg0, as per EPA method 1631 (U.S. Environmental 

Protection Agency, 2002). Quantification was carried out on 

an Automated Mercury Analyzer (Tekran Model 2600, Tekran, 

Inc., Canada), as described above. The quality assurance 

metrics for that batch run included the following: (a) recovery 
10 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

of the certified reference material (NIST 1641d, THg certified value of 15.9 ng/L) was 113 percent (n=1), and (b) matrix 

spike recovery was 96.7±0.4 percent (n=2). Because of limited 

volume, this sample was not assayed in duplicate.

Reactive Mercury

Sediment Hg(II)R is methodologically defined as the fraction of total Hg(II), that has not been chemically altered (for 

example, digested, oxidized, or chemically preserved apart 

from freezing), that is readily reduced to elemental Hg(0) by 

an excess of SnCl2 for an exposure time of 15 minutes. This 

operationally defined parameter was developed as a surrogate measure of the fraction of inorganic Hg(II) that is most 

likely available to Hg(II)-methylating bacteria responsible 

for MeHg production (Marvin-DiPasquale and Cox, 2007; 

Marvin-DiPasquale and others, 2014). The single whole sediment sample collected during 2011 was assayed in duplicate, 

with subsample weights of 0.55 g and 0.81 g. The freeze dried 

samples from the 2013 field and mesocosm experiments had 

subsample weights ranging from 0.11–0.83 g. Assay conditions were run as described in Marvin-DiPasquale and Cox, 

(2007). There is no certified reference material for this assay 

and matrix spikes are not appropriate because of the operationally defined nature of the assay. Quality controls included 

bubbler blanks, calibration standards prepared from a NISTcertified commercial HgCl2

 stock, and analytical duplicates. 

The RPD for analytical duplicates was 18.5 percent for the 

2011 sample and 6.5±1.8 percent (n=2) for the 2013 freeze 

dried samples.

Methylmercury

Methylmercury in both sediment and snail tissue was 

assayed by first digesting in an extractant of 25 percent 

potassium hydroxide (KOH) in methanol in a 60 °C oven for 

4 hours (Florida Department of Environmental Protection, 

2011) followed by ethylation with sodium tetraethyl borate 

of a diluted subsample in trace metal clean I-Chem vials 

and quantification on an automated MeHg analysis system 

(Brooks Rand Laboratories, Seattle, Washington) using coldvapor atomic fluorescence spectrometry detection. Sediment 

subsample masses used for analysis were approximately 0.5 g 

wet weight for the 2011 whole sediment sample, 0.1 g dry 

weight for the 2013 sediment samples, and 0.0015–0.022 g for 

the snail tissue. Quality assurance included the following: (a) 

certified reference material, (b) matrix spikes, (c) analytical 

duplicates, (d) method blanks, and (e) calibration standards 

prepared from commercial crystalline methlymercuric chloride 

and commercial MeHg standard solution.

The method detection limit at the level of the MERX 

analyzer is approximately 0.4 picograms of MeHg (absolute), 

and the detection limit of any particular sediment or tissue 

sample is fully dependent on the sample dry mass digested and 

subsample volume used. No samples were below detection at 

the level of the MERX analyzer. For sediment samples, the 

average (± standard deviation) recovery of the certified reference material (CC–580, estuarine sediment, MeHg certified 

value of 75 ng/g) was 106±6 percent (n=4, both batches combined). The average matrix spike recovery was 101±5 percent 

(n=8, both batches combined). The average RPD for duplicate 

analyses was 4.9±5.3 percent (n=4, both batches combined).

For snail samples, the average (± standard deviation) 

recovery of the certified reference material (TORT-2, Lobster 

Hepatopancreas, MeHg certified value of 152 ng/g) was 

92±1 percent (n=2). The average matrix spike recovery was 

98±4 percent (n=4). The average RPD for duplicate analyses 

was 2.6±1.5 percent (n=2).

Pore-water MeHg was assayed on the November 

2011 sample after distillation (DeWild and others, 2001), 

followed by ethylation and quantification on an automated 

MeHg analysis system (Brooks Rand Laboratories, Seattle, 

Washington), as described above. The quality assurance 

metrics for that batch run included matrix spike recovery, 

which was 105±11 percent (n=3). Due to limited volume, this 

sample was not assayed in duplicate, however RPD values for 

other samples in that batch run were 14.2±13.0 percent (n=3). 

Commercial certified reference material was not available for 

this assay.

Sediment Bulk Density, Percent Dry Weight, 

Porosity, and Organic Content

Sediment bulk density was assayed on the 2011 sediment 

sample first, precisely determining the volume and weight of 

n=5 replicate samples. The relative percent standard deviation 

(RPSD) of this analysis was 1.0 percent. Sediment percent dry 

weight was then determined on duplicate 3-g subsamples (wet 

weight) by weighing before and after oven drying at 105 °C 

until constant weight was achieved. The relative percent difference (RPD) of this analysis was 1.7 percent. Subsequently, 

the two samples were then combusted at 450 °C for 4 hours 

and reweighed for a percent LOI determination. The RPD of 

this analysis was 2.8 percent.

Total Reduced Sulfur

Sediment total reduced sulfur (TRS) analysis was carried 

out on the 2011 whole sediment sample using a modified 

version of a single-step distillation procedure using heated 

acid and chromium for reduction of sulfur to hydrogen sulfide 

(H2S) (Fossing and Jørgensen, 1989), where the evolved 

H2S gas is precipitated as zinc sulfide (ZnS) in a zinc acetate 

solution. The resulting ZnS precipitate was subsequently 

subsampled and assayed using spectrophotometric analysis 

(Cline, 1969). Duplicate subsamples of 1.0 g (wet weight) 

were analyzed. Calibration standards (ZnS) and matrix spike 

solutions were prepared by precipitating a known weight 

reagent grade sodium sulfide (Na2S) crystal in 10 percent zincacetate in anoxic conditions. The ZnS matrix spike recovery 
Results of the Laboratory and Field Experiments 11

was 90 percent (n=1). Certified reference material for the TRS 

assay is not available commercially. The RPD for duplicate 

analysis was 3.5 percent.

Iron Speciation

Sediment iron speciation was carried out both on the 

2011 whole sediment sample and the 2013 freeze dried field 

and mesocosm samples. Three Fe-fractions (acid-extractable 

ferrous iron (Fe(II)AE), amorphous (poorly crystalline) ferric 

iron (Fe(III)a), and crystalline ferric iron (Fe(III)c) were 

assayed as described in Marvin-DiPasquale and others 

(2009a,b). Sediment subsample masses used for analysis were 

approximately 0.5 g wet weight for the 2011 whole sediment 

sample (assayed in duplicate) and 0.07–0.21 g dry weight for 

the 2013 freeze dried sediment samples. The typical detection 

limit for each Fe-fraction is approximately 0.02 μg/mL at the 

level of the spectrophotometric analysis. Certified reference 

material for the various Fe-species is not commercially 

available. Calibration standards of FeSO4 were prepared in 

a solution of 0.25 M hydroxylamine-HCl to inhibit Fe(II) 

oxidation to Fe(III). The average RPD for duplicate assays 

associated with each of the three Fe fractions in sediment was 

15.1±7.7 percent (n=3) for Fe(II)AE, 8.2±1.3 percent (n=2) for 

Fe(III)a, and 2.1±0.8 percent (n=2) for Fe(III)c.

