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There has been increasing concern about mercury (Hg) levels in marine and freshwater organisms in the Arctic, due to the importance of traditional country foods such as fish and marine mammals to the diet of Northern Peoples. Due to its toxicity and ability to bioaccumulate and biomagnify in food webs, methylmercury (MeHg) is the form of Hg that is of greatest concern. The main sources of MeHg to Arctic aquatic ecosystems, the processes responsible for MeHg formation and degradation in the environment, MeHg bioaccumulation in Arctic biota and the human health implications for Northern Peoples are reviewed here. In Arctic marine ecosystems, Hg(II) methylation in the water column, rather than bottom sediments, is the primary source of MeHg, although a more quantitative understanding of the role of dimethylmercury (DMHg) as a MeHg source is needed. Because MeHg production in marine waters is limited by the availability of Hg(II), predicted increases in Hg(II) concentrations in oceans are likely to result in higher MeHg concentrations and increased exposure to Hg in humans and wildlife. In Arctic freshwaters, MeHg concentrations are a function of two antagonistic processes, net Hg(II) methylation in bottom sediments of ponds and lakes and MeHg photodemethylation in the water column. Hg(II) methylation is controlled by microbial activity and Hg(II) bioavailability, which in turn depend on interacting environmental factors (temperature, redox conditions, organic carbon, and sulfate) that induce nonlinear responses in MeHg production. Methylmercury bioaccumulation-biomagnification in Arctic aquatic food webs is a function of the MeHg reservoir in abiotic compartments, as well as ecological considerations such as food-chain length, growth rates, life-history characteristics, feeding behavior, and trophic interactions. Methylmercury concentrations in Arctic biota have increased significantly since the onset of the industrial age, and in some populations of fish, seabirds, and marine mammals toxicological thresholds are being exceeded. Due to the complex connection between Hg exposure and human health in Northern Peoples-arising from the dual role of country foods as both a potential Hg source and a nutritious, affordable food source with many physical and social health benefits--reductions in anthropogenic Hg emissions are seen as the only viable long-term solution.

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REVIEW

Methylmercury biogeochemistry: a review with special

reference to Arctic aquatic ecosystems

Igor Lehnherr

Abstract: There has been increasing concern about mercury (Hg) levels in marine and freshwater organisms in the Arctic, due to the

importance of traditional country foods such as ﬁsh and marine mammals to the diet of Northern Peoples. Due to its toxicity and

ability to bioaccumulate and biomagnify in food webs, methylmercury (MeHg) is the form of Hg that is of greatest concern. The main

sources of MeHg to Arctic aquatic ecosystems, the processes responsible for MeHg formation and degradation in the environment,

MeHg bioaccumulation in Arctic biota and the human health implications for Northern Peoples are reviewed here. In Arctic marine

ecosystems, Hg(II) methylation in the water column, rather than bottom sediments, is the primary source of MeHg, although a more

quantitative understanding of the role of dimethylmercury (DMHg) as a MeHg source is needed. Because MeHg production in marine

waters is limited by the availability of Hg(II), predicted increases in Hg(II) concentrations in oceans are likely to result in higher MeHg

concentrations and increased exposure to Hg in humans and wildlife. In Arctic freshwaters, MeHg concentrations are a function of two

antagonistic processes, net Hg(II) methylation in bottom sediments of ponds and lakes and MeHg photodemethylation in the water

column. Hg(II) methylation is controlled by microbial activity and Hg(II) bioavailability, which in turn depend on interacting environ-

mental factors (temperature, redox conditions, organic carbon, and sulfate) that induce nonlinear responses in MeHg production.

Methylmercury bioaccumulation–biomagniﬁcation in Arctic aquatic food webs is a function of the MeHg reservoir in abiotic com-

partments, as well as ecological considerations such as food-chain length, growth rates, life-history characteristics, feeding behavior,

and trophic interactions. Methylmercury concentrations in Arctic biota have increased signiﬁcantly since the onset of the industrial

age, and in some populations of ﬁsh, seabirds, and marine mammals toxicological thresholds are being exceeded. Due to the complex

connection between Hg exposure and human health in Northern Peoples—arising from the dual role of country foods as both a

potential Hg source and a nutritious, affordable food source with many physical and social health beneﬁts—-reductions in anthro-

pogenic Hg emissions are seen as the only viable long-term solution.

Key words: mercury, methylation, Arctic, bioaccumulation, freshwater, marine.

Résumé : On observe une préoccupation accrue au sujet des teneurs en mercure (Hg) chez les organismes marins et d'eau douce dans

l'Arctique, dues a

`l'importance de la nourriture traditionnelle du pays comme le poisson et les mammifères marins. À cause de sa

toxicité et de sa capacité d'accumulation et de bioampliﬁcation dans la chaîne alimentaire, le méthyle de mercure (MeHg) constitue

la forme de Hg soulevant les plus grandes préoccupations. L'auteur passe en revue les principales sources de MeHg pour les systèmes

aquatiques de l'Arctique, les procédés responsables de la formation et de la dégradation dans l'environnement, la bioaccumulation du

Hg dans le biote arctique et les implications pour la santé humaine des peuplades nordiques. Dans les écosystèmes arctiques, la

méthylation du Hg(II) dans la colonne d'eau, plutôt que dans sédiments de fond constitue la première source de MeHg, bien qu'une

compréhension plus quantitative du rôle du diméthyle mercure (DMHg) comme source de MeHg s'avère nécessaire. Parce que la

production du MeHg dans les eaux marines est limitée par la disponibilité du Hg(II), la prédiction des augmentations des concentra-

tions du Hg(II) dans les océans conduira vraisemblablement a

`de plus fortes concentrations en MeHg et une exposition accrue des

humains et de la faune. Dans les eaux douces de l'Arctique, les concentrations en MeHg dépendent de deux processus antagonistes,

méthylation nette de Hg(II) dans les sédiments de fond des marais et des lacs, et photodéméthylation du MeHg dans la colonne d'eau.

La méthylation du Hg(II) est contrôlée par l'activité microbienne et la disponibilité du Hg(II), laquelle a

`son tour dépend de l'interaction

de facteurs environnementaux (température, conditions redox, carbone organique et sulfates) induisant des réactions non linéaires

sur la production de MeHg. La bioaccumulation/bioampliﬁcation dans les chaines alimentaires aquatiques de l'Arctique est fonction

du réservoir du MeHg dans les compartiments abiotiques, aussi bien que de considérations écologiques telles que la longueur de la

chaîne alimentaire, des taux de croissance, des caractéristiques des cycles de vie, ainsi que des comportements nutritionnels et

interactions trophiques. Les teneurs en MeHg dans le biote arctique ont augmenté signiﬁcativement depuis le début de l'ère indus-

trielle, et chez certaines populations de poissons, d'oiseaux de mer et de mammifères marins, les seuils toxicologiques sont dépassés.

Dû a

`la relation complexe entre l'exposition au Hg et la santé humaine des peuples nordiques —provenant du double rôle des

aliments du pays a

`la fois comme source potentielle de Hg et a

`la fois comme source d'aliments nutritifs et abordables avec plusieurs

avantages pour la santé physique et sociale —les réductions des émissions anthropogènes du Hg s'avèrent comme la seule solution

viable a

`long terme. [Traduit par la Rédaction]

Mots-clés : mercure, méthylation, Arctique, bioaccumulation, eau douce, marin.

Received 16 August 2013. Accepted 10 January 2014.

I. Lehnherr. Department of Earth and Environmental Sciences, University of Waterloo, 200 University Ave. W., Waterloo, ON N2L 3G1, Canada.

E-mail for correspondence: ilehnher@uwaterloo.ca.

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Introduction

Mercury (Hg) is a global pollutant due to its ability to undergo

long-range transport from source regions to remote parts of the

world, and its ubiquitous presence in aquatic ecosystems and food

webs is an important environmental and human-health issue

(Fitzgerald and Clarkson 1991 ). This is particularly true in the

Arctic where elevated concentrations of Hg exceeding toxicologi-

cal thresholds have been reported in both wildlife and humans

(AMAP 2011 ; Dietz et al. 2013). This paper provides an introduction

to Hg biogeochemistry—including chemical and toxicological

properties of Hg, Hg uses and emissions, and basic Hg cycling—

before reviewing methylmercury (MeHg

) production and bioac-

cumulation in Arctic aquatic ecosystems. Due to its toxicity and

ability to bioaccumulate and biomagnify in food webs, MeHg is

the form of Hg of greatest concern (Morel et al. 1998). Therefore,

the various sources of MeHg in both Arctic marine and freshwater

ecosystems are reviewed as well as the environmental factors that

regulate the microbial methylation of inorganic Hg(II). Microbial

and photochemical pathways for the demethylation of MeHg are

also discussed. Methylmercury bioaccumulation and biomagniﬁ-

cation in Arctic aquatic food webs are examined, particularly in

the context of key ecological factors that, in addition to the stand-

ing stock of MeHg in abiotic compartments, regulate MeHg con-

centrations in higher trophic level organisms. Finally, the complex

connection between Hg exposure and human health in Northern

Peoples is described, highlighting the need for balanced messag-

ing on issues pertaining to contaminants.

Mercury as a contaminant

The incidents that took place in Minamata, Japan, provide a

now-infamous example of the potential dangers associated with

Hg contamination and led to the identiﬁcation of MeHg as the

cause of Minamata disease in 1958 (McAlpine and Araki 1958). As a

consequence of consuming locally caught ﬁsh and shellﬁsh con-

taminated with MeHg, many inhabitants from around Minamata

Bay developed symptoms consistent with organic Hg poisoning,

including numbness in hands and feet, loss of ﬁne motor skills,

memory loss, blindness, and in severe cases culminating in coma

and death (Takeuchi et al. 1962). The source of the MeHg was

determined to be a local chemical plant (Chisso Co. Ltd.) that

manufactured acetaldehyde using mercuric sulfate as a catalyst

(Irukayama 1977), resulting in the accidental production of MeHg.

The discharge of MeHg in efﬂuent resulted in extremely high

MeHg concentrations in sediments (up to 2010 ␮ gg

−1

) and marine

biota within Minamata Bay (Clarkson and Magos 2006), later lead-

ing scientists to realize that MeHg is transferred through aquatic

food webs, from seston and (or) sediments to phytoplankton, zoo-

plankton, and ﬁsh (Nishimura and Kumagai 1983). Indeed, MeHg

typically increases in concentration with each increase in trophic

level in the food chain, a process called biomagniﬁcation (Fig. 1).

Although inorganic Hg, typically through the inhalation of

elemental Hg vapor, also has deleterious health effects (e.g., mad

hatter syndrome), exposure to MeHg, primarily through the con-

sumption of ﬁsh and marine mammals, is the principal concern

with respect to human health (Clarkson and Magos 2006 ; Fitzgerald

and Lamborg 2004). Methylmercury has many detrimental human-

health effects, primarily neurological and including distal sensory

disturbances, constriction of visual ﬁelds, loss of muscle control,

dysarthria (speech disorder), auditory disturbances, and tremors

(Mergler et al. 2007). In addition to the nervous system, MeHg can

also affect the cardiovascular and immune systems (Mergler et al.

2007). There is increasing evidence that even low-level exposure to

MeHg can have detrimental health effects with respect to cardio-

vascular disease in adults and neurological development in fe-

tuses and young children (Karagas et al. 2012). Methylmercury is

actively transferred across the placenta (Kajiwara et al. 1996), with

concentrations becoming higher in the fetus compared with those in

maternal blood (Walker et al. 2006), which can lead to congenital

MeHg poisoning even in cases when the mother exhibits no mani-

festation of poisoning (Harada 1978). Exposure to MeHg in utero can

lead to slower neurologic development resulting in lower perfor-

mance on tests of language, attention, memory, and visuospatial and

(or) motor functions (Mergler et al. 2007). A recent study estimated

that MeHg exposure throughout countries of the European Union re-

sults in long-term intelligence quotient (IQ) deﬁcits in 1.8 million chil-

dren born each year, with an associated societal cost of €8–9 billion

annually (Bellanger et al. 2013). Similarily, the socioeconomic cost of

MeHg exposure is predicted to reach $3.7 billion annually in the

United States by 2020 (Sundseth et al. 2010).

