Critical review of mercury fates and contamination in the arctic tundra ecosystem

S C I E N C E O F T H E T O T A L E N V I R O N M E N T 4 0 0 ( 2 0 0 8 ) 1 7 3 – 2 1 1

ava i l ab l e a t www.sc i enced i rec t . com

www.e l sev i e r. com/ loca te / sc i to tenv

Critical review of mercury fates and contamination in the arctictundra ecosystem

Laurier Poissanta,⁎, Hong H. Zhanga, João Canáriob, Philippe Constanta

aEnvironment Canada, Science and Technology branch 105 McGill St. Montréal, Québec, Canada H2Y 2E7bNational Institute of Biological Resources (INRB/L-IPIMAR) Av. Brasilia, 1499-006 Lisbon, Portugal

A R T I C L E I N F O

⁎ Corresponding author. Tel.: +1 514 283 1140E-mail address: [email protected] (

0048-9697/$ – see front matter. Crown Copyrdoi:10.1016/j.scitotenv.2008.06.050

A B S T R A C T

Article history:Received 25 March 2008Received in revised form27 June 2008Accepted 27 June 2008Available online 15 August 2008

Mercury (Hg) contamination in tundra region has raised substantial concerns, especiallysince the first report of atmospheric mercury depletion events (AMDEs) in the Polar Regions.During the past decade, steady progress has been made in the research of Hg cycling in thePolar Regions. This has generated a unique opportunity to survey the whole Arctic in respectto Hg issue and to find out new discoveries. However, there are still considerable knowledgegaps and debates on the fate of Hg in the Arctic and Antarctica, especially regarding theimportance and significance of AMDEs vs. net Hg loadings and other processes that burdenHg in the Arctic. Some studies argued that climate warming since the last century hasexerted profound effects on the limnology of High Arctic lakes, including substantialincreases in autochthonous primary productivity which increased in sedimentary Hg,whereas some others pointed out the importance of the formation and postdepositioncrystallographic history of the snow and ice crystals in determining the fate andconcentration of mercury in the cryosphere in addition to AMDEs. Is mercury re-emittedback to the atmosphere after AMDEs? Is Hg methylation effective in the Arctic tundra?Where the sources of MeHg are? What is its fate? Is this stimulated by human made? Thispaper presents a critical review about the fate of Hg in the Arctic tundra, such as pathwaysand process of Hg delivery into the Arctic ecosystem; Hg concentrations in freshwater andmarine ecosystems; Hg concentrations in terrestrial biota; trophic transfer of Hg andbioaccumulation of Hg through food chain. This critical review of mercury fates andcontamination in the Arctic tundra ecosystem is assessing the impacts and potential risks ofHg contamination on the health of Arctic people and the global northern environment byhighlighting and “perspectiving” the various mercury processes and concentrations foundin the Arctic tundra.

Crown Copyright © 2008 Published by Elsevier B.V. All rights reserved.

Keywords:MercuryArcticAirSoilWaterBiotaVegetation

1. Introduction

Themainland of the Arctic is dominated by tundra. The Arctictundra is located in the northern hemisphere (Greenland,Alaska, Canada, Europe, and Russia), encircling the North Pole(Fig. 1) and extending south to the coniferous forests of thetaiga. Arctic tundra covers about one-tenth of the total surfaceof the Earth. The total human population in the whole Arctic

; fax: +1 514 283 8869.L. Poissant).

ight © 2008 Published by

currently exceeds 3.5 million; most of them are living in thetundra ecosystem.

Purely traditional life-styles are tending to disappear asindigenous people adopt “southern” life-styles to a greater orlesser extent. Nevetheless, indigenous peoples of the northstill have a large traditionally survival on a sustainable sys-tem of hunting, fishing, herding the local fauna and floracollection.

Elsevier B.V. All rights reserved.

mailto:[email protected]

http://dx.doi.org/10.1016/j.scitotenv.2008.06.050

Fig. 1 –The transport of Hg into the Arctic region through main physical pathways (wind, rivers and ocean currents) (adaptedfrom Macdonald et al., 2005).