Laboratory Analytical Procedures 

for Pore Water Collected From 

Experiments 1, 2, 4, and 5 and Sediment 

From Experiments 4 and 5

Analysis of pore water for total mercury was carried 

out with a Tekran 2600 cold-vapor atomic fluorescence 

spectrometer (CVAFS) following EPA method 1631 and 

using hydroxylamine hydrochloride (NH2OH–HCl) for 

oxidant pre-reduction and acidified SnCl2 for instrument 

reduction (U.S. Environmental Protection Agency, 2002a). 

All pore-water mercury samples were analyzed in clean room 

conditions using trace metal grade reagents and following 

appropriate and well-documented sample handling protocols 

(Mason and others, 1998; U.S. Environmental Protection 

Agency, 2002a).

Sediment samples from Experiments 4 and 5 were 

analyzed for THg using a Direct Mercury Analyzer (DMA) 

using thermal decomposition, amalgamation, and atomic 

absorption spectrophotometry following EPA Method 7473 

(U.S. Environmental Protection Agency, 2007).

MeHg analysis for sediment and pore water was carried 

out with a Tekran 2700 CVAFS following EPA Method 

1630 (U.S. Environmental Protection Agency, 1998) with 

modifications for sediment digestion methods (personal 

commun., Evangelos Kakouros, U.S. Geological Survey, 

Menlo Park, Calif.). Sample preparation involved the 

following processes: distillation (for pore waters), digestion 

(for sediment) ethylation, chromatographic separation, and 

thermal decomposition.

Sulfide was determined colorimetrically following Cline 

(1969) (detection limit 10 micromoles per liter (μmol/L). 

Fe(II) was determined colorimetrically following Stookey 

(1970) (detection limit 5 μmol/L). pH was measured with 

a portable gel probe (Accumet Gel-Filled Combination 

Electrode with an Orion Model 290Aplus meter). DOC was 

determined using a TOC analyzer (OI Corporation 1010; 

detection limit 0.5 parts per million [ppm]).

Results of the Laboratory and Field 

Experiments

The results of the laboratory and field experiments 

are described in 6 sections. The first section describes the 

chemical characteristics of the littoral sediment and pore 

water. The second and third sections describe the effects of 

ZVI on pore water chemistry from littoral and profundal 

sediment. The fourth section describes the effects of ZVI on 

pore sediment chemistry and the uptake of mercury in snails 

from Experiment 3. The fifth section describes the effects of 

ZVI and GAC on pore-water and sediment chemistry in the 

field mesocosms. The final section describes the effects of ZVI 

and GAC on sediment chemistry and uptake of mercury in 

snails in Experiment 5.

Chemical Characteristics of Littoral Sediment 

and Sediment Pore Water

Selected chemical characteristics of littoral sediment 

collected from the top 15 cm in a wetland fringing the eastern 

shore of Hodgdon Pond indicate this material to be rich in 

organic matter (LOI = 46.5 percent; table 1). The sediment 

had 8.36 μmol of total reduced sulfur gdw-1 and 2.0 mg gdw-1

total Fe in the Fe(II)AE, Fe(III)a and Fe(II)c (crystalline) 

fraction combined, most of which was Fe(II)AE with minor 

amounts in the other forms. The sediment contained 90.5 ng 

gdw-1 THg, 1.36 ng gdw-1 Hg(II)R, and 2.42 ng gdw-1 MeHg. 

Reactive mercury was 1.5 percent of THg and MeHg was 

2.68 percent of THg. This sediment THg concentration was 

substantially higher than that reported for four vernal pools 

at Acadia National Park (Loftin and others, 2012) but the 

MeHg concentrations are similar to those observed in a vernal 

pool in Massachusetts (Benoit and others, 2013). The pore 

water extracted from this sediment contained 49.3 ng/L THg 

and 0.76 ng/L MeHg and MeHg was 1.54 percent of THg in 

pore water.
12 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

Table 1. Chemical characteristics of littoral sediment and pore 

water extracted from sediment from wetlands fringing Hodgdon 

Pond, Mount Desert Island, Maine. 

[%, percent; wt, weight; μmol, micromole; μm, micrometer; gdw-1, per gram 

dry weight of sediment; S, sulfur; mg, milligram; ng, nanogram; ng L-1, nanogram per liter; Fe(II)AE, acid-extractable divalent iron or Fe2+; Fe(III), trivalent 

iron or Fe3+; Hg, mercury; Hg(II)R, reactive Hg; THg, total Hg; MeHg, methyl 

mercury; <, less than]

Constituent Units Value

Sediment loss on ignition (LOI) as 

percentage of dry wt

(%) 46.5

Sediment grain size <64 μm (%) 63.4

Sediment total reduced S (TRS) (μmol gdw-1) 8.36

Sediment acid-extractable Fe(II)AE (mg gdw-1) 1.84

Sediment amorphous Fe(III)a (mg gdw-1) <0.08

Sediment crystalline Fe(III)c (mg gdw-1) 0.09

Sediment THg (ng gdw-1) 90.50

Sediment Hg(II)R (ng gdw-1) 1.36

Sediment MeHg (ng gdw-1) 2.42

Sediment Hg(II)R as a percentage of THg (%) 1.50

Sediment MeHg as a percentage of THg (%) 2.68

Pore-water MeHg (ng L-1) 0.76

Pore-water THg (ng L-1) 49.30

Pore-water MeHg as a percentage of THg (%) 1.54

Effects of Zero-Valent Iron on Pore-Water 

Chemistry From Littoral Sediment (Experiment 1)

The addition of ZVI to blended sediment maintained in 

microcosms in the laboratory resulted in an increase in Fe(II) 

concentrations in the 50 times native total Fe treatments 

and an increase in pH in the 5 and 50 times native total iron 

treatments in the surface sediment where treatments had been 

applied (figs. 2A and B; table 2). No statistical comparison 

is warranted because of the small sample size; therefore 

these increases should be considered qualitative indications 

of changes in pore-water chemistry. These differences in 

Fe(II) concentrations and pH were observed at 2- and 4-cm 

depths but not at 6- and 8-cm depths (figs. 2C and D). Porewater THg and MeHg concentrations were below detection 

limits (approximately 0.10 ng/L for THg and 0.4 picograms 

[absolute] for MeHg) at all depths and for all sampling times. 

Dissolved hydrogen sulfide concentrations were measured 

before ZVI treatment and generally determined to be in the 

range of 0.5 to 4 μmol/L. Initial DOC concentrations 12 days 

after the microcosms were established were in the range of 

27 to 52 ppm but by 42 days after establishment had declined 

to 20 and 28 ppm (table 2). Following ZVI application, DOC 

concentrations changed little except at the 2- and 4-cm depths 

in the 50 times native Fe treatment where they declined to the 

range of 15 to 20 ppm indicating possible adsorption to Fe 

in these treatments. Substantial seasonal variation in Fe(II) 

and sulfide concentrations in pore water has been reported 

for organic matter-rich littoral lake sediment in Michigan 

(Koretsky and others, 2006). These seasonal variations suggest 

that temperature and biological activity substantially affect 

concentrations so that the concentrations measured in these 

unvegetated microcosms in the laboratory are not necessarily 

representative of field conditions.