Populations with high consumption of ﬁsh and (or) marine

mammals, such as Northern Peoples (e.g., Inuit), are particularly

at risk of elevated exposure to MeHg. For example, average hu-

man blood Hg concentrations up to 40 ␮ gL

−1

, well above the US

Environmental Protection Agency's blood Hg guideline of 5.8 ␮ gL

−1

have recently (2000–2004) been observed in northern commu-

nities where diets are composed predominantly of traditional

foods such as Arctic char (Salvelinus alpinus), ringed seals (Phoca

hispida), and beluga whales ( Delphinapterus leucas)( AMAP 2011).

Mercury chemistry

Mercury is a transition (d-block) metal, is the only metal to be

liquid at room temperature (melting point = −38.8 °C), and its

relatively high vapor pressure (0.18 Pa) means that elemental Hg

can exist as a monoatomic gas under natural conditions. Mer-

cury has seven stable isotopes (

196

Hg, 0.15% abundance;

198

Hg,

10.0%;

199

Hg, 16.7%;

200

Hg, 23.2%;

201

Hg, 16.7%;

202

Hg, 29.8%; and

204

Hg, 6.8%) (Lauretta et al. 2001 ) and two longer lived radioisotopes

(

194

Hg, t

1/2

= 444 years; and

203

Hg, t

1/2

= 46.6 days). Mercury occurs

in three different oxidation states: 0, 1, or 2, although Hg(I) com-

pounds are only rarely encountered in the environment. The stan-

dard reduction potential for the Hg(0)/Hg(II) redox pair (E

= 0.851 V,

Lide 2007) is in the range of commonly encountered environmental

redox conditions (as deﬁned by E

), resulting in dynamic reduction

and oxidation cycling in atmospheric, aquatic, and terrestrial com-

partments. In terms of reactivity and binding, Hg is considered a

soft acid and, therefore, according to the Hard–Soft Acid–Base

Theory, forms stable bonds and compounds with soft bases such

as thiols, sulﬁdes, and other ligands containing reduced sulfur.

Thus, it is not surprising that the principal mineralized form of Hg

is cinnabar (HgS) (Fitzgerald and Lamborg 2004), a mineral with

very low solubility (log K

= −36.8, Martell et al. 1998).

Mercury compounds containing at least one Hg–carbon covalent

bond are classiﬁed as organic Hg compounds, MeHg and dimethyl-

mercury (DMHg) being two examples that occur in the environment.

Methylmercury can interact with inorganic or organic ligands to

form, for example, MeHgCl, MeHgOH, or larger complexes with dis-

solved organic matter (DOM), the nature of the ligand dictating

whether the MeHg moiety will be hydrophilic or hydrophobic

(Hintelmann 2010). While Hg(II) can also form complex ions with

DOM, these are not considered organic Hg compounds, as they lack

a covalent Hg–carbon bond. Dimethylmercury is soluble (water sol-

ubility = 2.95 g L

−1

) and volatile, and its hydrophobic nature (K

180–380) combined with high toxicity makes it a dangerous

neurotoxin: a few drops coming into contact with skin proved to be

fatal in one infamous case (Nierenberg et al. 1998). Like MeHg, DMHg

Methylmercury (MeHg) is used here to indicate any compound containing the MeHg

cation; the term monomethylmercury (MMHg; not used in this

review) is sometimes used to distinguish MeHg from dimethylmercury (DMHg).

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is easily photodegraded but relatively chemically stable except to-

wards strong oxidizing agents (Hintelmann 2010). However, unlike

MeHg, there is no data at this point that shows that DMHg is bioac-

cumulated, and thus it is thought that DMHg does not biomagnify

through food webs (Mason et al. 1996).

Mercury uses and emissions

Due to the many useful properties of Hg, cinnabar deposits have

been mined by humans for thousands of years. For example, the

onset of cinnabar extraction from the Huancavelica mine in the

Peruvian Andes has been estimated at 1400 BC (Cooke et al. 2009 )

and the Almadén mine in Spain has been active for over

2500 years (Goldwater 1972). Historically, Hg in its various forms

was used as a pigment (red cinnabar), as a treatment for syphilis,

and by alchemists attempting to turn Hg into gold, HgNO

was

also used to treat the fur used in the preparation of felt for hat-

making. Hg(0) is still used to this day to extract gold and silver in

mining operations (Clarkson and Magos 2006) because of its abil-

ity to form alloys—known as amalgams—with these metals. In

the industrial era, Hg has been used in tooth ﬁllings, batteries,

ﬂuorescent light ﬁxtures, as an electrode in the production of

chlorine and caustic soda (i.e., chlor-alkali production), and in

applications requiring a heavy liquid, such as thermometers, ba-

rometers, and electrical switches (Clarkson and Magos 2006). Or-

ganic Hg compounds have been used in agricultural applications

as fungicides; however, their use has been discontinued following

the accidental mass MeHg poisoning of both humans and wildlife

in Iraq (Bakir et al. 1973).

Anthropogenic emissions of Hg to the environment are largely

from the combustion of fossil fuels, such as oil and coal which

contain Hg as a trace contaminant (e.g., in coal-powered electrical

generation plants), representing ⬃35%–45% of total anthropo-

genic Hg emissions (Pirrone et al. 2010 ; Pacyna et al. 2010). Other

important sources of anthropogenic Hg emissions include artisanal

gold mining, waste incineration, metal smelting and production

(e.g., iron, steel, copper, lead, and zinc), and cement production

(Pacyna et al. 2010 ; Pirrone et al. 2010). While emissions of anthro-

pogenic Hg to the atmosphere have decreased since 1990 in many

industrialized countries in Europe and North America, principally

due to the implementation of emission control technology, these

decreases have been negated by increased Hg emissions in devel-

oping countries, particularly in Asia, where coal combustion has

increased (Pacyna and Pacyna 2002 ; AMAP 2011). Mercury can also

be emitted to the atmosphere from natural sources, often in the

form of gaseous elemental Hg(0). Natural waters are usually su-

persaturated with Hg(0) (Morel et al. 1998), and thus the surface

oceans are the largest natural source of Hg(0) to the atmosphere.

Additional natural Hg sources include terrestrial surfaces (soil

and vegatation), forest ﬁres or biomass burning, and geogenic

sources such as volcanoes (Mason and Sheu 2002 ; Pirrone et al.

2010).

Estimates of natural (or pre-industrial) Hg emissions range be-

tween 2000 and 3700 Mg year

−1

, accounting for ⬃ 1/3 of all emissions

(Selin et al. 2008 ;Sunderland and Mason 2007). By comparison,

present-day anthropogenic emissions of Hg to the atmosphere are

similar to 2000 Mg year

−1

(Pirrone et al. 2010 ;Pacyna et al. 2010 ;

Streets et al. 2011). However, an additional component of anthropo-

genic Hg emissions that has to be considered is the re-emission of

previously deposited Hg (legacy Hg), which has been estimated to be

2500–4100 Mg year

−1

(Selin et al. 2008 ;Sunderland and Mason 2007 )

and is, therefore, equal to or greater than current direct anthropo-

genic emissions. As a consequence of human perturbations to the Hg

biogeochemical cycle, Hg concentrations have approximately tripled

in the atmosphere and surface ocean since the beginning of the

industrial age (Mason et al. 2012 ; Mason et al. 1994 ; Lamborg et al.

2002a ). Increases in the size of the subsurface (20%–25%) and deep

ocean (⬃10%) Hg reservoirs have been more modest, due to the lon-

ger time required for anthropogenic Hg to be transported to those

depths (Mason et al. 2012 ; Lamborg et al. 2002a ). Furthermore, the

increase in anthropogenic Hg emissions since the onset of in-

dustrialization is reﬂected in environmental archives of atmo-

spheric Hg deposition such as lake sediments, peat bogs, and

Fig. 1. Methylmercury (MeHg) bioaccumulation and biomagniﬁcation in a typical Arctic marine food web. The inset shows the range of MeHg

concentrations in aqueous abiotic compartments, as well as in different groups of organisms, illustrating how MeHg concentrations increase

with each trophic level as a result of biomagniﬁcation.

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ice or ﬁrn (Lamborg et al. 2002b ;Muir et al. 2009;Goodsite et al.

2013). Because recently deposited Hg is more reactive and tends to

bioaccumulate in aquatic biota more rapidly, it is expected that

reductions in Hg emissions will be effective in reducing environ-

mental Hg contamination on meaningful time scales (Harris et al.

2007). To this end, the United Nations Environment Programme's

Minamata Convention on Mercury was recently agreed (January

2013) with the goal of reducing anthropogenic Hg sources. The

treaty, signed by all circumpolar nations, includes a ban on new

Hg mines, a provision for the phase-out of existing mines, control

measures for atmospheric emissions, and regulations pertaining

to Hg use in artisanal gold mining (www.mercuryconvention.org).

The mercury biogeochemical cycle with special

reference to Arctic aquatic ecosystems

The chemical stability and long residence time of Hg(0) in the

atmosphere (0.5–1 year) allow it to be widely distributed across the

globe (Pacyna et al. 2006 ; Schroeder and Munthe 1998). Therefore,

anthropogenic Hg(0) emitted from point sources in industrialized

regions can undergo long-range transport and reach remote and

seemingly pristine regions such as the Arctic. The majority of Hg

in the atmosphere is in the form of Hg(0) (typically >95%) (Fitzgerald

and Lamborg 2004); however, Hg(0) can be oxidized by strong oxi-

dants such as ozone and halogens (e.g., Br radicals) through vari-

ous gas-, aerosol-, and aqueous-phase reactions (e.g., Shia et al.

1999). Oxidation of Hg(0) results in the formation of reactive gas-

eous Hg(II) and (or) particulate Hg(II), both of which have short

atmospheric residence times (minutes to weeks) and are thus ef-

ﬁciently removed from the atmosphere via wet and dry deposi-

tion to nearby landscapes (Mason and Sheu 2002). Once deposited,

Hg(II) can be either sequestered in soils and aquatic sediments or

reduced to Hg(0) by biotic (e.g., Barkay et al. 2003 ; Poulain et al.

2007) and photochemical (e.g., Amyot et al. 1994) mechanisms.

Hg(0) can then be re-emitted to the atmosphere through diffusion

and air–water gas exchange, thus completing the inorganic Hg

cycle (Fig. 2).

The oxidation and subsequent deposition of atmospheric Hg is

greatly enhanced in the Arctic during polar sunrise by atmo-

spheric mercury depletion events (AMDEs) (Schroeder et al. 1998).

Atmospheric mercury depletion events arise from the rapid

photo-oxidation of tropospheric Hg(0) catalyzed by halogen radi-

cals in marine aerosols (Lindberg et al. 2002). Because AMDEs

potentially result in the deposition of very large quantities of

atmospheric Hg to polar regions, they have been the focus of

intensive scientiﬁc investigation (Schroeder et al. 1998 ; Lindberg

et al. 2002;Skov et al. 2004;Steffen et al. 2008;Berg et al. 2003;

Gauchard et al. 2005). Early estimates of atmospheric Hg deposi-

tion to the Arctic ranged from 208 to 325 Mg year

−1

, 20%–57% of

which was attributed to AMDEs (Christensen et al. 2004 ; Ariya

et al. 2004;Travnikov 2005). However, it has now been demon-

strated that the majority of Hg(II) deposited during AMDEs is lost

from snowpacks within days (Kirk et al. 2006 ; Durnford and

Dastoor 2011), as a result of photoreduction (Lalonde et al. 2003;

Sherman et al. 2010), and emitted back to the atmosphere as Hg(0).

Therefore, more recent estimates of Hg(II) deposition to the area

north of the Arctic Circle, which account for snowpack processes

such as photoreduction, are in the range of 60–108 Mg year

−1

(Holmes et al. 2010 ; Dastoor and Durnford 2013).