174 S C I E N C E O F T H E T O T A L E N V I R O N M E N T 4 0 0 ( 2 0 0 8 ) 1 7 3 – 2 1 1

Because of the harsh climatic conditions on land, thebiodiversity of the Arctic tundra region is relatively poor. Thelow productivity of the terrestrial ecosystemsmakes the Arctictundra particularly sensitive to land degradation and erosion.By comparison, the marine environment is far more produc-tive. Ice edges and associated waters are key areas ofproductivity in all regions of the Arctic (UNEP, 1997). Thismatter of fact is mainly explaining the sustainable geographi-cal extension of indigeneous peoples along the tundra marinecoasts.

On the other hand, fauna associated with the ice edge forman important pathway for contaminants, especially mercury(Hg) to enter the food web between primary producers andfish, sea birds, and mammals (UNEP, 1997). Whereas, inlandtundra might be further exposed to local (e.g., mining) andatmospheric long-range transport contamination.

During the past two decades, steady progress has beenmade in the research of contaminants cycling in the PolarRegions. For example, in 1991, the Northern ContaminantsProgram (NCP) was established by the Canadian researchers inresponse to concerns about human exposure to elevatedlevels of contaminants in fish and wildlife species that areimportant in the traditional diets of northern Aboriginalpeoples. At the same year, an international organization:Arctic Monitoring and Assessment Programme (AMAP), was

established as well to implement components of the ArcticEnvironmental Protection Strategy (AEPS). These programshelped to better understand the fate of contaminants in theArctic regions.

However, there are still considerable knowledge gaps onthe pathways and processes delivery of contaminants toArctic environments and that issue has been the subject ofincreased scientific investigation (e.g., ArcticNet; InternationalPolar Year).

Accordingly, mercury (Hg) contamination in tundra regionhas raised substantial concerns, especially since the firstreport of atmospheric mercury depletion events (AMDEs) inthe Arctic region (Schroeder et al., 1998). Indeed, the Arctic isbelieved as an important global sink for atmospheric Hg,especially during the AMDEs that are occurring in spring time(Schroeder et al., 1998; Lindberg et al., 2002; Poissant et al.,2002; Ariya et al., 2004).

During AMDEs, deposited Hg may enter into arctic foodchains, resulting in elevated Hg levels in ecosystem.Moreover,some evidences have been accumulated during the pastdecade to indicate that global change may alter the exposurerisks to contaminants, such Hg, delivered in the Arctic (Figs. 1and 2).

The main purpose of this study is to critical review themercury fates and contamination in the arctic tundra

Fig. 2 –Schematic pathways for the delivery and accumulation of mercury to Arctic ecosystems (Adapted from Lindberg et al.,2002; Fisk et al., 2003; Steffen et al., 2007).

175S C I E N C E O F T H E T O T A L E N V I R O N M E N T 4 0 0 ( 2 0 0 8 ) 1 7 3 – 2 1 1

ecosystems and to assess the impacts and risks of Hgcontamination on the health of Arctic people and the globalnorthern environment.

2. Pathways and process of Hg delivery intothe Arctic ecosystem

Pathways of Hg entering to the Arctic region include itstransport in the troposphere under gas phase and particulateforms, as well as via ocean currents and rivers (Fig. 1)(Macdonald et al., 2000). Dry deposition and precipitationscavenging of gaseous and particulate phases of Hg are alsosignificant pathways of Hg input into the ecosystem (Poissantet al., 2002; Braune et al., 2005).

It is thought that the primary mechanism for atmosphericelemental Hg removal in Polar Regions is by atmosphericmercury depletion events (AMDEs) during polar sunrise(Schroeder et al., 1998; Lindberg et al., 2002). During AMDEs,

gaseous elemental mercury is oxidized to reactive gaseousmercury, is lost from the atmosphere, and results in largeseasonal fluxes of Hg onto snow surface (e.g., Poissant, 2000;St. Louis et al., 2005; Constant et al., 2007; Steffen et al., 2007). Aunique series of photochemical reactions between ozone andhalogens during polar spring create reactive halogen speciessuch Br radicals that oxidize gaseous elementalmercury (GEM)to reactive gaseous mercury (RGM) (e.g., Lindberg et al., 2002;Hoenninger, 2002; Poissant and Hoenninger, 2004) (Fig. 2).