Effects of Zero-Valent Iron on Pore-Water 

Chemistry From Profundal Sediment 

(Experiment 2)

In this experiment, the concentration of Fe(II) varied 

from 0.16 to 2.0 μmol/L in the overlying water before 

application of the ZVI treatment and Fe(II) varied from 

4 to 27 μmol/L in the pore water from 1 to 7 cm below the 

sediment water interface (table 3). pH in the pore water varied 

from 5.1 to 6.1, increasing with depth below the surface 

the sediment water interface. No consistent change in pH 

was observed 10 days after ZVI application. Pore-water 

THg and MeHg were not detected, were below detection 

limit, or were only slightly above detection limits before 

and after ZVI treatment applications. In this experiment a 

different type of pore-water sampler (draw-peeper) was used 

compared with Experiment 1 where Rhizon samplers were 

used and differences in sampler types could affect the pore 

water sampled. Similarly, it was not possible to ensure that 

the samplers were always sampling from the same depths 

below the sediment water interface because of variability in 

the surface topography and sediment density among replicate 

microcosms. Additionally, the sediment water interface 

was observed to migrate upwards in some instances as the 

surface sediment became less dense and more of a floc. These 

differences indicate the need for caution in comparing porewater chemistry between experiments and even throughout 

time in any single experiment.

Effects of Zero-Valent Iron on Sediment 

Chemistry and Uptake of Mercury in Snails 

Using Vegetated Littoral Sediment 

(Experiment 3)

The application of ZVI to vegetated littoral sediment 

from wetlands fringing Hodgdon Pond maintained in glass 

aquaria resulted in an increase in surficial (top 3 cm) sediment 

MeHg and no change in snail MeHg (table 4). For the 5 times 

native Fe treatment the concentration of Fe(II)AE in the surficial sediment approximately doubled and the concentrations 

of Fe(III)a, Fe(III)c, and total Fe increased by factors of 3.7, 

15, and 5.0, respectively. These differences were statistically 
Results of the Laboratory and Field Experiments 13

0

50

100

150

200

250

Pretreatment Posttreatment

Concentrations of Fe(II), in micromoles

Control

5 times Fe

50 times Fe

5.2

5.4

5.6

5.8

6.0

6.2

6.4

6.6

pH

0

50

100

150

200

250

300

2

4

6

8

5.0

5.2

5.4

5.6

5.8

6.0

6.2

6.4

6.6

pH

EXPLANATION

A. B.

C. D.

Pretreatment Posttreatment

Pretreatment Posttreatment Pretreatment Posttreatment

Concentrations of Fe(II), in micromoles

EXPLANATION

Iron treatments

Control

5 times Fe

50 times Fe

EXPLANATION

Iron treatments

Depth below sediment-water

interface, in centimeters

2 

4 

6

8

EXPLANATION

Depth below sediment-water

interface, in centimeters

Figure 2. Pore water concentration of A, ferrous iron for 5 and 50 times iron treatments, B, pH for 5 and 50 times iron treatments, 

C, ferrous iron collected from 2, 4, 6, and 8 centimeters depth below the sediment-water interface, and D, pH collected from 2, 4, 6, and 

8 centimeters depth below the sediment-water interface. Fe(II), divalent iron.
14 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

Table 2. Pore-water chemistry from laboratory microcosms with blended littoral sediment (Experiment 1) from Mount Desert Island, 

Maine.

[Each value is the average of two samples. If the value was less than the minimum detection limit (MDL), then one-half the MDL was used to compute the average; MDL varied among analysis dates. Fe(II), divalent iron; DOC, dissolved organic carbon; cm, centimeter; μM, micromolar; ppm, parts per million; Fe, iron; 

NA, not analyzed]

Treatment Depth

cm

Fe(II)

μmol/L

Sulfide

μmol/L

DOC

ppm pH

 2/14/2012 12 days postestablishment

MDL 1.29 0.59 1.11

Reference 2 41.38 1.91 39.37 5.42

Reference 4 66.27 2.89 36.67 5.63

Reference 6 115.70 2.66 43.58 5.51

Reference 8 138.60 1.39 48.02 5.42

5 times Fe 2 54.99 2.71 32.01 5.45

5 times Fe 4 65.46 0.39 44.38 5.28

5 times Fe 6 72.59 2.23 47.70 5.45

5 times Fe 8 90.94 0.08 52.30 5.35

50 times Fe 2 65.56 3.79 30.48 5.52

50 times Fe 4 89.67 2.79 27.13 5.46

50 times Fe 6 122.09 0.84 39.19 5.39

50 times Fe 8 123.75 0.30 45.52 5.35

3/15/2012 42 days postestablishment

MDL 0.50 0.59 1.11

Reference 2 60.72 NA 27.10 5.69

Reference 4 66.38 NA 23.19 5.84

Reference 6 75.89 NA 23.17 5.77

Reference 8 79.65 NA 27.95 5.89

5 times Fe 2 66.39 NA 22.80 5.95

5 times Fe 4 67.22 NA 19.50 5.97

5 times Fe 6 81.72 NA 23.23 6.17

5 times Fe 8 81.69 NA 23.63 5.85

50 times Fe 2 45.29 NA 26.60 5.99

50 times Fe 4 58.52 NA 24.23 5.78

50 times Fe 6 77.68 NA 23.16 5.80

50 times Fe 8 82.52 NA 24.17 5.76

3/26/2012 53 days postestablishment

MDL 1.76 0.10 1.11

Reference 2 62.16 0.45 23.62 5.76

Reference 4 66.63 2.67 21.82 5.75

Reference 6 69.89 0.39 21.82 5.64

Reference 8 79.73 0.51 24.68 5.63

5 times Fe 2 66.59 1.62 20.29 5.69

5 times Fe 4 64.07 2.38 21.16 5.81

5 times Fe 6 79.21 4.08 27.55 5.74

5 times Fe 8 82.39 4.09 24.47 5.58

50 times Fe 2 61.42 1.74 20.28 5.69

50 times Fe 4 50.97 2.83 20.51 5.60

50 times Fe 6 65.87 1.75 19.84 5.56

50 times Fe 8 79.12 1.64 22.53 5.66

Treatment Depth

cm

Fe(II)

μmol/L

Sulfide

μmol/L

DOC

ppm pH

4/17/2012 22 days posttreatment

MDL 0.30 0.10 1.11

Reference 1.50 5.68 NA 20.11 5.55

Reference 2 39.83 NA 23.47 5.79

Reference 4 60.35 NA 31.09 5.77

Reference 6 67.58 NA 26.68 5.75

Reference 8 67.85 NA 26.57 5.68

5 times Fe 1.50 13.28 NA 15.42 6.02

5 times Fe 2 71.77 NA 20.37 5.94

5 times Fe 4 91.45 NA 25.72 5.86

5 times Fe 6 66.91 NA 28.26 5.71

5 times Fe 8 82.95 NA 28.83 5.60

50 times Fe 1.50 25.27 NA 14.89 6.62

50 times Fe 2 263.63 NA 16.18 6.39

50 times Fe 4 234.56 NA 19.04 6.42

50 times Fe 6 84.10 NA 27.28 5.92

50 times Fe 8 66.16 NA 25.12 5.78

5/7/2012 42 days posttreatment

MDL 0.30 0.10 1.11

Reference 1.50 2.75 NA 18.43 5.57

Reference 2 30.03 NA 20.29 5.62

Reference 4 57.01 NA 24.10 5.68

Reference 6 64.19 NA 25.85 5.59

Reference 8 72.65 NA 26.73 5.60

5 times Fe 1.50 3.69 NA 14.42 5.93

5 times Fe 2 78.19 NA 17.21 6.10

5 times Fe 4 88.34 NA 22.32 5.80

5 times Fe 6 64.35 NA 25.11 5.82

5 times Fe 8 81.43 NA 29.36 5.61

50 times Fe 1.50 6.02 NA 10.75 6.49

50 times Fe 2 233.46 NA 14.09 6.41

50 times Fe 4 270.00 NA 15.92 6.31

50 times Fe 6 83.83 NA 23.75 5.83

50 times Fe 8 65.63 NA 27.08 5.72
Results of the Laboratory and Field Experiments 15

Table 3. Pore-water chemistry from laboratory microcosms with profundal sediment (Experiment 2) from Mount Desert Island, Maine. 