Because the Arctic Ocean is semi-enclosed sea, it is strongly

inﬂuenced by riverine inputs of water, nutrients, and trace ele-

ments including Hg. The export of Hg to Arctic marine waters has

been quantiﬁed for various North American rivers, such as the

Mackenzie River, which delivers 1.9–3.5 Mg year

−1

of total Hg and

13–23 kg year

−1

of MeHg to the Arctic Ocean (Emmerton et al. 2013;

Graydon et al. 2009;Leitch et al. 2007). However, due to a lack of

measurements in the large Siberian rivers, the pan-Arctic input of

riverine Hg to the Arctic Ocean is still highly uncertain. For exam-

ple, the export value of 12.5 Mg year

−1

calculated from the limited

number of observations available (Outridge et al. 2008) is much

lower than the 80 Mg year

−1

estimated using a coupled ocean–

atmosphere model (Fisher et al. 2012). A third estimate, calculated

from more extensive measurements of riverine DOC exports to

the Arctic Ocean, and Hg-DOC ratios in northern rivers, yielded a

value of 50 Mg year

−1

(Dastoor and Durnford 2013). This value

supports the conclusion made by Fisher et al. (2012) that the input

of riverine Hg to the Arctic Ocean has previously been underesti-

mated. The fate of this Hg in marine waters remains largely un-

known and understudied; and because riverine Hg is primarily

bound to particulate (⬃90% of total Hg (THg) and ⬃50% of MeHg)

(Emmerton et al. 2013 ; Schuster et al. 2011), it may be less biogeo-

chemically active. Nonetheless, current data suggest that atmo-

spheric and riverine inputs are the two most important sources of

Hg to Arctic marine ecosystems (Kirk et al. 2012 ; Fisher et al. 2012).

In addition to the processes described earlier in the paper (de-

position, sequestration, reduction, and volatilization), Hg(II) can

also be methylated to form MeHg. While MeHg production in

terrestrial systems is thought be negligible (Hintelmann 2010),

aquatic environments, especially anaerobic environments such as

freshwater and marine sediments (e.g., Gilmour and Riedel 1995;

Hollweg et al. 2010;Macalady et al. 2000;Lehnherr et al. 2012b ) and

hypolimnetic waters (Eckley et al. 2005 ; Eckley and Hintelmann

2006), are important sites of Hg(II) methylation. Hg(II) methylation is

primarily carried out by microorganisms such as sulfate-reducing

bacteria (King et al. 2000), iron-reducing bacteria (Fleming et al.

2006), and methanogenic archea (Parks et al. 2013 ; Hamelin et al.

2011). Additional microbial taxa capable of methylating Hg(II) have

also been identiﬁed (Gilmour et al. 2013) following the discovery of a

gene cluster responsible for MeHg production (Parks et al. 2013).

Abiotic mechanisms of Hg(II) methylation have been proposed

(Weber 1993), but it is generally thought that these are not likely to

be environmentally relevant in most aquatic ecosystems.

Although both Hg(II) and MeHg can be taken up and retained

(bioaccumulated) by lower trophic level organisms such as mi-

crobes and algae (Mason et al. 1996), Hg(II) is not biomagniﬁed

through food webs in the same manner as MeHg (Fitzgerald and

Clarkson 1991;Foster et al. 2012). Thus, most of the Hg occurring in

tissues of upper trophic level organisms, such as predatory spe-

cies, is in the form of MeHg (Morel et al. 1998), and the uptake and

biomagniﬁcation of Hg through food webs are initially con-

strained by the production of MeHg via the methylation of Hg(II).

Unfortunately, because Hg exists as a number of interconverting

chemical species that have the ability to transfer across environ-

mental compartments, the ultimate source of MeHg to aquatic

food webs has remained elusive in many instances. This is espe-

cially true in the Arctic, where MeHg sources to both marine and

freshwater ecosystems remained poorly characterized and quan-

tiﬁed, until recently.

Methylmercury sources in Arctic marine ecosystems

Seafood consumption remains the primary source of MeHg to

humans globally (Mergler et al. 2007 ; Sunderland 2007). This is a

concern for many Northern Peoples, for example, whose tradi-

tional diets are composed primarily of marine mammals and ﬁsh

(Donaldson et al. 2010). The biogeochemical cycling of Hg in ma-

rine environments is distinctly different than in freshwaters be-

cause DMHg, which is not usually detected in freshwaters, can

often be the dominant organic Hg species in seawater (Mason et al.

1995; Mason and Sullivan 1999;Mason and Fitzgerald 1993). As

discussed later in the paper, DMHg may represent both a direct

and indirect source of MeHg to marine organisms and is, there-

fore, a key component of the marine Hg cycle. In estuaries and

near-shore environments, rivers (Sunderland et al. 2012) and coastal

marshes (Mitchell et al. 2012 ) are small-to-moderate sources of

MeHg. Hg(II) methylation in shelf sediments on the other hand

can result in signiﬁcant export of MeHg to overlying waters (Hollweg

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et al. 2009, 2010; Hammerschmidt et al. 2004;Hammerschmidt and

Fitzgerald 2006a) and, through lateral transport, to waters near the

continental shelf break (Hammerschmidt and Fitzgerald 2004).

However, in open-ocean environments, increasing evidence points

to Hg(II) methylation in the water column, rather than bottom sedi-

ments, as the principal source of MeHg (Mason and Fitzgerald 1990 ;

Sunderland et al. 2009;Cossa et al. 2009, 2011; Mason et al. 2012;Kirk

et al. 2012).

It has long been suggested that, in marine environments, sig-

niﬁcant MeHg production occurs at intermediate depths in the

water column (e.g., Mason and Fitzgerald 1990). Sunderland et al.

(2009) demonstrated that methylated Hg (i.e., MeHg + DMHg) con-

centrations at various depths and locations in the Paciﬁc Ocean

were correlated with indices of organic carbon remineralization,

suggesting that MeHg production in the water column was asso-

ciated with heterotrophic respiration in intermediate waters

(150–1000 m depths). Similarly, methylated Hg concentrations in

the Mediterranean Sea were correlated with apparent oxygen uti-

lization and phosphate concentrations throughout the water col-

umn (Cossa et al. 2009; see also Cossa et al. 2011), as well as the

abundance of nano- and pico-phytoplankton in surface waters

(Heimbürger et al. 2010), supporting the hypothesis that MeHg

production is controlled by both the availability and remineral-

ization of organic matter. Hg(II) methylation in marine waters is

thus closely linked to the carbon cycle through primary produc-

tion and heterotrophic microbial respiration. It is also interesting

to note that none of the methylated Hg depth proﬁles in the open

ocean showed any indication of a signiﬁcant sedimentary source.

Even though the highest methylated Hg concentrations are found

in subsurface waters, experiments using Hg stable isotope tracers

showed that Hg(II) methylation also occurs in surface seawater

(Monperrus et al. 2007). It has recently been suggested that the ma-

jority of MeHg bioaccumulated into pelagic food webs is actually

produced within the surface mixed layer (Hammerschmidt and

Bowman 2012). This hypothesis is supported by measurements of Hg

stable isotope ratios in ﬁsh from the Gulf of Mexico, which suggest

that the MeHg that is bioaccumulated and biomagniﬁed in pelagic

foodwebs has previously undergone photochemical degradation

near the ocean's surface (i.e., photodemethylation, see MeHg demeth-

ylation section) (Senn et al. 2010). The seven stable isotopes of Hg

are fractionated in a mass-dependent manner during most

chemical and biological transformations, such that the lighter

isotopes react preferentially compared with the heavier iso-

topes (Kritee et al. 2009). However, during photochemical processes

such as MeHg photodemethylation, Hg isotopes also fractionate in a

mass-independent manner because odd isotopes, which have a mag-

netized nucleus, react more slowly (Bergquist and Blum 2007 ). The

isotope signature imparted by mass-independent fractionation is re-

tained during Hg bioaccumulation and biomagniﬁcation and can,

therefore, be used as a tracer of photochemical processes for Hg in

biota (Bergquist and Blum 2007). Using this approach, Senn et al.

(2010) demonstrated that MeHg in pelagic ﬁsh from the Gulf of

Mexico had undergone substantial photodemethylation (⬃50%)

prior to bioaccumulation, suggesting that most of the MeHg in ﬁsh

originated from the euphotic zone near the ocean's surface.

Although less data have been collected in Polar Regions, similar

patterns are also emerging with respect to MeHg sources. For

example, in the Canadian Arctic Archipelago, depth proﬁles of

methylated Hg species showed an increase in concentration in

subsurface waters, but the proﬁle shape was not consistent with a

sediment MeHg source, again suggesting that the source of MeHg

to polar marine waters was production within the water column

(Kirk et al. 2008). Furthermore, methylated Hg concentrations in

the Southern Ocean (Cossa et al. 2011) and Canadian Arctic marine

waters (Lehnherr et al. 2011 ; Wang et al. 2012) also correlated with

apparent oxygen utilization, consistent with the hypothesis that

the remineralization of organic matter plays an important role in

MeHg production. The biogenic origin of methylated Hg com-

pounds in the polar marine waters is supported by both laboratory

and ﬁeld measurements. For instance, laboratory incubations of

pure cultures of polar marine bacteria (Pongratz and Heumann 1999 )

and macroalgae from an Arctic fjord (Pongratz and Heumann

Fig. 2. Biogeochemical cycle of mercury and methylmercury depicting the main transport and transformation processes, including the

methylation of inorganic Hg to MeHg, in Arctic ecosystems (modiﬁed from Kirk 2006 and used with permission from the Artic Institute of

North America).

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1998a ) resulted in the production of both MeHg and DMHg. In

addition, DMHg and MeHg concentrations in Arctic and Antarctic

waters were correlated with chlorophyll a concentrations (Pongratz

and Heumann 1998b), linking primary production to Hg(II) meth-

ylation. Direct measurements of Hg(II) methylation in polar ma-

rine waters using Hg stable isotope tracers showed that Hg(II) can

be methylated to form both MeHg and DMHg and modeled water

column MeHg production rates based on these measurements are

high enough to account for approximately half of the MeHg pres-

ent in marine waters of the Canadian Arctic Archipelago (Lehnherr

et al. 2011). Furthermore, Hg(II) methylation at the depth of maxi-

mum phytoplankton abundance accounted for ⬃80% of MeHg

occurring there (Lehnherr et al. 2011), highlighting the previously

underestimated importance of methylation in oxic surface wa-

ters. Demethylation, or decomposition, of methylated Hg in ma-

rine waters limit how far MeHg can be transported via ocean

currents; it is estimated that 90% of MeHg in a particular water

mass would be demethylated by the time it travelled 200 km

through the Canadian Arctic Archipelago, demonstrating the im-

portance of local MeHg sources to Arctic marine food webs

(Lehnherr et al. 2011). Because MeHg production in marine waters

is limited by the availability of Hg(II), increases in Hg(II) concen-

trations, as have been observed in certain parts of the Paciﬁc

Ocean due to increased anthropogenic emissions (Sunderland et al.

2009), will lead to higher MeHg concentrations and increased expo-

sure to Hg in human and wildlife populations (Lehnherr et al. 2011).

Experiments with diatoms suggest that DMHg is not bioaccu-

mulated into phytoplankton (Mason et al. 1996), despite its hy-

drophobic properties (log K

= 2.59), which contrarily suggest

that DMHg is likely to be retained in biological tissues. However,

DMHg can be an indirect, but important, source of Hg to biota, as

it likely decomposes to MeHg in marine waters (Mason et al. 1995)

or after evading to the atmosphere. The efﬂux of DMHg from

Arctic marine waters combined with its photochemical decompo-

sition to MeHg in the atmosphere (Niki et al. 1983 ; Sommar et al.

1996) means that DMHg is a potential source of MeHg to Arctic

marine and coastal ecosystems (Kirk et al. 2008 ; St. Louis et al.

2007). Dimethylmercury is also thought to be the source of ele-

vated MeHg concentrations in fogwater in coastal California, result-

ing in signiﬁcant deposition of MeHg to coastal areas (Weiss-Penzias

et al. 2012). High MeHg concentrations in fog were associated with

upwelling events, which enhance the vertical transport of DMHg-

enriched subsurface waters (Conaway et al. 2009) and the subse-

quent evasion of DMHg to the atmosphere (Weiss-Penzias et al.