There is great debate whether springtime AMDEs increasethe Hg concentration in surface snow and how long this inputmay be preserved (e.g., Poissant, 2000; St. Louis et al., 2005;Constant et al., 2007). Although snowmay act as a sink duringAMDEs, Hg can be re-emitted back to the atmosphere and actas a source to the atmosphere. It was shown that Hg is highlymobile in snow, particularly in exposed snow to solar radia-tion (Lalonde et al., 2002).

Also, an interesting question that remains unclear yet:Why AMDEs do not appear in Polar fall season. Recently,


Poissant et al., 2006 did measurements of speciated mercuryspecies (GEM, total gaseous mercury (TGM), RGM, particulatemercury (Hgp)) onboard the Amundsen icebreaker in 2005 inthe whole Canadian Arctic during ArcticNet 2005's missionand did not observe active marine boundary layers in respectto AMDEs (August to October).

Several reports have demonstrated that anthropogenic Hgemissions have considerably increased the Hg burden of theArctic biota. For instance, analyses of Hg concentrations insediment cores (Lockart et al., 1998; Fitzgerald et al., 2005) andanimal tissues collections such as prey bird feathers (Dietzet al., 2006a), polar bear hairs (Dietz et al., 2006b) and belugateeth (Outridge et al., 2002) revealed that industrialization hadcaused a pulse in Hg concentrations. Recently, Outridge et al.,2007 published a “new direction paper” giving the evidence forcontrol of Hg accumulation rates in Canadian High Arctic Lakesediments by variations of aquatic primary productivity. Theauthors claimed that climate warming since the last centuryhas exerted profound effects on the limnology of High Arcticlakes, including substantial increases in autochthonousprimary productivity (APP). This finding suggests that theatmospheric contribution of long-range anthropogenic Hg toHigh Arctic lakes may have been overestimated by several-fold because of this climate-driven process.

Recently, Douglas et al., 2008 highlighted the importance ofthe formation and postdeposition crytollographic history ofthe snow and ice crystals in determining the fate andconcentration of mercury in the cryosphere in addition toAMDEs. On the basis of many snow and ice samples collectedin the Alaskan Arctic, they suggested that kinetic crystalsgrowing from vapor phase, including surface hoar, frostflowers, and diamond dust, yield mercury concentrationsthat are typically 2–10 times higher than that reported forsnow deposited during AMDEs. Interestingly, the latterprocesses should be effective in Arctic fall season as well,but no data was presented in that respect.

2.1. Atmospheric Hg concentrations

More than 98% of atmospheric Hg is generally under the formof gaseous elemental mercury (GEM; Hg0), the remaining beingreactive and present either in the gaseous form (reactivegaseous mercury; RGM) or adsorbed onto airborne particles(particulatemercury; pHg) (e.g., Berg et al., 2001; Poissant et al.,2005). While GEM has a long atmospheric residence time,reactive Hg species are rapidly scavenged from the atmo-sphere by particles and precipitations. Chemical reactionsinvolved in redox transformations of atmospheric Hg are thencrucial to control the fate of this global pollutant. Develop-ment of automated instrumentation for continuous TGMmonitoring enabled the establishment of extensive datasetassessing the temporal and spatial variation of atmosphericHg concentrations. TGM atmospheric concentrations hadincreased between the 1970 and 1980, reaching a peak valuein the 1980's (Slemr et al., 2003). Reduction of the global Hgemissions (Pacyna et al., 2006) was observed and the lowestTGM concentrations were recorded in 1996. Since then, TGMlevels remain steady at 1.5–1.7 ng/m3 and 1.2–1.3 ng/m3 in thenorthern and the southern hemispheres, respectively (Slemret al., 2003; Kim et al., 2005; Steffen et al., 2005;Wängberg et al.,

2007). Interestingly, the analysis of long-term TGM atmo-spheric concentrations measured at Alert (Canada) suggestedan increasing retention of Hg in the Arctic (Steffen et al., 2005).