[Each value is the average of two samples. If the value was less than the minimum detection limit (MDL), then one-half the MDL was used to compute the average; MDL varied among analysis dates. Fe(II), divalent iron; THg, total mercury; MeHg, methyl mercury; cm, centimeter; μM, micromolar; ng L-1, nanograms 

per liter; Fe, iron; NA, not analyzed]

Treatment Depth

(cm)

Fe(II)

μmol/L

THg

(ng/L)

MeHg

(ng/L) pH

7/2/2012 pretreatment

MDL 0.16 1.0 0.45

Reference 1.5 0.19 NA NA 4.60

Reference -1 12.65 2.13 0.37 5.33

Reference -3 10.97 4.72 NA 5.57

Reference -5 15.25 NA NA 5.85

Reference -7 14.16 9.55 NA 5.77

5 times Fe 1.5 1.08 0.50 NA 4.26

5 times Fe -1 24.09 1.37 NA 5.69

5 times Fe -3 5.68 NA NA 5.68

5 times Fe -5 8.77 5.09 NA 5.78

5 times Fe -7 9.89 NA NA 5.71

50 times Fe 1.5 0.22 0.50 NA 4.56

50 times Fe -1 8.07 2.62 0.55 5.49

50 times Fe -3 8.79 6.12 NA 5.77

50 times Fe -5 8.52 5.60 NA 5.77

50 times Fe -7 11.60 5.18 NA 5.73

7/12/2012 10 days posttreatment

Reference 1.5 NA NA NA 4.45

Reference -1 NA 0.94 0.23 5.46

Reference -3 NA NA NA 5.73

Reference -5 NA NA NA 5.74

Reference -7 NA NA NA 5.83

5 times Fe 1.5 NA NA NA 4.73

5 times Fe -1 NA 1.20 0.23 5.47

5 times Fe -3 NA NA NA 5.86

5 times Fe -5 NA NA NA 5.95

5 times Fe -7 NA NA NA 5.81

50 times Fe 1.5 NA NA NA 4.58

50 times Fe -1 NA 0.50 0.23 5.63

50 times Fe -3 NA NA NA 5.88

50 times Fe -5 NA NA NA 5.81

50 times Fe -7 NA NA NA 5.85

Treatment Depth

(cm)

Fe(II)

μmol/L

THg

(ng/L)

MeHg

(ng/L) pH

11/12/2012 133 days posttreatment

Reference 1.5 NA NA NA NA

Reference -1 NA NA 0.08 NA

Reference -3 NA 1.25 1.12 NA

Reference -5 NA 2.01 NA NA

Reference -7 NA NA NA NA

5 times Fe 1.5 NA NA NA NA

5 times Fe -1 NA NA 0.08 NA

5 times Fe -3 NA 1.00 0.20 NA

5 times Fe -5 NA 1.47 NA NA

5 times Fe -7 NA 2.96 NA NA

50 times Fe 1.5 NA NA NA NA

50 times Fe -1 NA NA 0.07 NA

50 times Fe -3 NA 1.81 0.28 NA

50 times Fe -5 NA 1.98 NA NA

50 times Fe -7 NA 2.21 NA NA
16 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine Table 4. Summary of sediment and snail chemistry from Experiment 3 vegetated mesocosms from Mount Desert Island, Maine, with great pond snails. [The mean values and standard errors of the means are shown where there were four replicate analyses in all cases. Values of differences in bold indicate statistically significant differences at 

p less than (<) 

0.05, values of differences in italics indicate statistically significant differences at 

p<0.1. MeHg, methyl mercury; Sed., sediment; Fe(II)AE, acid-extractable divalent iron; Fe(III)a, trivalent amorphous iron; 

Fe(III)c, trivalent crystalline iron; Fe total, sum of Fe(II), Fe(III)a, and Fe(II)AE; ng, nanogram; mg, milligram; %, percent; ZVI, zero-valent iron; GAC, granular activated carbon; Std Err; standard error of the mean]

Reference microcosms, experimental run 1

Variable Snail MeHg ng/g Sed. MeHg ng/g Snail MeHg Sed MeHg . Fe(II)AE mg/g Fe(III)a mg/g Fe(III)c mg/g Fe total mg/g Fe(II)

%

Fe II

Fe III a Fe III c

AE ( )

( )

+ ( ) ( )

Fe II

Fe II Fe III a

AE

AE

( )

( )

+ ( ) ( )

Mean 58.8 2.05 30.4 3.62 1.93 1.26 6.82 53.2 1.37 0.64Std Err 6.89 0.23 5.9 0.79 0.37 0.40 0.62 8.99 0.42 0.08

Microcosms treated with ZVI at 5 times native total Fe, Experiment 3, run 1

Mean 65.7 3.90 16.5 8.63 7.23 18.3 34.2 25.2 0.35 0.53Std Err 13.7 0.39 8.4 1.52 0.80 2.80 4.36 3.91 0.07 0.06

Net change (ZVI – Reference), Experiment 3, run 1

Mean 6.89 1.85 -13.9 5.01 5.30 17.0 27.3 -28.0 -1.02 -0.11Std Err 7.68 0.23 5.13 0.86 0.44 1.41 2.20 4.90 0.21 0.05

Reference microcosms, Experiment 3, run 2

Mean 55.6 1.26 39.9 2.34 2.01 1.12 5.47 43.1 0.76 0.56Std Err 22.5 0.21 8.4 0.30 0.58 0.08 0.82 1.38 0.04 0.04

Microcosms treated with ZVI at 10 times native total Fe, Experiment 3, run 2

Mean 51.1 3.11 18.1 12.7 7.75 32.3 52.7 32.6 0.55 0.60Std Err 4.35 0.65 6.6 3.16 2.06 21.5 25.2 7.90 0.18 0.07

Net change (ZVI – Reference), Experiment 3, run 2

Mean -4.48 1.85 -21.8 10.4 5.74 31.1 47.3 -10.5 -0.21 0.04Std Err 11.4 0.34 5.31 1.59 1.07 10.7 12.6 4.01 0.09 0.04
Results of the Laboratory and Field Experiments 17

significant (p less than [<] 0.01). Fe(II)AE as a proportion of 

total Fe increased from 25 to 53 percent in the ZVI-treated 

sediment compared with the reference treatments. Concentrations of Fe(II)AE, Fe(III)c, and total Fe increased substantially 

more for the 10 times native Fe treatment by factors of 5, 

9.6, and 29, but Fe(III)a increased by a very similar factor 

to what was observed in the 5 times native Fe treatment. For 

the 10 times native Fe treatment the amount of Fe(II)AE as a 

proportion of total Fe decreased from 43 to 33 percent in the 

ZVI-treated sediment compared with the reference treatments.

In this experiment with vegetated littoral sediment, 

when all treatments were combined, the sediment MeHg 

concentration correlated positively with the Fe(III)

a concentration (r squared [R2

]=0.76) and the regression 

between these two variables was statistically significant 

(p<0.001) (fig. 3A). Similarly, sediment MeHg concentration 

was positively correlated with Fe(II)AE concentration 

(R2

=0.39) and the regression between these two variables 

also was statistically significant (p<0.01) (fig. 3B). The 

MeHg concentration in snail tissue was positively related to 

the amount of sediment MeHg but the regression was not 

statistically significant. Snail tissue MeHg was not related to 

the concentration of any iron species.