2012). No measurements of MeHg in fog have been conducted in

the Arctic to date, although both DMHg and MeHg have been de-

tected in air above the seawater–air interface (Baya and Hintelmann

2013). An important knowledge gap impeding our ability to quan-

tify the importance of DMHg as a source of MeHg to marine food

webs is the lack of rate measurements for the interconversion

between MeHg and DMHg in the water column (Lehnherr et al.

2011). To solve the puzzle posed by the fact that MeHg concentra-

tions can simultaneously be relatively low in seawater yet rela-

tively high in marine biota, it is necessary to understand what role

DMHg plays either as a direct or indirect source of Hg to marine

food webs.

In coastal areas, Arctic rivers also contribute signiﬁcant quan-

tities of Hg (see earlier in the paper) including MeHg (Graydon

et al. 2009;Kirk and St. Louis 2009;Leitch et al. 2007;Emmerton

et al. 2013). However, MeHg exports have been measured in only a

handful of Arctic rivers and the total riverine input of MeHg to

Arctic marine waters cannot, therefore, be reliably calculated.

Kirk et al. (2012) derived a very approximate estimate from the

observation that MeHg makes up 0.5%–5% of THg in Arctic rivers

and assuming a THg export of ⬃50 Mg year

−1

to the Arctic Ocean

(see earlier in the paper), the corresponding MeHg export is, there-

fore, somewhere in the range of 0.25–2.5 Mg year

−1

. Furthermore,

the fate of riverine MeHg upon entering marine waters and the

importance of this MeHg source to organisms feeding in estuaries

have not yet been examined. Thus, in many ways our knowledge

of MeHg sources and the mechanisms resulting in high MeHg

concentrations in Arctic marine food webs is still incomplete, and

it was recently suggested that research on the biogeochemical

cycle of Hg in Arctic regions needs to focus on methylation pro-

cesses within the ocean (Macdonald and Loseto 2010).

Methylmercury sources in Arctic freshwater ecosystems

Methylmercury production in freshwater ecosystems can take

place in bottom sediments (e.g., Gilmour and Riedel 1995), anoxic

hypolimnetic lake waters (Eckley and Hintelmann 2006 ; Eckley

et al. 2005), periphyton bioﬁlms (Desrosiers et al. 2006), and moss

mats (St. Louis et al. 2004 ; Yu et al. 2010). Water column MeHg

production is unlikely to be signiﬁcant in most Arctic lakes, as

they often do not develop an anoxic hypolimnion in the summer

in the same manner as temperate dimictic lakes. In temperate and

boreal regions, wetlands in particular have been identiﬁed as im-

portant sources of MeHg to downstream ecosystems (St. Louis

et al. 1994, 1996 ). However, most of our knowledge pertaining to

MeHg sources in freshwaters is derived from studies conducted in

the summer, and data collected in other seasons are required to

complete our understanding of MeHg cycling on an annual scale.

Wetlands are relatively abundant in the North, covering ap-

proximately 5% of the Canadian Arctic landscape (North American

Wetlands Conservation Council 2003) for example, and therefore

represent a potentially signiﬁcant source of MeHg. Unfortunately,

the limited data that are available on the importance of wetlands

as a MeHg source in Arctic regions are often contradictory. Exper-

iments have indicated that soils from Arctic wetlands can meth-

ylate Hg(II) under laboratory conditions, even if sulfate reducing

bacteria may not be the primary methylating bacteria (Loseto

et al. 2004a ;Oiffer and Siciliano 2009). However, in situ measure-

ments in the same soils showed a decline in MeHg concentration

over the summer growing season, suggesting that wetland soils

are in fact MeHg sinks, except perhaps for a short period during

snowmelt (Oiffer and Siciliano 2009). On the other hand, small

(<1 ha), warm wetland ponds in the High Arctic were shown to

have higher concentrations of aqueous MeHg and a higher per-

centage of THg as MeHg (%MeHg) compared with larger, colder

ponds and lakes in the same area (%MeHg = 19.5% versus 4.5%)

(St. Louis et al. 2005). This suggests that small wetland ponds may be

important for Hg(II) methylation and MeHg production. Methyl-

mercury mass-balance budgets constructed for a subset of these wet-

land ponds conﬁrmed this hypothesis, with the measured net areal

MeHg export to the overlying water column following Hg(II) meth-

ylation in the sediments ranging between 1.8 and 40 ng m

−2

day

−1

(Lehnherr et al. 2012a ). These ﬂux values are comparable to what

has been measured in temperate wetlands and lakes, and the very

high MeHg concentrations measured in zooplankton at these sites

(up to 880 ng MeHg g

−1

dry weight, Lehnherr et al. 2012a ) demon-

strate that High Arctic ponds are a source of MeHg to local aquatic

food webs. Lake sediments can also be signiﬁcant sites of MeHg

production in the Arctic. For example, the major source of MeHg

to Alaskan tundra lakes with no wetlands in their watershed was

shown to be in situ benthic production and subsequent diffusion

from the sediments (Hammerschmidt et al. 2006).

An alternate source of MeHg, which may be signiﬁcant in the

Arctic, is Hg(II) methylation within the snowpack, especially dur-

ing spring-melt. Elevated MeHg concentrations in snow have been

measured in the Canadian Arctic (St. Louis et al. 2005), Greenland

(Ferrari et al. 2004), and Svalbard (Dommergue et al. 2010); and

snowmelt-fed tributaries are sometimes more important than

wetland sources in determining MeHg concentrations in down-

stream lakes (Loseto et al. 2004b ). Possible sources for MeHg in

Arctic snow include (1) deposition of atmospheric MeHg originat-

ing from marine DMHg emissions (see earlier in the paper), which

is supported by the relationship sometimes observed between

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MeHg and marine halogens (e.g., chloride) in surface snow (St. Louis

et al. 2005); and (2) methylation of Hg(II) within the snowpack

(Larose et al. 2010 ; Constant et al. 2007). The hypothesis that Hg(II)

could be methylated in snow was proposed following the observa-

tion that %MeHg increased in Subarctic snow during spring melt,

with MeHg concentrations reaching 0.7 ng L

−1

(Constant et al. 2007).

Furthermore, MeHg concentrations were correlated to the abun-

dance of heterotrophic bacteria, suggesting that Hg(II) meth-

ylation in snow is also a biotic process (Constant et al. 2007). Based

on the positive correlation between MeHg and methanesulfonate

observed in coastal snow in Svalbard, a possible mechanism for

Hg(II) methylation has been proposed which implicates the sulfur

cycle (Larose et al. 2010). However, simple calculations based on

Hg loads in snow and MeHg concentrations in ponds following

snowmelt suggest that Hg(II) methylation in snow is not a signif-

icant source of MeHg to aquatic systems, at least in certain loca-

tions (Lehnherr et al. 2012a ). Until Hg(II) methylation rates in snow

are measured directly, it remains challenging to evaluate the rel-

ative importance of atmospheric deposition and in situ methylation

to MeHg levels in snow (Douglas et al. 2012).

Biogeochemical controls on Hg(II) methylation

Whether in marine or freshwater ecosystems, Hg(II) methylation

and net MeHg production are subject to biogeochemical controls

which are important to identify to mitigate Hg contamination in

aquatic ecosystems. Conceptually, methylation rates can be thought

of as being a function of the quantity of bioavailable Hg(II) and the

activity of methylating bacteria (Hintelmann 2010):

(1) MeHg production ⫽bioavailable Hg共 II兲

× microbial activity

Furthermore,microbialactivity governs the rate constantsofmeth-

ylation (k

) and demethylation (k

), such that net methylation can

be expressed as

(2) Net MeHg production ⫽ k m Hg共II兲– k dMeHg

It was recently demonstrated that, in Arctic wetland ponds,

MeHg concentrations in sediments were primarily controlled by

the measured value of k

and the concentration of Hg(II) (Lehnherr

et al. 2012b). On the other hand, the demethylation potential (k

)

did not appear to be as important in controlling sediment MeHg

concentrations (Lehnherr et al. 2012b ). Quantifying the impact of

biogeochemical controls on Hg(II) methylation is complicated by

the fact that many environmental and biogeochemical parame-

ters can inﬂuence either, and often both, microbial activity and

Hg(II) bioavailability. For example, MeHg production increases at

lower pH values, either because Hg(II) bioavailability increases as

pH decreases (Golding et al. 2007) or because Hg(II) methylating

microorganisms dominate the microbial community at lower pH

(Winch et al. 2008). Redox conditions, sulfate, DOM, and temper-

ature are other important factors that control MeHg production

(Ullrich et al. 2001).

Redox is a key parameter, controlling Hg bioavailability by

controlling chemical speciation and ligand characteristics (Morel

et al. 1998). Furthermore, microbial demethylation is thought to

be favored over methylation under aerobic and oxidative condi-

tions (Ullrich et al. 2001). Recent data support this hypothesis, as

Hg(II) methylation was elevated in Arctic wetlands sustaining

greater levels of anaerobic microbial respiration and with limited

potential for oxidation (Lehnherr et al. 2012b ). Because sulfate-

reducing bacteria are often implicated in Hg(II) methylation (King

et al. 2000, 2001 ), MeHg production is favoured when redox con-

ditions and growth substrate availability favour sulfate reduction.

For instance, the occurance of elevated concentrations of nitrate

(NO

3−

) and other electron acceptors, such as Mn

and Fe

, which

are thermodynamically favoured over sulfate during microbial

respiration of organic matter, can lead to decreased activity of

sulfate-reducing bacteria and by extension-decreased rates of

Hg(II) methylation (Todorova et al. 2009).

The role of sulfate in Hg(II) methylation is particularly complex.

Low sulfate concentrations (up to 200– 500 ␮M) limit sulfate reduc-

tion and, by extension, Hg(II) methylation (Gilmour et al. 1992).

However, at high sulfate concentrations, higher rates of sulfate

reduction lead to increased sulﬁde concentrations, decreasing the

bioavailability of Hg(II) due to the precipitation of HgS and (or) the

formation of charged Hg-S complexes (e.g., Hg(SH)

−

and HgS

22−

which are less bioavailable than neutral Hg-S complexes (Gilmour

et al. 1998;Benoit et al. 2003). Thus, sulfate is often limiting in

freshwater environments, such that methylation is stimulated by

sulfate additions (Gilmour et al. 1992 ; Mitchell et al. 2008a ) but

present in excess quantities in estuarine and marine environ-

ments (Gilmour et al. 1998). Other metals, such as iron, which can

also compete with Hg to bind sulﬁde, may moderate the extent to

which sulﬁde limits Hg(II) bioavailability (Han et al. 2008).

Like sulfate, organic carbon controls MeHg production on mul-

tiple levels. The availability of organic carbon substrates can limit

microbial respiration and hence Hg(II) methylation rates (Drott

et al. 2008;Lambertsson and Nilsson 2006;Windham-Myers et al.

2009). In Swedish lakes, for example, Hg(II) methylation rates

were an order of magnitude higher in productive lakes compared

with unproductive lakes (Drott et al. 2008), and the experimental

removal of emergent vegetation in Californian wetlands led to

decreased MeHg production (Windham-Myers et al. 2009). Emer-

gent vegetation exude labile DOM, such as acetate, which stimu-

lates microbial sulfate and iron reduction in the rhizosphere and

in turn enhances Hg(II) methylation (Windham-Myers et al. 2009).

However, DOM also complexes Hg(II), thereby decreasing its bio-

availability to methylating bacteria (Miskimmin et al. 1992), ex-

plaining why porewater %MeHg was negatively correlated with

DOM in a boreal peatland (Mitchell et al. 2008b ). Therefore, bio-

geochemical controls on MeHg production are often site-speciﬁc.

In carbon-rich environments (e.g., peatlands), the availability of

electron acceptors for key microbial processes (e.g., sulfate for

sulfate reduction) is likely to control MeHg production. On the

other hand, in oligotrophic marine environments, it is the avail-

ability of labile organic carbon (electron donor) that limits MeHg

production, with the presence of excess sulfate a further con-

strain on Hg(II) bioavailability and methylation.