Multi-year observations (7–11 years) of TGM along a Southto North transect (parallels: 45°N; 50°N; 55°N) in Québec(Canada) (Poissant et al., in preparation) showed that southernTGM concentrations at monitoring sites in St. Anicet (45°N)and Mingan (50°N) are decreasing whereas the northern site(Kuujjuarapik (55°N)) showed stable even a small increase ofTGM concentrations over time. The latter information issuggesting that southern and northern sites are decoupled.N–E America Hg policy reduction strategies seem to beeffective to reduce TGM concentrations in southern locationsin Québec whereas the northern region seem to be stillimpacted by Hg long-range transport from Eurasia, especiallyin winter time.

Discovery of AMDEs in Polar Regions changed perspectivesabout GEM atmospheric lifetime and chemistry. Duringsunrise at the polar region, the autocatalytic release of seasalt aerosols is changing the oxidative photochemistry of thestratified planetary boundary layer (Fig. 2). Reactive halogensare oxidizing GEM to RGM resulting in a reduction of GEMconcentrations from 1.6–1.8 to ~0 ng/m3. During theseepisodes, atmospheric Hg speciation is then dominated byshort-lived reactive Hg species. At Barrow (71°19'N; 156°37'W),AMDEs are typically accompanied by an increase of RGMconcentration from 0.000–0.002 ng/m3 to 0.300–0.900 ng/m3,with no significant pHg enrichments (Lindberg et al., 2001;Skov et al., 2004; Brooks et al., 2006). At Kuujjuarapik (55°17′N;77°46′W), both RGM and pHg increased at the onset of AMDEs,reaching concentrations as high as 0.190 and 0.600 ng/m3,respectively (Poissant and Hoenninger, 2004). Accordingly,RGM and pHg concentrations of 0.200 and 0.160 ng/m3 havebeen detected at Ny-Ålesund, Svalbard (78°54′N; 11°53′W)during AMDEs (Berg et al., 2003; Aspmo et al., 2005). Finally,extreme RGM and pHg concentrations of 1.2 and 2.3 ng/m3

have been reported during AMDEs recorded at Churchill(58°45'N; 94°4'W) (Kirk et al., 2006). Opposite to GEM whichhas a deposition velocity b0.2 cm/s, RGM and pHg aredeposited at rates N1.0 cm/s (Skov et al., 2006; Poissant et al.,2004). AMDEs are then a significant pathway involving a directtransfer of atmospheric Hg to the arctic biota.

2.2. AMDEs: dry deposition of bioavailable Hg

Ecotoxicological potential of the AMDEs has been assessed byusing Vibrio anguillarum pRB28, amer-lux bioreporter. Since thelatter strain has been transformedwith a plasmid encoding forthe luxCDABE operon under the control of the regulationregion of themer operon, it is emitting a bioluminescent signalin the presence of bioavailable Hg (Selifonova et al., 1993). InAlaska,mer-lux bioreporter positive signals were detected onlyfor snow samples collected following AMDEs, and bioavailableHg accounted for 13–55% of the total Hg (Scott, 2001; Lindberget al., 2002). Since reactive was formed during AMDEs, it wasready to deposit due to its short atmospheric lifetime asdiscussed above. The deposited reactive Hg contributed to theincreasing Hg levels at surface. A clear pattern of total Hgconcentration gradients had been observed for surface snowsamples collected around polynya and sea ice leads; the sites