The results from the two separate experimental runs 

with 5 times total Fe in the first run and 10 times total Fe in 

the second run (table 4) were similar in that the concentration 

of MeHg in snail tissue was unaffected by the ZVI treatment 

regardless of the amount of ZVI added. In both cases, the 

MeHg concentration in the sediment was increased compared 

with the reference treatment.

Effects of Zero-Valent Iron and Granular 

Activated Carbon Treatments on Pore-Water 

and Sediment Chemistry in Field Mesocosms 

(Experiment 4)

There were no statistically significant (p<0.05) 

differences in the concentrations of MeHg, Fe(II) or dissolved 

hydrogen sulfide in pore water between the treated mesocosms 

and the reference mesocosms in Experiment 4 at 25 or 

91 days after treatment application for any of the depths 

sampled (fig. 4). However, the concentration of MeHg in 

pore water on September 6, 2013 (25 days after application 

of treatments) was lower, but not significantly, in the GACtreated mesocosms than in the reference mesocosms for the 

average concentration of MeHg in samples collected at the 

2 and 4 cm depths (0.1<p<0.12) (fig. 5). No statistically 

significant differences (p<0.05) were observed in porewater concentrations of MeHg, hydrogen sulfide, or Fe(II) 

between samplings 25 and 91 days after application of ZVI 

and GAC treatments, but the reference mesocosms indicated 

a significant decrease in pore-water MeHg between the two 

sampling dates (fig. 4).

y = 0.3138x + 1.0924

R2 = 0.7563

0

1

2

3

4

5

6

0 5 10 15

Concentrations of sediment MeHg,

in nanograms per gram dry weight

Concentrations of Fe(III)a, in milligrams per gram dry weight

ZVI

Reference

A

y = 0.1492x + 1.5586

R2 = 0.3946

0

1

2

3

4

5

6

0 5 10 15 20

Concentrations of sediment MeHg,

in nanograms per gram dry weight

Concentrations of Fe(II)AE, in milligrams per gram dry weight

B

EXPLANATION

ZVI

Reference

EXPLANATION

Linear regression

trend line

Linear regression

trend line

Figure 3. The relation between concentrations of A, sediment 

methylmercury (MeHg) and amorphous trivalent iron (Fe(III)a) and 

B, sediment MeHg and acid-extractable divalent iron (Fe(II)AE), 

Experiment 3 for Mount Desert Island, Maine, for zero-valent iron 

(ZVI) treatment and untreated (reference) samples. R2

, r squared; 

mg/g dwt, milligram per gram dry weight.

No statistically significant differences were observed in 

sediment THg, MeHg, or MeHg:THg ratio among mesocosms 

on August 12, 2013, before the application of ZVI and GAC 

treatments (data not shown). No significant differences were 

observed in sediment THg, MeHg, or MeHg:THg ratio 

between treatments or in comparison with the pretreatment 

condition when mesocosms were resampled on September 6, 

2013 (fig. 6). When mesocosms were resampled on 

November 11, 2013 (91 days posttreatment) THg was 

unchanged but MeHg and MeHg:THg ratio decreased 

significantly (p<0.05) on both the ZVI-treated and reference 

plots. There was a trend towards decreasing MeHg and 

MeHg:THg ratio on the GAC-treated plots but the decreases 

were not significant (0.1<p<0.12).
18 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

0

20

40

60

80

100

120

Concentrations of Fe(III), in micromoles per liter

0

10

20

30

40

50

60

70

09/06/2013 11/11/2013

 Fe(II)

0

2

4

6

8

10

12

A. Reference 

2

4

6

8

0

2

4

6

8

10

12

Concentrations of MeHg, in nanograms per liter and S(II), in micromoles per liter

B. ZVI

0

5

10

15

20

09/06/2013 11/11/2013 09/06/2013 11/06/2013

C. GAC

Standard error of the mean

Depth below sediment-water

interface, in centimeters

 Me Hg S(II)

0

20

40

60

80

100

120

140

160

180

EXPLANATION

Figure 4. Concentrations of methylmercury (MeHg), sulfide (S(II)), and divalent iron (Fe(II)) in pore water from field mesocosms 

following A, untreated (reference; Experiment 4), and treatments of B, zero-valent iron (ZVI) treatment and C, granular activated carbon 

(GAC), sampled at Mount Desert Island, Maine, September 6 and November 11, 2013, at four depths below the sediment-water interface.
Discussion of Experimental Results 19

0.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

09/06/2013 11/11/2013

Concentrations of MeHg,

in nanograms per liter

2 to 4

6 to 8

+ +

EXPLANATION

p-value is < 0.12 for a paired 2-tailed

Students t-Test for the difference

between the treatment and the reference.

+

Standard error of the mean

Refence ZVI GAC

09/06/2013 11/11/2013 09/06/2013 11/11/2013

Depth below sediment-water

interface, in centimeters

Figure 5. Concentrations of methylmercury (MeHg) in pore water collected from field mesocosms from Mount 

Desert Island, Maine, on September 6 and November 11, 2013, for Experiment 4. Samples included untreated 

(reference) and treated zero-valent iron (ZVI) and granular activated carbon (GAC), at 2 to 4 and 6 to 8 cm depth 

below the sediment-water interface.

The distribution coefficient for MeHg (MeHg KD) 

increased (p<0.05) in sediment at 6 cm depth below the sediment water interface from August 12, 2013, to September 6, 

2013, following application of GAC (fig. 7). The KD decreased 

in the sediment at the 4 cm depth (p<0.1) following application of ZVI from August 12 to September 6, 2013 (fig. 7).

Effects of Zero-Valent Iron and Granular 

Activated Carbon on Sediment Chemistry and 

Uptake of Mercury in Snails Using Vegetated 

Littoral Sediment (Experiment 5)

Analysis of MeHg in snail tissue from snails of the same 

cohort that were used in this experiment indicated that snails 

accumulated a substantial amount of MeHg in reference, ZVItreated, and GAC-treated microcosms. Before the exposure 

to treatments, the snail tissue contained 1.2 ng MeHg/g dwt 

MeHg, and after exposure the tissue contained 103, 41.8, and 

64.5 ng MeHg/g dwt in the reference, ZVI-treated and GACtreated microcosms, respectively. The application of ZVI to 

vegetated littoral sediment from wetlands fringing Hodgdon 

Pond maintained in glass aquaria in this second experiment 

using snails resulted in a significant decrease in MeHg concentration in snail tissue compared with the reference treatments 

(table 5). The GAC application appeared to reduce the concentration of MeHg in snail tissue compared with the reference 

treatments but the decrease was not statistically significant 

(p=0.17). The application of ZVI did not affect the concentration of MeHg in the sediment, but the application of GAC 

increased the concentration of MeHg in the sediment and 

lowered the ratio of MeHg in snail tissue to MeHg in the sediment (table 5). There were few differences in the Fe fractions 

in this experiment. The concentration of Fe(II)AE was lower in 

the GAC treatments compared with the reference treatments 

and ZVI treatments (table 5). The ratio of Fe(II)AE to the sum 

of Fe(III)a and Fe(III)c was lower in the ZVI treatments compared with the reference and GAC treatments.

The concentration of MeHg in surface sediment tended 

to be lower in the ZVI-treated mesocosms and higher in the 

GAC-treated mesocosms than in the reference treatments 125 

and 154 days following treatment applications (fig. 8) but 

these differences were not statistically significant (p<0.05). 