Increasing temperatures generally stimulate microbial activity,

resulting in increased Hg(II) methylation during the summer sea-

son (Hintelmann and Wilken 1995 ; Watras et al. 1995); however,

demethylation rates may also be higher at higher temperatures

(Matilainen et al. 1991), meaning that the effect of temperature is

unclear and may also be site-dependent. In the Arctic, however,

higher temperatures in a warmer climate are likely to translate

into increased MeHg production by increasing the duration of the

time period during which methylation can take place (i.e., the

time period between the onset of the spring thaw and fall freeze-up)

(Hintelmann 2010) or altering the distribution of wetlands on the

landscape as well as increasing the release of Hg(II) from perma-

frost (Macdonald et al. 2005).

Methylmercury demethylation

While many scientiﬁc studies have focused on Hg(II) methylation,

MeHg demethylation is equally important in regulating net MeHg

production and MeHg reservoirs in aquatic ecosystems (eq. 2). Meth-

ylmercury can be microbially demethylated in freshwater and ma-

rine sediments (Oremland et al. 1991 ; Marvin-DiPasquale et al. 2000),

as well as marine waters (Lehnherr et al. 2011 ; Monperrus et al. 2007),

but the importance of microbial demethylation in the water column

of lakes is unclear (Schaefer et al. 2004 ; Eckley and Hintelmann

2006). Microbial demethylation can occur through either an oxida-

tive pathway (Marvin-DiPasquale and Oremland 1998), producing

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Hg(II), CO

, and CH

, or a reducing mechanism mediated by the

enzyme organomercurial lyase, producing Hg(0) and CH

(Spangler

et al. 1973). Reductive demethylation dominates in marine and (or)

contaminated environments (e.g., Schaefer et al. 2004) and is consid-

ered a detoxiﬁcation mechanism, while oxidative demethylation

dominates in freshwater ecosystems with nonelevated MeHg con-

centrations and is likely mediated by methanogenic and sul-

fate-reducing bacteria (Oremland et al. 1991 ; Marvin-DiPasquale and

Oremland 1998;Marvin-DiPasquale et al. 2000).

In addition to microbial demethylation, MeHg can also be photo-

lytically (i.e., abiotically) decomposed by solar radiation in surface

waters of lakes (Sellers et al. 1996 ; Lehnherr and St. Louis 2009) and

oceans (Monperrus et al. 2007 ; Lehnherr et al. 2011 ; Whalin et al.

2007), resulting in MeHg being converted to Hg(II) and Hg(0). Mass-

balance considerations have demonstrated that photodemeth-

ylation is the most important sink of MeHg in oligotrophic boreal

lakes (Sellers et al. 2001), tundra lakes (Hammerschmidt et al. 2006),

and High Arctic wetland ponds (Lehnherr et al. 2012a ) and is, there-

fore, an important mechanism by which the transfer of MeHg into

aquatic food webs is reduced. Methylmercury photodemethylation

rates are directly proportional to the concentration of MeHg in the

water and are also strongly dependent on both irradiation intensity

and wavelength (Hammerschmidt and Fitzgerald 2010 ; Lehnherr

and St. Louis 2009). Incubation experiments have shown that, in the

absence of solar radiation, no loss of MeHg is observed, demonstrat-

ing that, at least in surface lake waters, MeHg demethylation is en-

tirely light-mediated (Sellers et al. 1996 , 2001 ; Lehnherr and St. Louis

2009; Hammerschmidt and Fitzgerald 2006b , 2010).

Methylmercury was degraded by ultraviolet (UV; 100–400 nm) ra-

diation in laboratory experiments using both artiﬁcial and natural

light sources (Inoko 1981 ; Suda et al. 1993 ; Gårdfeldt et al. 2001 ; Chen

et al. 2003). Therefore, it is not surprising that photodemethylation

rates throughout the water column of lakes are correlated to the

intensity of solar radiation at those depths. However, photodemeth-

ylation rates are sometimes best correlated with UV-A radiation

(320–400 nm; Lehnherr et al. 2012a , 2012b ) and other times with

photosynthetically active radiation (PAR, 400–700 nm, i.e., visible

light) (Hammerschmidt and Fitzgerald 2006b ). Controlled ﬁeld ex-

periments have demonstrated that UV radiation degrades MeHg

more rapidly; but due to the ability of PAR to penetrate deeper into

the water column, MeHg photodemethylation is mediated by a

combination of visible and UV radiation (Lehnherr and St. Louis

2009; Hammerschmidt and Fitzgerald 2010). For example, it was

estimated that UV radiation is responsible for 79% of MeHg photo-

demethylated in a DOM-rich dark water lake, and 58% in a clear

water lake where PAR is not attenuated as rapidly with depth

(Lehnherr and St. Louis 2009). Therefore, whole-lake MeHg photo-

demethylation ﬂuxes depend on both the demethylation rate con-

stant and the attenuation coefﬁcient for each type of radiation

(Lehnherr and St. Louis 2009). The attenuation of UV and PAR in

the water column is in turn controlled by solutes (e.g., DOM) and

particles, which absorb and scatter light, respectively.

In the Arctic, MeHg photodemethylation is not only limited by the

optical properties of the water column and incident radiation, which

exhibits dramatic seasonality at high latitudes, but also by ice cover

which plays an important role in regulating light levels in both fresh

and marine waters. It was recently shown that sea ice is the domi-

nant control on MeHg photodemethylation in the spring season

(April–June) and climate-change-induced sea-ice loss will result in

increased MeHg loss from photodemethylation in polar surface ma-

rine waters (Point et al. 2011).

DOM and other solutes such as Fe(III) and NO

3−

play a key role in

MeHg photodemethylation because when exposed to UV radia-

tion, they generate reactive species, such as hydroxyl radicals and

singlet oxygen, that degrade MeHg (Hammerschmidt and Fitzgerald

2010; Zhang and Hsu-Kim 2010). Finally, rates of photodemethylation

decrease at higher salinity and chloride concentrations (Black et al.

2012; Zhang and Hsu-Kim 2010). However, this does not necessarily

mean that MeHg photodemethylation ﬂuxes are lower in marine

waters, relative to freshwaters, because marine waters tend to be

more transparent to UV compared with most lakes.

Methylmercury bioaccumulation and

biomagniﬁcation in Arctic biota

The accumulation of MeHg in organisms' tissues over time (bio-

accumulation) and its efﬁcient transfer from prey to predators

across trophic levels (biomagniﬁcation) means that, in many

aquatic ecosystems around the world, higher trophic level spe-

cies, such as piscivorous ﬁsh and apex predators, have elevated

concentrations of MeHg. While MeHg concentrations are rela-

tively low in natural waters (usually subparts per trillion), they

can reach parts per million levels in ﬁsh, representing a biomag-

niﬁcation of more than a million-fold (Fig. 1). The greatest increase

in concentration during the transfer of MeHg in food chains

occurs not from biomagniﬁcation between trophic levels, but

through the uptake of MeHg from abiotic matrices, such as water

or sediments, into organisms at the base of the food chain (e.g.,

algae) (Fitzgerald et al. 2007 ; Pickhardt and Fisher 2007 ). The up-

take of MeHg into phytoplankton can occur through both diffu-

sion and active transport (Pickhardt and Fisher 2007). Conversely,

for organisms occupying higher trophic positions, direct uptake

of MeHg from surrounding water represents a small but some-

times nonnegligible intake; rather, these organisms acquire their

MeHg body burden primarily form the ingestion of prey items (Hall

et al. 1997;Hrenchuk et al. 2012). Methylmercury is efﬁciently bioac-

cumulated and biomagniﬁed due to the high afﬁnity of the MeHg

cation for thiol groups in proteins and its lipophilic nature (when

bound to ligands such as Cl

−

)( Harris et al. 2003), promoting its reten-

tion in both muscles and fatty tissues. As a result, the average pro-

portion of Hg in the MeHg form increases from ⬃10% in the water

column to 15% in phytoplankton, 30% in zooplankton and >95% in

ﬁsh (Watras and Bloom 1992). The transfer and biomagniﬁcation of

MeHg through food webs is also demonstrated by the good correla-

tion normally observed between an organism's trophic position, as

inferred from its ␦

N value, and Hg concentration (Cabana and

Rasmussen 1994;Atwell et al. 1998).

Concentrations of MeHg in Arctic biota are controlled by vari-

ous complex interacting processes which act on one or both of the

primary regulating factors: (1) the size of the abiotic pool of MeHg

available for bioaccumulation, and (2) the transfer of MeHg

through aquatic food webs (Loseto et al. 2008 ; St. Louis et al. 2011;

Kirk et al. 2012). As discussed earlier in the paper, the size of the

abiotic MeHg pool, which represents a bottom-up control on

MeHg concentrations in biota (AMAP 2011 ; St. Louis et al. 2011;

Kirk et al. 2012), is determined by the delivery of Hg(II) to Arctic

aquatic ecosystems from atmospheric, riverine, and oceanic

sources, as well as the efﬁciency with which this Hg(II) is methyl-

ated within Arctic aquatic ecosystems. Numerous drivers, such as

climate, primary production, and sea-ice cover affect the delivery

and methylation of Hg(II) (see also Stern et al. 2012). Ecological

factors, such as food-chain length, habitat selection, foraging be-

havior, and life history, are equally important in controlling

MeHg concentrations in biota (Loseto et al. 2008 ; St. Louis et al.

2011; Swanson and Kidd 2010). Because MeHg is biomagniﬁed,

longer food chains with more trophic levels tend to result in

higher MeHg concentrations in apex predators. For example, dif-

ferences in hair Hg concentrations between two populations of

polar bear (Ursus maritimus) in the Canadian Arctic were best ex-

plained by differences in food-chain length (St. Louis et al. 2011).

Methylmercury concentrations in water and the contribution of

pelagic versus benthic resources to the bears' diet were other

important determinants of hair Hg concentrations (St. Louis et al.

2011). In lower trophic levels such as zooplankton, it is important

to understand food-web linkages because the extent to which

MeHg is biomagniﬁed across trophic levels varies for different

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planktonic food webs (Foster et al. 2012). For example, In High

Arctic ponds, zooplankton communities containing Daphnia had

MeHg concentrations that were ﬁve times higher than copepod-

dominated communities because of species-speciﬁc processes which

allowed Daphnia to more strongly bioaccumulate MeHg (Chételat

and Amyot 2009).

The importance of ecological processes such as habitat selection

and foraging behavior on Hg exposure has also been demonstrated

for Beaufort Sea beluga whales (Loseto et al. 2008). Beaufort Sea be-

luga whales sexually segregate into groups with different summer

ranges, and thus different feeding ecology (Loseto et al. 2006 ), di-

rectly impacting their dietary Hg exposure (Loseto et al. 2008).

Belugas using shallow estuarine habitats and feeding on estuarine-

shelf food webs, which include coastal ﬁsh species such as Paciﬁc

herring, Arctic cisco, Least cisco, Saffron cod, and Arctic cod, had

the lowest Hg concentrations, followed by belugas selecting off-

shore ice-edge habitats and feeding primarily on Arctic cod at a

slightly higher trophic level (Loseto et al. 2008). Finally, individu-

als selecting deep waters with heavy ice cover and feeding on

benthic and epibenthic food webs, which included benthic am-

phipods and ﬁsh (e.g., ﬂounder and sculpin), had the highest Hg

concentrations. Belugas in this last group did not feed at a higher

trophic level compared with those in the other feeding groups,

rather their higher Hg concentrations was attributed to higher Hg

levels in potential prey items, which were associated with high Hg

environments (i.e., sediments) (Loseto et al. 2008). The difference

in Hg concentration between whales from different groups dem-

onstrates the importance of ecological processes such as habitat

selection and feeding behavior (Kirk et al. 2012). Similarly, for fresh-

water ﬁsh such as Arctic char and lake trout (Salvelinus namaycush),

life-history characteristics also impact Hg exposure. For instance,

anadromous (sea-run) ﬁsh tend to have lower Hg concentrations

than nonanadromous ﬁsh of the same species, due to a combina-

tion of faster growth and higher lipid content (higher C:N in mus-

cle tissue) in ﬁsh spending part of their life cycle feeding at sea

(Swanson et al. 2011). Climate-related shifts in diet can also affect

Hg exposure in Arctic biota, and Hg concentrations in eggs of

thick-bill mures (Uria lomvia) breeding in northern Hudson Bay

have recently increased as a result of a shift in prey availability

(Braune et al. 2011).