where chemical reactions initiating AMDEs are taking place(Bargagli et al., 2005; Douglas et al., 2005). For instance, total Hgconcentrations of 820 ng/L have been detected in snow andfrost flowers collected along large open water sea ice leads ofthe Arctic Ocean (Douglas et al., 2005). This is an extreme highconcentration, which is one order of magnitude larger thanconcentrations detected in snow samples collected at 20 km ofthe open water sea ice leads (Douglas et al., 2005). The sameconcentration distribution pattern has been observed atKuujjuarapik (55°17′N; 77°46′W) where total Hg snow concen-trations decreased as a function of the distance from theHudson Bay coast (Constant et al., 2007). At Ellesmere Island(82°29′N; 62°20′W), total Hg concentrations of surface snowsamples were 5–22 ng/L before the polar sunrise and increasedto 121–182 ng/L following AMDEs (Steffen et al., 2002). Thesame trend has been observed at Barrow (71°19′N; 156°37′W)where total Hg snow concentrations increased form thetypical background level of 1 ng/L to 90 ng/L followingAMDEs (Lindberg et al., 2002; Brooks et al., 2006). AMDEsrecorder at Ny-Ålesund, Svalbard (78°54′N; 11°53′W) increasedalso total Hg snow concentration from the backgroundconcentration of 1 ng/L to levels of 30 ng/L (Berg et al., 2003).Finally, AMDEs caused Hg enrichments to surface snowsamples collected in sub-arctic ecosystems such as Kuujjuar-apik and Churchill (58°45'N; 94°4'W) where total Hg concen-trations increased from 1–5 ng/L to 60–80 ng/L followingAMDEs (Poissant, 2000; Kirk et al., 2006). At several occasions,the occurrence of AMDEs was not accompanied by any Hgsnow enrichments (Ferrari et al., 2005; Constant et al., 2007).Analysis of air mass backward trajectories showed that theabsence of Hg snow enrichments following AMDEs werecaused by influx of air masses already depleted in Hg ratherthan from local chemical reactions (Gauchard et al., 2005a,b;Constant et al., 2007).

It is estimated that 90 to 450metric tons of Hg are depositedannually in the Arctic due to AMDEs (Ariya et al., 2004; Skov etal., 2004). This deposition term is generally well accepted, butthe fate and processes of deposited Hg are strongly debated(Lindberg et al., 2002; Outridge et al., 2007; Douglas et al., 2008).

As discussed above Douglas et al. (2008) pointed out theimportance of the formation and postdeposition crytollo-graphic history of the snow and ice crystals in determining thefate and concentration of mercury in the cryosphere inaddition to AMDEs.

Lahoutifard et al. (2006) demonstrated that solar irradiationof acidic snow (pH 4–5) containing the photochemical oxidanthydrogen peroxide (H2O2) results in Hg deposition.

Even if AMDEs are often followed by snowpack GEMemissions and total Hg snow concentrations diminution(Poissant, 2000; Steffen et al., 2002; Lahoutifard et al., 2005;St. Louis et al., 2005; Kirk et al., 2006; Poulain et al., 2007a),there was evidence that AMDEswere causing a net input of Hgin the Arctic ecosystems. As observed for spatial distributionof total Hg surface snow concentrations, total Hg concentra-tion in the snowmelt water are spatially heterogeneous. Forsnowmelt water collected in ponds formed on ocean ice or atEllesmere Island, total Hg concentrations varied between 1and 10 ng/L (Lahoutifard et al., 2005; Aspmo et al., 2006). On theother hand, snowmelt water total Hg concentrations werebetween 10 and 30 ng/L at Kuujjuarapik and Barrow (Lindberg

et al., 2002; Dommergue et al., 2003a). Actually, more data areneeded to investigate the spatial variation of total Hgconcentrations in snow and snowmelt water under differentlandscapes. Considering that tree shading is diminishing thephotoreduction of Hg within the snow cover of mid-latitudeecosystems (Poulain et al., 2007c), significant differenceshould be observed for total Hg snow concentrations andspringtime Hg discharges along the ecosystem boundaries ofthe sub-Arctic and Arctic ecosystems. Even if snow is oftenutilized as an indicator of Hg deposition during AMDEs, anexamination of Hg concentrations in vegetation has revealedinteresting insights about the impact of AMDEs on thepolar biota. Indeed, direct transfer of deposited Hg has beenobserved in Antarctica where lichen and moss samplescollected near the coast facing a polynya contained highertotal Hg concentrations than the background level (Bargagliet al., 2005).