The p-value for the difference between the reference and 

ZVI-treated mesocosms for 154 days post treatment was 0.17. 

The concentration of MeHg in pore water tended to be lower 

in the ZVI-treated mesocosms than in the reference treatments 

125 and 154 days following treatment applications (fig. 8) but 

these differences were not statistically significant (p<0.05).

Discussion of Experimental Results

The laboratory experiments where ZVI was applied to 

blended littoral and profundal sediment in columns indicated 

that the concentration of Fe(II) increased following application of ZVI. The increase in Fe(II) was expected based on the 

anticipation of two-electron oxidation of elemental Fe to Fe(II) 

and its subsequent one-electron oxidation to poorly crystalline 

Fe(III)a. Following the application of ZVI, pH increased from 

about 5.6 to 6.4. The concentration of MeHg in pore water was 

at or below detection in these conditions.

At the higher ZVI addition treatment, and in the 2- and 

4-cm depths, pore-water DOC concentration appeared to 

decrease, possibly a result of DOC sorption to solid-phase 

Fe(III) surfaces. An increase in pH would normally be 

expected to result in an increase in DOC as has been observed 
20 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

0

1

2

3

4

5

6

7

8

0

20

40

60

80

100

120

140

Pretreatment 25 days posttreatment 91 days posttreatment

MeHg (ng/g) and MeHg:THg (%)

Total Hg (ng/g)

0

1

2

3

4

5

6

7

8

0

20

40

60

80

100

120

140

Pretreatment 25 days posttreatment 91 days posttreatment

Pretreatment 25 days posttreatment 91 days posttreatment

MeHg (ng/g) and MeHg:THg (%)

Total Hg (ng/g)

THg

MeHg

MeHg:THg

** **

0

1

2

3

4

5

6

7

0

20

40

60

80

100

120

MeHg (ng/g) and MeHg:THg (%)

Total Hg (ng/g)

 

** 

A. Reference 

B. ZVI

C. GAC

Standard error of the mean

p-value of <0.01 for a paired

2-tailed Students t-Test

for the difference between

the treatment and the reference

EXPLANATION

Figure 6. Concentrations of total mercury (THg), methylmercury (MeHg), and MeHg:THg for sediment 

sampled from field mesocosms (Experiment 4) from Mount Desert Island, Maine. Samples were collected 

at 0 to 3 cm depth below the sediment-water interface for A, untreated (reference), and B, treated with 

zero-valent iron (ZVI), and C, treated with granular activated carbon (GAC) at pretreatment and 25 and 

91 days posttreatment. ng/g, nanograms per gram; %, percent; THg, total mercury; MeHg:THg, ratio of 

methylmercury to total mercury. 
Discussion of Experimental Results 21

1

10

100

Reference ZVI GAC

A. 9/6/2013 (25 days posttreatment)

2

4

6

8

*

* **

1

10

100

Reference ZVI GAC

B. 11/11/2013 (91 days posttreatment)

*

Log sediment

MeHg/pore water MeHg (KD)

Log sediment

MeHg/pore water MeHg (KD)

Depth below sediment-water

interface, in centimeters

EXPLANATION

*

**

p-value of <0.05 for a paired 2-tailed

Students t-Test for the difference

between the treatment and the reference

p-value of <0.01 for a paired 2-tailed

Students t-Test for the difference

between the treatment and the reference

Figure 7. Log of the distribution coefficient (KD) for concentrations of methylmercury (MeHg) between 

sediment and pore water sampled from field mesocosms (Experiment 4) for A, 25 days posttreatment; and 

B, 91 days posttreatment from Mount Desert Island, Maine. Treatments included untreated (reference); 

ZVI, zero-valent iron; and GAC, granular activated carbon.

in other aquatic systems (Monteith and others, 2007). In these 

experiments, without vegetation, there were no new root exudates released to the pore water, so pore-water DOC concentrations may have been more affected than they would have 

been if live vegetation were present. The decrease in DOC 

observed at the higher rate of ZVI addition (table 2) may have 

resulted from an increase in DOC sorption to a substantially 

increased abundance of solid-phase Fe(III) surfaces that have 

a high affinity for binding DOC.

Results of the analysis of Fe speciation in Experiment 3 

indicate that a substantial fraction of the added ZVI (elemental 

Fe0

) was oxidized to Fe(II)AE and subsequently to amorphous 

or poorly crystalline, Fe(III)a and crystalline Fe(III)c. This 

oxidation is reported to be largely abiotic in the presence 

of sufficient oxygen but can occur in anoxic environments 

facilitated by sulfur reducing bacteria or methanogenic 

bacteria (Enning and Garrelfs, 2014; Uchiyama and others, 

2010). The Fe(III)a pool is reported to be the primary substrate 

for microbial reduction of solid-phase Fe(III) to Fe(II) (Roden 

and Wetzl, 2002). These abiotic oxidation reactions and 

microbial reduction were anticipated to create conditions that 

could result in an inhibition of the methylation of mercury as 

discussed in the section on potential remediation mechanisms. 

However, the application of ZVI tended to increase the amount 

of MeHg associated with surficial sediment in one experiment 

(Experiment 3); therefore it is possible that for the conditions 

of Experiment 3 Fe reducing bacteria were stimulated 

to methylate Hg. Apparently, this stimulation of MeHg 

production outweighed any effect of decreasing methylation 

by sulfate reducing bacteria or reduction of availability 

of Hg(II) by sorption to iron oxide surfaces or conversion 

of Hg(II) to Hg0

 and gaseous evasion. These results from 

Experiment 3 contrasted with those of Experiment 5 where 

ZVI did not increase the amount of MeHg associated with 

the sediment and tended to decrease it, though the differences 

were not statistically significant.

There were no significant differences in MeHg uptake 

in snails in Experiment 3, comparing the effects of ZVI 

additions to reference treatments. However, in Experiment 5 

in similar conditions to those in Experiment 3, ZVI treatments 
22 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine Table 5. Summary of sediment and great pond snail chemistry from Experiment 5 for Mount Desert Island, Maine. [The mean values and standard errors of the means are shown where there were four replicate analyses in all cases except for the sediment values for the reference treatment where there were three replicates. Values of differences in bold were indicated statistically significant differences at 

p less than (<) 0.05, values of differences in italics were indicated statistically significant differences at 

p<0.1. MeHg, methyl 

mercury; Sed., sediment; Fe(II)AE, acid-extractable divalent iron; Fe(III)a, trivalent amorphous iron; Fe(III)c, trivalent crystalline iron; Fe Total, sum of Fe(II)AE, Fe(III)a, and Fe(III)c; ng, nanogram; mg, milligram; %, percent; ND, no data; ZVI, zero valent iron; GAC, granular activated carbon; Std Err; standard error of the mean]

Reference microcosms

Variable Snail MeHg ng/g Sed. MeHg ng/g Snail MeHg Sed MeHg . Fe(II)AE mg/g Fe(III)a mg/g Fe(III)c mg/g Fe total mg/g Fe(II)AE

%

Fe II

Fe III a Fe III c

AE ( )

( )

+ ( ) ( )

Fe II

Fe II Fe III a

AE

AE

( )

( )

+ ( ) ( )

Mean 103 2.24 57.2 4.77 0.97 1.87 7.61 65.7 2.96 0.83Std Err 22.4 0.33 12.7 0.06 0.20 0.85 1.11 10.4 1.72 0.03