It has been hypothesized that the high Hg concentrations ob-

served in Arctic biota are due to recent changes in the global Hg

cycle caused by industrialization and climate change (Dietz et al.

2006a ; Outridge et al. 2002, 2005; Stern et al. 2012). Analyses of

hard tissues (teeth, hair, and feathers) of historical and modern

Arctic animals, such as beluga whales (Outridge et al. 2002 , 2005),

polar bears, and birds (Dietz et al. 2006a , 2006b ), suggest that Hg

concentrations in Arctic biota have increased 10-fold, starting in

the mid- to late-19th century (Dietz et al. 2009). Furthermore, it is

estimated that ⬃ 90% of Hg in present-day Arctic marine animals is

of anthropogenic origin (Dietz et al. 2009). What is most alarming

about these trends is that Hg concentrations in Arctic marine

biota have increased disproportionally compared with the esti-

mated increases of Hg in abiotic reservoirs (e.g., tripling of the

global surface ocean Hg pool; Mason et al. 2012).

While there are clear increases in Hg concentrations in certain

marine mammals and birds of prey over longer time scales (cen-

tury to millennia), the current short-term temporal trends for

circumpolar biota are less clear (see Rigét et al. 2011 for a review).

For marine mammals (polar bears, beluga whales, and ringed

seals) as well as seabirds (thick-billed murres and northern ful-

mars (Fulmarus glacialis)), there are examples of populations in

different areas of the Arctic demonstrating signiﬁcant increases

in Hg concentrations in the past 30 years or so (Rigét et al. 2011,

2007; Braune 2007). However, there are also many examples of

marine mammal populations not showing any signiﬁcant in-

creases in Hg over time, in part because small sample sizes and

large variability sometimes limit our ability to detect such trends

(AMAP 2011 and reference therein; Rigét et al. 2011 ; Braune et al.

2005). Similarly, most freshwater ﬁsh of the Canadian Arctic are

not showing any recent increases in Hg concentrations (Muir et al.

2005; Evans et al. 2005), but signiﬁcant increases have been re-

ported for a limited number of landlocked Arctic char and burbot

(Lota lota ) populations (Rigét et al. 2011 ; AMAP 2011). In general, a

larger proportion of biota in the Canadian and Greenland Arctic

are exhibiting recent increases in Hg concentrations than in the

North Atlantic and European Arctic, where Hg concentrations in

some reindeer and Arctic cod populations have actually decreased

over the past couple of decades (Rigét et al. 2011).

A recent review of Hg concentrations in Arctic biota revealed

that certain populations of species such as polar bear, beluga and

pilot (Globicephala melas) whales, hooded seal (Cystophora cristata),

certain seabird species, and land-locked Arctic char, have Hg con-

centrations in their tissues that exceed toxicological thresholds

(Dietz et al. 2013). Brain Hg concentrations in beluga whales from

the Beaufort Sea exceeded levels associated with neurochemical

effects in 14% of individuals (Ostertag et al. 2013). Furthermore,

certain pilot whale and polar bear populations have Hg concen-

trations in liver and kidneys, respectively, that are above thresh-

old values for liver and kidney damage (Sonne et al. 2010 ; Dietz

et al. 2013). Mercury concentrations in eggs of most seabird spe-

cies can also reach levels that may be high enough to have nega-

tive impacts on reproductive success, with possible impacts

including reduced hatchability and clutch size (Braune et al. 2006;

Wolfe et al. 1998;Dietz et al. 2013).

Mercury and human-health considerations for

Northern Peoples

Bioaccumulation and biomagniﬁcation of MeHg in Arctic food

webs have resulted in some ﬁsh and marine mammals, particu-

larly arctic char, ringed seals, and beluga whales, having Hg con-

centrations that are above Health Canada's ﬁsh consumption

guideline of 0.5 ␮ gg

−1

(AMAP 2011 ; NCP 2012 ; Braune et al. 2005;

Campbell et al. 2005;Lockhart et al. 2005). In the Arviat region of

western Hudson Bay, for example, average concentrations of Hg

in the muscle, liver, and kidney of beluga whales ranged from 0.88

to 12.5 ␮ gg

−1

in the 1980s and 1990s (Lockhart et al. 2005). Mercury

concentrations in beluga whale muscle and muktuk, which is the

edible portion of the skin and blubber and is often the most

sought after portion of beluga whales for consumption, are also

generally above 0.5 ␮ gg

−1

across the Canadian High Arctic (NCP

2012; Lockhart et al. 2005). While ringed seal liver Hg concentra-

tions are typically high, Hg concentrations in ringed seal muscle

tissue is much lower (⬃30x) than those in liver and, therefore,

usually do not exceed 1 ␮ gg

−1

(NCP 2012). Mercury concentrations

in freshwater ﬁsh of the Canadian Arctic generally fall below the

0.5 ␮ gg

−1

consumption guideline, with average length-adjusted

Hg concentrations usually in the range of 0.2–0.5 ␮ gg

−1

for lake

trout and <0.2 ␮ gg

−1

for Arctic char and whiteﬁsh (Coregonus

clupeaformis and Prosopium cylindraceum) in most lakes (NCP 2012;

Swanson and Kidd 2010). However, for all four of these ﬁsh spe-

cies, there are lakes where the average Hg concentration exceeds

the 0.5 ␮ gg

−1

consumption guideline (NCP 2012).

The elevated Hg concentrations in some ﬁsh and marine mam-

mal populations (e.g., ringed seal and beluga whale) that are used

as traditional food sources pose potential health risks for Northern

Peoples across the circumpolar Arctic (see Donaldson et al. 2010 for a

review). Blood Hg concentrations greater than 100 ␮ gL

−1

, consid-

ered high enough to pose risks of irreversible neurological dam-

age, were observed over the period of 1971–1992 in Northern

Peoples whose diets were composed of predominantly traditional

country foods (Wheatley and Wheatley 2000). By comparison, the

U.S. Environmental Protection Agency recommends maintaining

blood Hg concentrations below 5.8 ␮ gL

−1

(Clarkson and Magos

2006; Fitzgerald and Lamborg 2004). More recently (2000–2007),

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average human blood Hg concentrations up to 40 ␮ gL

−1

have been

reported in women of childbearing age in certain northern com-

munities where diets are also composed primarily of traditional

foods (AMAP 2011). However, blood Hg levels are decreasing in

certain parts of the Canadian Arctic, such as Nunavik, and a

health survey conducted in 2004 demonstrated that mean blood

concentrations (10.3 ␮ gL

−1

) were 32% lower than those in 1992

(Fontaine et al. 2008). Similarly, blood Hg concentrations have

decreased in many circumpolar regions, including parts of Alaska

and northern Sweden (Donaldson et al. 2010 ; AMAP 2011). This

decline has been attributed to a decreased consumption of tradi-

tional country foods and not to a decrease in Hg concentrations in

traditional foods, which appear to have remained generally un-

changed (Kirk et al. 2012). Despite the general decreasing trend,

Hg exposure continues to be high in certain regions of the Arctic,

such as parts of Greenland, where 90% of women of childbearing

age had blood Hg levels exceeding the 5.8 ␮ gL

−1

guideline (AMAP

2011). By comparison, the 50th percentile for blood Hg in U.S.

women was recently reported to be 0.72 ␮ gL

−1

(Department of

Health and Human Services 2012). Elevated Hg exposure in North-

ern Peoples has been linked with adverse effects on recognition

memory (Boucher et al. 2011) and cardiovascular health (Valera

et al. 2009).

Despite the link between consumption of traditional foods and

Hg exposure, it is critical to remember that the consumption of

traditional foods has many social, cultural, and physical health

beneﬁts and is a vital component of the culture of Northern Peoples

(Kuhnlein and Chan 2000). Some of the nonnutritional beneﬁts of

traditional foods include lower cost and improved mental health,

cultural morale, and social cohesion (Wheatley and Wheatley

2000), as well as reduced risk of obesity, diabetes, and loss of

ﬁtness, due to the high level of physical activity required to har-

vest traditional foods (Kuhnlein et al. 2004 ). Nutritionally, tra-

ditional foods are rich in many important nutrients, such as

omega-3 fatty acids, and a recent health survey conducted in

36 Arctic communities showed that the consumption of traditional

foods was associated with higher intake of proteins, micronutri-

ents, vitamins A and C, a lower intake of carbohydrates, saturated

fat, ﬁber, and a lower sodium:potassium ratio (Egeland et al. 2011).

It is apparent from these ﬁndings that the consumption of tradi-

tional foods has many beneﬁts, despite potentially resulting in

increased exposure to Hg. Therefore, communication about con-

taminants in traditional foods should be done very carefully to

ensure that it does not result in confusion, fear, changes to tradi-

tional lifestyles, and (or) unnecessary decreases in the consump-

tion of traditional foods (Furgal et al. 2005 ; Kirk et al. 2012).

Summary

Mercury has a unique combination of properties—such as high

vapor pressure, long atmospheric residence time, and rich redox

chemistry— which together facilitate deposition of anthropogenic

Hg to remote regions like the Arctic and generally dictate how Hg

cycles through the various earth system components. The com-

plex biogeochemical cycle of Hg makes source attribution com-

plex, and predicting MeHg concentrations in biota based on rates

of regional Hg deposition is extremely challenging due to the

number of postdepositional processes that affect both Hg(II) meth-

ylation and MeHg bioaccumulation and biomagniﬁcation.

Hg(II) methylation rates are a function of the quantity of bio-

available Hg(II) and the metabolic activity of microbial organisms

capable of producing MeHg. Key taxa implicated in Hg(II) meth-

ylation include sulfur-reducing and iron-reducing bacteria and

methanogenic archea (Parks et al. 2013), highlighting the importance

of anaerobic processes for MeHg production. Hg(II) methylation is

often, but not always, greatest near oxic–anoxic boundaries, where

substrate (labile carbon) and electron acceptors (e.g., sulfate) are

available. Methylmercury production varies nonlinearly with a

number of interacting variables, including temperature, redox

conditions, organic carbon, and sulfate.

In Arctic marine ecosystems, Hg(II) methylation in the water

column, rather than bottom sediments, is the primary source of

MeHg (Lehnherr et al. 2011 ; St. Pierre et al. 2014). Because a large

proportion of the Arctic Ocean is composed of productive conti-

nental shelves, Hg(II) methylation in Arctic marine waters, which

is closely linked to primary production and organic carbon remin-

eralization, is likely efﬁcient and sensitive to perturbations to

the carbon cycle. Dimethylmercury is a potentially important

source of Hg to biota, through decomposition to MeHg in air and

water. However, the importance of DMHg cannot be evaluated

until the potential for direct uptake of DMHg by marine plankton

and the interconversion of MeHg and DMHg in air and water are

better quantiﬁed. The surface ocean has not yet reached equi-

librium with the atmosphere and, therefore, seawater Hg(II) con-

centrations are predicted to continue increasing in the near

future (Mason et al. 2012). Because MeHg production in marine

waters is limited by the availability of Hg(II) (Lehnherr et al. 2011),

this will lead to higher MeHg concentrations and increased expo-

sure to MeHg in marine biota and humans.

In Arctic freshwaters, MeHg concentrations are a function of

two antagonistic processes, net Hg(II) methylation in bottom sed-

iments of ponds and lakes and MeHg photodemethylation in the

water column. The importance of alternate MeHg sources, such as

Hg(II) methylation in snow and the water column of lakes that

develop under-ice anoxia, has not yet been properly evaluated.