Microcosms treated with ZVI at 5 times native total Fe

Mean 41.8 1.78 25.3 3.84 1.66 4.31 9.81 43.7 0.90 0.70Std Err 6.4 0.19 5.93 0.62 0.29 2.47 2.91 7.58 0.30 0.02

Net change (ZVI – Reference)

Mean -61.3 -0.46 -31.9 -0.93 0.69 2.44 2.20 -22.0 -2.06 -0.13Std Err 11.6 0.19 7.0 0.31 0.18 1.31 1.56 6.42 0.87 0.02

Microcosms treated with GAC

Mean 64.5 6.07 12.7 2.81 0.44 1.12 4.36 64.4 1.89 0.86Std Err 10.3 2.25 7.6 0.20 0.07 0.15 0.23 3.35 0.29 0.02

Net change (GAC – Reference)

Mean -38.6 3.83 -44.5 -1.96 0.53 -0.75 -3.25 -1.3 -1.07 0.03Std Err 12.3 0.57 6.6 0.11 0.11 0.43 0.57 5.45 0.87 0.02
Discussion of Experimental Results 23

0

1

2

3

4

5

6

125 days

posttreatment,

sediment

154 days

posttreatment,

sediment

125 days

posttreatment,

pore water

154 days

posttreatment,

pore water

Concentrations of sediment MeHg, in nanograms per gram

dry weight total of pore water, in nanograms per liter

Reference

ZVI

GAC

Standard error of the mean

EXPLANATION

Figure 8. A comparison of concentrations of methylmercury (MeHg) in sediment (four samples) and pore water (two 

samples) sampled 125 and 154 days posttreatment of zero-valent iron (ZVI) and granular activated carbon (GAC) in 

laboratory mesocosms containing vegetated sediment to untreated (reference) samples, Experiment 5, for Mount Desert 

Island, Maine. ng/g dwt, nanograms per gram dry weight; ng/L, nanograms per liter.

resulted in significantly less uptake of MeHg in snail tissue 

than observed in reference treatments. In Experiment 5, the 

application of GAC did not significantly reduce the uptake 

of MeHg in snail tissue in comparison with the reference 

treatments. We cannot explain the fact that ZVI did not 

reduce MeHg uptake in Experiment 3 but did in Experiment 

5. The vegetated clumps of sediment were collected on 

September 10, 2012, for Experiment 3 and on October 14, 

2013, for Experiment 5. In Experiment 5, there was less total 

sediment mass and substantially more overlying water in 

the microcosms, which may have affected the redox regime 

and therefore methylation. Both experiments were carried 

out in the laboratory under similar light and temperature 

regimes. The snails tended to remain on the surface of the 

sediment, on the vegetation, or on the walls of the aquaria 

near the sediment water interface. Because the snails did not 

seem to ingest live vegetation, it is assumed that during the 

exposure they subsisted on a diet of periphyton and detritus. 

In both experiments there were similar amounts of vegetation, 

and presumably live roots, but the actual amounts were not 

quantified. Both experiments had similar periods of ZVI 

equilibration with surface sediment before adding snails; 

121 days for Experiment 3 and 112 days for Experiment 5. 

One difference between experiments was that in Experiment 

3 the snail exposure period was shorter (28 days) compared to 

the exposure period in Experiment 5 (41 days). It is possible 

that the snails ingested substantially more detritus and 

periphyton in Experiment 5 than in Experiment 3, especially 

so if the snails initially avoided the lower quality food in the 

mesocosms compared with their diet of romaine lettuce before 

the exposure in the aquaria.

The results of the field mesocosm experiment 

(Experiment 4) indicated that there was a decreasing trend 

in pore-water MeHg concentration after application of GAC, 

but because there was no corresponding decrease in sediment 

MeHg concentration, mercury methylation was apparently 

not affected. Rather, the data indicate that the application 

of GAC resulted in the sorption of MeHg to GAC thereby 

increasing the partitioning of MeHg to GAC and decreasing 

MeHg concentration in pore water but not inhibiting mercury 

methylation rate. Sorption of MeHg by GAC has been 

reported in contaminated freshwater and estuarine sediment 

with organic matter contents similar to those in this study 

(Gilmour and others, 2013).

The seasonal decrease in MeHg concentration in sediment from summer to fall in the field mesocosm (Experiment 4) is consistent with reports from other temperate wetlands and estuarine sediment (Heim and others, 2007; Mitchell 

and Gilmour, 2008; Selvendiran and others, 2008; Bradley 

and others, 2011; Bergman and others, 2012). This pattern of 

seasonal decrease in sediment MeHg has been attributed to 

higher activity of sulfate and iron reducing bacteria in spring 

and summer than in fall and winter. Lower microbial activity in the fall is consistent with our observations that DOC 
24 Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine

concentrations were substantially lower in the fall compared 

with in the summer (data not shown). Seasonal variability 

in the activity of the macrophytes in these field mesocosms 

also may have played a role in the observed pattern of MeHg 

concentration in sediment. Macrophytes can affect rhizosphere 

processes by affecting pH, redox potential (Eh) and by the exudation of organic ligands (Cosio and others, 2014) and these 

effects would be more pronounced during periods of active 

growth in the summer than when plants were senescing in the 

fall. Microbial activity also may be decreased in the fall as 

temperatures decline and plants senesce and there is less labile 

dissolved organic matter.

The fact that ZVI did not lower MeHg concentrations in 

pore water compared with the reference treatments could be 

explained by ZVI having a stimulating effect on iron reducing bacteria that methylated Hg. Results from the laboratory 

mesocosm Experiment 5 indicated that in that application of 

ZVI concentrations of MeHg in pore water tended to be lower 

than those measured in the reference treatments. Because 

MeHg concentrations in sediment were also lower in the ZVI 

treatments of Experiment 5 it suggests that the ZVI may have 

reduced methylation rather than acted solely as a sorbent in 

that case. In Experiment 5 the MeHg concentrations in pore 

water were lower in GAC-treated plots compared to reference 

plots after 125 days, consistent with GAC acting as a sorbent, 

but after 154 days MeHg concentrations were higher.

The observed increase in KD for MeHg following the 

application of GAC also indicates that the addition of GAC 

shifted the equilibrium from the dissolved phase towards the 

sorbed phase. The application of ZVI did not increase the 

KD for MeHg and in one case, at the 4 cm sampling depth, 

decreased it, which would be consistent with ZVI increasing 

mercury methylation more than sorption in that experiment.

Limitations of This Study and Potential 

for Future Research

This study was designed to evaluate the potential for the 

amendments ZVI and GAC to decrease mercury methylation, 

pore-water and sediment MeHg concentrations and biological uptake of Hg and MeHg. The study was not designed to 

determine the mechanisms that would explain decreases in 

methylation, MeHg concentrations or uptake. Further research 

with additional measurements could improve understanding 

of the mechanisms behind decreases in pore-water MeHg 

concentrations. For example, monitoring of redox conditions 

using probes to measure oxidation reduction potential as the 

season progressed and water levels and plant activity varied 

during the course of the summer and fall would be particularly helpful in understanding the fate of mercury in these 

systems. Similarly more frequent sampling for monitoring 

mercury and iron species, sulfide, and DOC would support a 

more mechanistic understanding. The field experiment began 

in mid-August with an unusually dry period during which 

total precipitation between August 12 and November 11 was 

19.6 cm compared with the average for that period from 1982 

to 2013, which was 41.9 cm (standard deviation, 13.3 cm; 

National Atmospheric Deposition Program, undated). It would 

be very useful to monitor MeHg in a more typical year where 

precipitation was higher and where the water level in Hodgdon 

Pond did not vary as much as it did in 2013.

The length of the experimental period is of potential 

concern because ideally it should be long enough to observe 

a response in the variables of interest and to evaluate the 

persistence of the effect of the chemical addition. Our 

experiments involving 1 to 3 months of exposure were long 

enough to observe responses in sediment, pore-water, and 

snail tissue chemistry. Results from other studies on the 

biouptake of trace metals in the great pond snail L. stagnalis

(Croteau and Luoma, 2009; Croteau and others, 2011), 

mercury in L. stagnalis (Tessier and others, 2007) or mercury 

in the oligochaete Lumbriculus variegatus (Gilmour and 

others, 2013) have indicated that relatively short term 

experiments of hours to weeks are sufficient to determine if 

the uptake of toxic metals is affected by the conditions of the 

experiment. However, 1 to 3 month periods were not sufficient 

to evaluate if the effects observed would be observed for 

longer periods. Studies in which pore-water and sediment 

chemistry are monitored for several years following additions 

would inform resource managers of the potential persistence 

of these effects. Longer term monitoring also would be helpful 

to track effects on sediment pH, phosphorus availability, and 

potential effects on algal growth. The short term laboratory 

column experiments were carried out in conditions of little 

variation in temperature or water level and therefore may not 

be representative of actual field conditions. The laboratory 

experiments with unvegetated microcosms obviously lacked 

the rhizosphere environments where live macrophytes affect 

microbial activity and pore-water and sediment geochemistry 

and redox conditions (Cosio and others, 2014).

Future experiments could include microbial assays and 

the use of specific inhibitors to reveal existing populations 

and activity of sulfur and iron reducing bacteria. These kinds 

of experiments could help to identify bacteria responsible 

for methylation of mercury in this sediment and how they 

were affected by additions of ZVI and GAC. More frequent 

sampling of MeHg in pore water and sediment would help in 

understanding variations in rates of methylation or demethylation in different experimental conditions. It also could 

be helpful to monitor the uptake of mercury and MeHg in 

periphyton and macrophytes over the course of a summer and 

fall season.

Summary

The U.S. Geological Survey in cooperation with Acadia 

National Park, carried out a series of laboratory and field 

experiments to evaluate the potential of zero-valent iron and 

granular activated carbon to reduce the rate of the bacterially 
References Cited 25

mediated process of mercury methylation and subsequent 

biological uptake in the great pond snail Lymnaea stagnalis. It 

was apparent that the addition of zero-valent iron resulted in 

an increase in divalent iron and especially poorly crystalline 

amorphous trivalent iron. Our original hypothesis that these 

reactions would reduce methylation by binding the substrates 

for bacterial methylation was not consistently supported. 

In our initial column experiments without vegetation, 

the application of zero-valent iron tended to increase the 

amount of methylmercury associated with surficial sediment; 

therefore, it is more likely that iron reducing bacteria were 

stimulated to methylate mercury. Apparently, this stimulation 

of methylmercury production outweighed any effect of 

decreasing methylation by decreasing concentrations of 

the substrates sulfide or divalent mercury through chemical 

binding to iron, sorption to amorphous trivalent iron surfaces 

or evasion of mercury.

The results of the field mesocosm experiments indicated 

that application of granular activated carbon decreased 

pore-water methylmercury concentration, but there was 

no corresponding decrease in sediment methylmercury 

concentration; hence, the sediment-pore water distribution 

coefficient or methylmercury increased. These results 

indicated that mercury methylation was not affected by the 

addition of granular activated carbon. Rather, the data indicate 

that the application of granular activated carbon resulted in 

the sorption of methylmercury to granular activated carbon 

thereby decreasing methylmercury concentration in pore 

water but not inhibiting the mercury methylation rate. The 

application of zero-valent iron did not increase the sedimentpore water distribution coefficient for methylmercury.

The results of our two experiments to test the effect of 

zero-valent iron on the uptake of methylmercury in snails in 

laboratory mesocosms were mixed. The fact that the great 

pond snail L. stagnalis accumulated a substantial amount of 

methylmercury in relatively short exposure periods (41 days 

or less) in laboratory mesocosm experiments with intact vegetated sediment indicates that these snails are useful for evaluating the potential of zero-valent iron and granular activated 

carbon as amendments to reduce biouptake of methylmercury 

in this wetland sediment.

Although zero-valent iron addition did not consistently 

result in a decrease in the methylation of divalent mercury in 

the Hodgdon Pond sediment based on the analysis of methylmercury in sediment, it is possible that it would be effective in estuarine sediment where sulfide concentrations were 

higher than in the Hodgdon Pond freshwater sediment. Further 

research could evaluate the potential for zero-valent iron to 

inhibit methylation in estuarine sediment. The results reported 

here indicate that, in certain conditions, zero-valent iron and 

granular activated carbon may have potential as sorbents to 

reduce the biomagnification of methylmercury in freshwater 

wetland sediment at Acadia National Park. The experiments 

with zero-valent iron indicated that in certain conditions zerovalent iron may inhibit the methylation of mercury and the 

biouptake of methylmercury but the results were inconclusive 

because of differences between experiments. Further research 

could be helpful in understanding the conditions under which 

zero-valent iron can be effective at decreasing mercury 

methylation and the biouptake of methylmercury. Additional 

research also is needed to evaluate the potential detrimental 

effects of zero-valent iron and granular activated carbon on 

biota in this sediment.

References Cited

Acadia National Park, 2010, Mercury contamination at Acadia 

National Park: National Park Service issue brief, June, 2 p., 

accessed September 30, 2014, at http://www.nps.gov/acad/

naturescience/upload/ACAD_Mercury_FactSheet_2010.

pdf.

Amirbahman, Aria, Kent, D.B., Curtis, G.P., and MarvinDipasquale, M.C., 2013, Kinetics of homogeneous and 

surface-catalyzed mercury(II) reduction by iron(II): 

Environmental Science and Technology, v. 47, no. 13, 

p. 7204–7213.

Amirbahman, Aria, Ruck, P.L., Fernandez, I.J., Haines, T.A., 

and Kahl, J.S., 2004, The effect of fire on mercury cycling 

in the soils of forested watersheds—Acadia National 

Park, Maine, USA: Water, Air, and Soil Pollution, v. 152, 

nos. 1–4, p. 315–331.

Bacon, Linda, 2013, Maine lakes transparency, color & 

chemistry—Overall mean values: Maine Department 

of Environmental Protection and Maine Volunteer Lake 

Monitoring Program dataset, accessed March 29, 2014, 

at http://www.gulfofmaine.org/kb/2.0/

record.html?recordid=9678.

Bank, M.S., Burgess, J.R., Evers, D.C., and Loftin, C.S., 

2007a, Mercury contamination of biota from Acadia 

National Park, Maine—A review: Environmental Monitoring and Assessment, v. 126, no. 1–3, p. 105–115.

Bank, M.S., Crocker, Jeff, Connery, Bruce, and Amirbahman, 

Aria, 2007b, Mercury bioaccumulation in green frog (Rana 

clamitans) and bullfrog (Rana catesbeiana) tadpoles from 

Acadia National Park, Maine, USA: Environmental Toxicology and Chemistry, v. 26, no. 1, p. 118–125.

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Prepared by the Pembroke and Rolla Publishing Service 

Centers.

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Huntington and others—Assessment of Sorbent Amendments for Reduction of Mercury Methylation at Acadia National Park, Maine—Scientific Investigations Report 2014–5234

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