Any methylation or demethylation pathways that are associated

with the cryosphere (e.g., Hg(II) delivery from melting permafrost

and MeHg photodemethylation in water bodies with decreasing

ice-cover extent and duration) will be particularly sensitive to

climate change.

Mercury concentrations in ﬁsh and apex predators (e.g., polar

bears and beluga whales) depend on both the reservoir of MeHg in

abiotic compartments (i.e., water and sediment) and ecological

factors such as food-chain length, growth rates, feeding behavior,

and trophic interactions. Slow growth rates, long life spans, and

(or) long food chains make certain Arctic ﬁsh and marine mam-

mals particularly sensitive to Hg contamination. Therefore, it fol-

lows that MeHg concentrations in Arctic aquatic food webs are

impacted by inputs of anthropogenic Hg(II), the efﬁciency with

which it is subsequently methylated, and to any ecological changes

brought about by climate and environmental change. Mercury con-

centrations in Arctic biota have increased approximately 10-fold

since the onset of industrialization (Dietz et al. 2009 )— in some

cases exceeding toxicological thresholds (Dietz et al. 2013 )— as a

result of anthropogenic emissions, which have increased the quan-

tity of Hg actively cycling in the atmosphere and hydrosphere. What

is particularly alarming about these trends is that Hg concentrations

in Arctic biota have increased disproportionally compared with

those in abiotic reservoirs (e.g., tripling of the global surface ocean

Hg pool; Mason et al. 2012). Compelling evidence suggest that MeHg

concentrations in aquatic biota respond to reductions in Hg deposi-

tion; however, the response time is longer for large systems such as

oceans (decades and longer) compared with smaller lakes (years)

(Harris et al. 2007 ;Mason et al. 2012).

Despite the link between the consumption of country foods and

exposure to Hg, it is clear that harvesting and consuming tradi-

tional country foods have many social, cultural, nutritional, and

physical health beneﬁts for Northern Peoples. Therefore, commu-

nication pertaining to contaminants should be done with consid-

eration for the cultural context surrounding this issue. Reducing

the consumption of traditional foods is not a viable long-term

solution for decreasing Hg exposure in Northern Peoples; therefore,

the reduction of anthropogenic Hg emissions remains the best

means of minimizing Hg concentrations in Arctic ecosystems.

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Acknowledgements

The author is grateful to Vincent St. Louis for providing valu-

able insights and feedback and three anonymous reviewers

whose comments helped to improve the manuscript. Financial

support for this work was provided by the Weston Foundation

and Alberta Ingenuity (now Alberta Innovates).

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... Bioaccumulation within specific species drives concentration, with the large glaucous gull species maintaining high levels of Hg (4.9 μg g −1 ), PCB (3,326 ng g −1 ) and DDT (2,367 μg g −1 ) in both tissue and shell samples 110 . In heterogeneous wetland and pond environments that are reliant on the influx of nutrients from seabird guano, this toxic input may create additional ecosystem stress and degradation over time 108,111 . ...

... Quantifying the risks to rapidly changing Arctic ecosystems from entrained permafrost biochemical compounds is an ongoing challenge 9 . The dominant Arctic transport mechanisms of atmospheric remobilization 159,160 , hydrological transport 83,161,162 , direct transport (through human or animal migration) 110,163 and trophic-level flow all need to be assessed for dispersion potential 86,88,111 . The transport rates and efficiency differ tremendously between these mechanisms, releasing sequestered compounds into recently thawed ecosystems 160,164,165 . ...

... The worst effects of known toxins on human populations in the Arctic may be yet to come. The increasing load of chemical mixtures 158 from permafrost can bioaccumulate across generations, weakening the immune system and increasing vulnerability to new pathogens 37,38,88,111,119,125 . ...

The Arctic cryosphere is collapsing, posing overlapping environmental risks. In particular, thawing permafrost threatens to release biological, chemical and radioactive materials that have been sequestered for tens to hundreds of thousands of years. As these constituents re-enter the environment, they have the potential to disrupt ecosystem function, reduce the populations of unique Arctic wildlife and endanger human health. Here, we review the current state of the science to identify potential hazards currently frozen in Arctic permafrost. We also consider the cascading natural and anthropogenic processes that may compound the impacts of these risks, as it is unclear whether the highly adapted Arctic ecosystems have the resilience to withstand new stresses. We conclude by recommending research priorities to address these underappreciated risks.

... Anthropogenic Hg emissions have been distributed ubiquitously, even reaching remote ecosystems such as the Arctic (AMAP/ UN Environment, 2019). Mercury is easily methylated in the aquatic environment (Jernelöv and Martin, 1975;Paranjape and Hall, 2017;Poulain and Barkay, 2013) and the resulting Hg species, methyl Hg (MeHg), is prone to bioaccumulation in biota followed by biomagnification through the food web (Atwell et al., 2011;Lavoie et al., 2013;Lehnherr, 2014). The highest concentrations are found in aquatic biota, and especially in top predators, integrating the contamination of their food web (AMAP/ UN Environment, 2019;Scheuhammer et al., 2007). ...

... Inter-but also intra-specific differences in Hg burden depend on diet, geographic location, large-and small-scale habitat characteristics and point sources of contamination (Lodenius and Solonen, 2013). The main diet factors affecting Hg burdens include trophic position (Lavoie et al., 2013;Lehnherr, 2014;Sebastiano et al., 2017) and the presence of fish species Da Silva et al., 2005) or other aquatic prey in the diet (Barnes and Gerstenberger, 2015;Lodenius and Solonen, 2013). Large-scale habitat differences in Hg-burden occur between aquatic and terrestrial habitats and freshwater and coastal habitats (Scheuhammer et al., 2007;Swanson et al., 2011). ...

Human-induced mercury (Hg) contamination is of global concern and its effects on wildlife remain of high concern, especially in environmental hotspots such as inland aquatic ecosystems. Mercury (Hg) biomagnifies through the food web resulting in high exposure in apex predators, such as the white-tailed eagle (Haliaeetus albicilla), making them excellent sentinel species for environmental Hg contamination. An expanding population of white-tailed eagles is inhabiting a sparsely populated inland area in Lapland, northern Finland, mainly around two large reservoirs flooded 50 years ago. As previous preliminary work revealed elevated Hg levels in this population, we measured Hg exposure along with dietary proxies (δ¹³C and δ¹⁵N) in body feathers collected from white-tailed eagle nestlings in this area between 2008 and 2018. Mercury concentrations were investigated in relation to territory characteristics, proximity to the reservoirs and dietary ecology as potential driving factors of Hg contamination. Mercury concentrations in the nestlings (4.97–31.02 μg g⁻¹ dw) were elevated, compared to earlier reported values in nestlings from the Finnish Baltic coast, and exceeded normal background levels (≤5.00 μg g⁻¹) while remaining below the tentative threshold of elevated risk for Hg exposure mediated health effect (>40.00 μg g⁻¹). The main drivers of Hg contamination were trophic position (proxied by δ¹⁵N), the dietary proportion of the predatory fish pike (Esox lucius), and the vicinity to the Porttipahta reservoir. We also identified a potential evolutionary trap, as increased intake of the preferred prey, pike, increases exposure. All in all, we present results for poorly understood freshwater lake environments and show that more efforts should be dedicated to further unravel potentially complex pathways of Hg exposure to wildlife.

... The enrichment of the biosphere in anthropogenic mercury (Hg) impacts wildlife and humans across the globe. [1][2][3][4] Despite detailed understanding of the environmental conditions that promote the uptake of neurotoxic methylmercury (MeHg) in food webs, key knowledge gaps remain on the internal transformations, redistribution, and toxicologic mechanisms of Hg in higher organisms. Recently, advancement in the application of high energy-resolution X-ray absorption (HR-XANES) spectroscopy in wildlife identified that MeHg is detoxified to nontoxic mercury selenide (HgSe) through an intermediary Hg-tetraselenolate (Hg(Sec)4) species. ...

... 46 We explain the steady state isotopic fractionation of Hg(Sec)4 and HgSe by (1) In plots (d) and €, error bars represent 10% and 90% percentiles and outliers are shown as data points. 3 data. Cys and Sec stand for cysteine and selenocysteine residues within a polymeric chain of peptide or protein, not for free amino acids. ...

... Hg bioaccumulates in organisms and biomagnifies in aquatic food webs largely as the neurotoxin methylmercury (MeHg, CH 3 Hg). Overexposure to Hg, primarily due to consumption of seafood, has major environmental and human health implications with a socio-economic cost estimated to exceed US $5 billion per year 1 . The global Hg problem has worsened significantly due to anthropogenic pollution 2,3 , although there are natural sources of Hg to the environment such as volcanic emissions and weathering of Hg-bearing minerals in rock (for example, cinnabar) 2 . ...

... These differences suggest that variations in subglacial hydrology and/or organic matter content and supply beneath each catchment are important for dMeHg cycling. In addition, a greater fraction of meltwaters may originate from hypoxic/anoxic regions of the glacier bed at RG (hypothesized due to the high CH 4 concentrations 22,29 ) where potential for methylation is higher 1,30 . There may also be an influence of river-marginal inputs in the RG catchment (for example, proglacial lakes and tundra) given that the meltwater river travels along the ice margin for ~9 km before arriving at the sampling site. ...

The Greenland Ice Sheet is currently not accounted for in Arctic mercury budgets, despite large and increasing annual runoff to the ocean and the socio-economic concerns of high mercury levels in Arctic organisms. Here we present concentrations of mercury in meltwaters from three glacial catchments on the southwestern margin of the Greenland Ice Sheet and evaluate the export of mercury to downstream fjords based on samples collected during summer ablation seasons. We show that concentrations of dissolved mercury are among the highest recorded in natural waters and mercury yields from these glacial catchments (521–3,300 mmol km−2 year−1) are two orders of magnitude higher than from Arctic rivers (4–20 mmol km−2 year−1). Fluxes of dissolved mercury from the southwestern region of Greenland are estimated to be globally significant (15.4–212 kmol year−1), accounting for about 10% of the estimated global riverine flux, and include export of bioaccumulating methylmercury (0.31–1.97 kmol year−1). High dissolved mercury concentrations (~20 pM inorganic mercury and ~2 pM methylmercury) were found to persist across salinity gradients of fjords. Mean particulate mercury concentrations were among the highest recorded in the literature (~51,000 pM), and dissolved mercury concentrations in runoff exceed reported surface snow and ice values. These results suggest a geological source of mercury at the ice sheet bed. The high concentrations of mercury and its large export to the downstream fjords have important implications for Arctic ecosystems, highlighting an urgent need to better understand mercury dynamics in ice sheet runoff under global warming. Meltwaters from the southwestern margin of the Greenland Ice Sheet contain exceptionally high concentrations of mercury, exporting up to more than 200 kmol of dissolved mercury every year, suggest mercury measurements from three glacial catchments.

... 2,3 Methylmercury bioaccumulates in aquatic food webs, reaching fish tissue concentrations that can exceed water concentrations by 6 orders of magnitude. 4,5 Because trace concentrations of Hg in water and sediments contribute to MeHg biomagnification, remediation strategies emphasize decreasing the production and bioaccumulation of MeHg in addition to source control. 6 Many remediation methods for Hg have been investigated, including sediment capping, phytoremediation, sorbent amendments, and redox controls. ...

... MeHg is considered to be a highly toxic compound. Much attention has been drawn to MeHg formation in aquatic sediments (Bravo & Cosio, 2020;Fleck et al., 2016;Lehnherr, 2014;Li & Cai, 2013). The inorganic Hg can be biomethylated by the biological species. ...

Traceability and reliable results are the two pillars of analytical methods; certified reference materials (CRMs) meet this requirement. ISO 17034:2016 credentials provide brief information on general requirements for the competence of Reference Material Producers (RMPs). The different types of CRMs have been produced in recent years for chemical analysis in food, water, soil, and sediment matrices in recent years. This review provides a detailed overview of the development of CRMs in the field of marine environment, as matrix CRMs play an important role in the field of environmental monitoring. COMAR database, EVISA database: materials, LGC standards, and JRC catalogs are very helpful online resources to find various types of CRMs according to the application requirements.

... The metal with the greatest bioaccumulation potential is Hg in its methylated form (MeHg), which is released largely under anoxic conditions by sulphur-reducing bacteria (Matilainen, 1995). MeHg also has the potential to biomagnify in food chains, a phenomenon where species at a higher trophic level are accumulating greater concentrations of toxic elements than organisms lower in the food chain (Lehnherr, 2014). There is now growing scientific evidence that exposure to low levels of contaminants in the environment is contributing to society's cancer burden and health hazard (Saaristo et al., 2018). ...

Nicolas Pelletier

Human-released metals are present to some extent in soil and sediment from even the remotest areas, including the Canadian Arctic. The cumulative impact of legacy pollution, ongoing release of contaminants and climate change could lead to important modifications to metal transport and transformation processes in the environment that can affect the exposure of biota and humans to metals. Large uncertainties remain regarding the future transport of metals on the subarctic landscape, including those that can be toxic in low dose like lead and mercury. Paleoecology is a powerful tool to evaluate changes in metal pollution and recovery in lakes by providing long-term records of environmental conditions at a relatively low cost and with rapid analysis. Paleoecological records can help fill important research gaps that current monitoring approaches can't address because of the lack of temporal perspective. In this thesis, the records of multiple environmental archives were analysed and compared to understand the changes in metal accumulation and transport that occurred over the last centuries to millennia in subarctic Canada, from the Yellowknife (Northwest Territories) and the Whitehorse Regions (Yukon). Multiple approaches to times series analysis were developed to evaluate the individual and cumulative impacts of specific sources and processes commonly affecting subarctic boreal lakes. These processes include local point-source emissions, catchment retention and transport of contaminants, and contaminants released by wildfires. This thesis provides quantification for processes that are seldom addressed in the literature so far, especially for subarctic environments. Subarctic lakes will continue receiving anthropogenic metal for years regardless of future emissions because of the impact of catchment retention. Terrestrial heavy metals retained in catchments are susceptible to remobilisation toward aquatic environments by natural processes such as land erosion, permafrost thaw and wildfires; and these processes may be enhanced by climate change. Recovery of any specific site from heavy metal pollution is also dependent on local parameters, explaining the necessity to characterise ecosystem recovery from heavy metal pollution in different types of ecosystems, including subarctic environments.

Estuaries are sinks for mercury, in which the most toxic mercury form, neurotoxic methylmercury (MeHg), is produced by mercury methylators and accumulates in estuarine sediments. In the same area, the microbial sulfur cycle is triggered by sulfate-reducing bacteria (SRB), which is considered as the main mercury methylator. In this review, we analyzed the sulfur and mercury speciation in sediments from 70 estuaries globally. Abundant mercury and sulfur species were found in the global estuarine sediments. Up to 727 μg THg/g dw and 880 ng MeHg/g dw were found in estuarine sediments, showing the serious risk of mercury to aquatic ecological systems. Significant correlations between sulfur and MeHg concentrations were discovered. Especially, the porewater sulfate concentration positively correlated to MeHg production. The sulfur cycle affects MeHg formation via activating mercury methylator activities and limiting mercury bioavailability, leading to promote or inhibit MeHg formation at different sulfur speciation concentrations. These results suggest that sulfur biogeochemical cycle plays an important role in mercury methylation in estuarine sediments, and the effect of the sulfur cycle on mercury methylation deserves to be further explored in future research.

Zhike Li
Jie Chi
Zhenyu Wu
Yindong Tong

The bioaccumulation of mercury (Hg) in aquatic ecosystem poses a potential health risk to human being and aquatic organism. Bioaccumulations by plankton represent a crucial process of Hg transfer from water to aquatic food chain. However, the current understanding of major factors affecting Hg accumulation by plankton is inadequate. In this study, a data set of 89 aquatic ecosystems worldwide, including inland water, nearshore water and open sea, was established. Key factors influencing plankton Hg bioaccumulation (i.e., plankton species, cell sizes and biomasses) were discussed. The results indicated that total Hg (THg) and methylmercury (MeHg) concentrations in plankton in inland waters were significantly higher than those in nearshore waters and open seas. Bioaccumulation factors for the logarithm of THg and MeHg of phytoplankton were 2.4–6.0 and 2.6–6.7 L/kg, respectively, in all aquatic ecosystems. They could be further biomagnified by a factor of 2.1–15.1 and 5.3–28.2 from phytoplankton to zooplankton. Higher MeHg concentrations were observed with the increases of cell size for both phyto- and zooplankton. A contrasting trend was observed between the plankton biomasses and BAFMeHg, with a positive relationship for zooplankton and a negative relationship for phytoplankton. Plankton physiologic traits impose constraints on the rates of nutrients and contaminants obtaining process from water. Nowadays, many aquatic ecosystems are facing rapid shifts in nutrient compositions. We suggested that these potential influences on the growth and composition of plankton should be incorporated in future aquatic Hg modeling and ecological risk assessments.

Methylmercury (MeHg) forms in anoxic environments and can bioaccumulate and biomagnify in aquatic food webs to concentrations of concern for human and wildlife health. Mercury (Hg) pollution in the Arctic environment may worsen as these areas warm and Hg, currently locked in permafrost soils, is remobilized. One of the main concerns is the development of Hg methylation hotspots in the terrestrial environment due to thermokarst formation. The extent to which net methylation of Hg is enhanced upon thaw is, however, largely unknown. Here, we have studied the formation of Hg methylation hotspots using existing thaw gradients at five Fennoscandian permafrost peatland sites. Total Hg (HgT) and MeHg concentrations were analyzed in 178 soil samples from 14 peat cores. We observed 10 times higher concentrations of MeHg and 13 times higher %MeHg in the collapse fen (representing thawed conditions) as compared to the peat plateau (representing frozen conditions). This suggests significantly greater net methylation of Hg when thermokarst wetlands are formed. In addition, we report HgT to soil organic carbon ratios representative of Fennoscandian permafrost peatlands (median and interquartile range of 0.09 ± 0.07 μg HgT g-1 C) that are of value for future estimates of circumpolar HgT stocks.

Christopher D. Holmes
Daniel J. Jacob
E. S. Corbitt
Franz Slemr

Global models of atmospheric mercury generally assume that OH and ozone are the main oxidants converting Hg0 to HgII and thus driving mercury deposition to ecosystems. However, thermodynamic considerations argue against the importance of these reactions. We demonstrate here the viability of atomic bromine (Br) as an alternative Hg0 oxidant. We conduct a global 3-D simulation with the GEOS-Chem model assuming Br to be the sole Hg0 oxidant (Hg + Br model) and compare to the previous version of the model with OH and ozone as the sole oxidants (Hg + OH/O3 model). We specify global 3-D Br concentration fields based on our best understanding of tropospheric and stratospheric Br chemistry. In both the Hg + Br and Hg + OH/O3 models, we add an aqueous photochemical reduction of HgII in cloud to impose a tropospheric lifetime for mercury of 6.5 months against deposition, as needed to reconcile observed total gaseous mercury (TGM) concentrations with current estimates of anthropogenic emissions. This added reduction would not be necessary in the Hg + Br model if we adjusted the Br oxidation kinetics downward within their range of uncertainty. We find that the Hg + Br and Hg + OH/O3 models are equally capable of reproducing the spatial distribution of TGM and its seasonal cycle at northern mid-latitudes. The Hg + Br model shows a steeper decline of TGM concentrations from the tropics to southern mid-latitudes. Only the Hg + Br model can reproduce the springtime depletion and summer rebound of TGM observed at polar sites; the snowpack component of GEOS-Chem suggests that 40% of HgII deposited to snow in the Arctic is transferred to the ocean and land reservoirs, amounting to a net deposition flux of 60 Mg a−1. Summertime events of depleted Hg0 at Antarctic sites due to subsidence are much better simulated by the Hg + Br model. Model comparisons to observed wet deposition fluxes of mercury in the US and Europe show general consistency but the Hg + Br model is unable to capture the summer maximum over the southeast US because of low subtropical Br concentrations. Vertical profiles measured from aircraft show a decline of Hg0 above the tropopause that can be captured by both the Hg + Br and Hg + OH/O3 models, except in Arctic spring where the observed decline is much steeper than simulated by either model; we speculate that oxidation by Cl species might be responsible. The Hg + Br and Hg + OH/O3 models yield similar global budgets for the cycling of mercury between the atmosphere and surface reservoirs, but the Hg + Br model results in much larger fraction of mercury deposited to the Southern Hemisphere oceans.

This paper provides an up-to-date assessment of global mercury emissions from anthropogenic and natural sources. On an annual basis, natural sources account for 5207 Mg of mercury released to the global atmosphere, including the contribution from re-emission processes, which are emissions of previously deposited mercury originating from anthropogenic and natural sources, and primary emissions from natural reservoirs. Anthropogenic sources, which include a large number of industrial point sources, are estimated to account for 2320 Mg of mercury emitted annually. The major contributions are from fossil-fuel fired power plants (810 Mg yr−1), artisanal small scale gold mining (400 Mg yr−1), non-ferrous metals manufacturing (310 Mg yr−1), cement production (236 Mg yr−1), waste disposal (187 Mg yr−1) and caustic soda production (163 Mg yr−1). Therefore, our current estimate of global mercury emissions suggests that the overall contribution from natural sources (primary emissions + re-emissions) and anthropogenic sources is nearly 7527 Mg per year, the uncertainty associated with these estimates are related to the typology of emission sources and source regions.

D. Mcalpine
S. Araki

Minamata is a small industrial town situated near the southwest coast of Kyushu, the most southerly of the three main islands of Japan. A number of villages are located on or near a neighboring bay. Into this bay flows the effluent from a large fertilizer factory. Between 1953 and 1956 a mysterious nervous disease affected the fishing community living near the bay. The outbreak was investigated by a number of departments of Kumamoto University, and the results were published in 1957, in Japanese in two supplements of the Journal of the Kumamoto Medical Society. The following brief account is taken from a recent article, summarizing the Japanese accounts of the outbreak (McAlpine and Araki¹). Between November, 1953, and December, 1956, a total of 56 persons living on or near Minamata Bay were affected by a neurological illness characterized by an acute or a subacute onset of numbness and cerebellar signs.

We present a detailed study of the biogeochemical factors controlling mercury (Hg) distribution, methylmercury (MeHg) production, and MeHg efflux in sediments of the mid-Atlantic continental shelf and slope. The mildly reduced surface sediments of the shelf and slope provide ideal conditions for MeHg production. They are sufficiently reduced to support microbial sulfate reduction, but contain very low dissolved sulfide concentrations. The redox zonation of sediments determined the depth distribution of MeHg production, whereas the bioavailability of inorganic Hg for methylation appeared to be the dominant driver of spatial patterns across the shelf and slope. Sediment total Hg concentrations were well predicted by sediment organic matter (SOM) content, with the highest concentrations of Hg and MeHg in the fine-grained organic clays of the slope. However, SOM-normalized Hg concentrations decreased with distance from shore. The changing character of organic matter with distance from shore appeared to affect Hg partitioning and bioavailability for methylation. The percentage of Hg in sediments as MeHg was well predicted by measured methylation rates, but not by demethylation rates. On the basis of measured concentrations in bottom waters and surficial pore waters, the average diffusive efflux of Hg(II) and MeHg from sediments to coastal waters was estimated to be 26 and 0.8 pmol m -2 d -1, respectively. Extrapolated globally, the diffusive input of MeHg from shelf and slope sediments is estimated to be 0.01 Mmol per year. As the actual fluxes can be substantially higher than diffusive fluxes, we suggest that shelf and upper slope sediments are a major source of MeHg to the coastal ocean. © 2010, by the American Society of Limnology and Oceanography, Inc.

Britta Culbertson Nature Conservancy What's in a Word

Source: https://www.researchgate.net/publication/260105766_Methylmercury_Biogeochemistry_A_review_with_Special_Reference_to_Arctic_Aquatic_Ecosystems

Goethe Pheine1949

Britta Culbertson Nature Conservancy What's in a Word

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