Asawhole,more than95%of theHg is scavenged in thesnowgrains,while less than5% isunder the formofGEM in interstitialair of the snowpack (Dommergue et al., 2003a; Ferrari et al.,2004). GEMmay reach high levels within the surface interstitialair due to photoreduction of reactive Hg species (Steffen et al.,2002; Lalonde et al., 2002), but GEM concentrations are decreas-ing as a function of the depth in the snowpack, suggesting theoccurrence of GEM oxidation reactions within the snow cover(Dommergue et al., 2003b). The exact mechanisms involved inGEM oxidation have not yet been identified but should involveoxidative agents such as halogen radicals (Ferrari et al., 2004) orH2O2 (Lahoutifard et al., 2006).

2.3. The fate of MeHg within the snow cover

Since AMDEs add bioavailable Hg to the snow cover of PolarRegions, investigations have been conducted to detect thepresenceofMeHg in the snowpack. Indeed, theanalysis ofMeHgconcentration in snowmelt water discharges on EllesmereIsland revealed that snowmelt water was the most importantsource ofMeHg for theArctic ecosystems (Loseto et al., 2004a,b).Further investigations showed that MeHg detected in the snowrepresented up to 7.5% of the total Hg (Ferrari et al., 2004;Lahoutifard et al., 2005; St. Louis et al., 2005; Constant et al.,2007). On Cornwallis Bay (74°42′N; 94°58′W), snow MeHgconcentrations were between b10 and 140 pg/L (Lahoutifardet al., 2005), while levels up to 280 pg/L have been reported forsamples collected on Ellesmere Island (St. Louis et al., 2005,2007). At Station Nord, Greenland (81°36′N; 16°40′W), MeHgconcentrations between b10 and 113 pg/L has been detected inthe snow cover (Ferrari et al., 2004). Interestingly, snow samplescollected at Kuujjuarapik (55°17′N; 77°46′W) revealed a signifi-cant riseofMeHgconcentrationduring thesnowmeltingperiod,with the detection of concentrations as high as 700 pg/L(Constant et al., 2007).

There are currently two pathways proposed to explain thefate of MeHg in the snow cover: transport of MeHgwithmarineaerosols or a production of MeHg within the snow cover.Marine dimethylmercury (DiMeHg) biogenic emissions werepostulated to be a significant source of MeHg for the snowcovers of the Arctic and sub-arctic ecosystems. Indeed, MeHgand chlorine snow concentrations were significantly corre-lated on Ellesmere Island (St. Louis et al., 2005; St. Louis et al.,


2007), Corwallis Island (Lahoutifard et al., 2005) and Kujjuar-apik (Constant et al., 2007). DiMeHg is insoluble and thenemitted to the atmosphere before being oxidized to MeHg (Linand Pehkonen, 1999) which is transported before its deposi-tion onto the snow cover. DiMeHg was detected in open waterof sea ice leads formed in the Canadian High Arctic, and thedetected concentrations corresponded to an emission factor of4.8±0.6 ng/m2d (St. Louis et al., 2007). Laboratory studies haveshown that some algae produce MeHg and DiMeHg (Pongratzand Heumann, 1998). DiMeHg production and fluxes frompolynya is then presently attributed to the presence of algaebeneath the ice cover of the sea and the presence of anaerobicwater masses sustaining potential microbial-mediatedDiMeHg and MeHg production (St. Louis et al., 2007). Even ifmarine aerosols are potential sources of MeHg, experimentsperformed at Kuujjuarapik suggested the existence of otherpathways to explain the variation of MeHg snow concentra-tions (Constant et al., 2007). During the latter study, nocorrelation was observed between chlorine and MeHg snowconcentrations for the snow samples collected during thesnow melting period. Correlation between MeHg snow con-centrations and culturable bacteria or snow particles, as wellas the increasing trends of [MeHg]/[THg] ratio during the snowmelting period suggested the occurrence of an active methy-lation process within the snow cover. However, further worksare needed to identify the exact nature of the biotic/abioticMeHg production in the melting snow cover.

2.4. Hg in precipitations

Previous investigations have shown the link between Hg wetdeposition and Hg concentrations detected in the aquaticbiota (Downs et al., 1998). An interesting case study has beenrealized in an ecosystem base experiment (ExperimentalLakes Area) involving amendments of different labeled Hgspecies in a lake and its watershed. Less than one monthfollowing the amendments, labeled Hg added in the lake(202Hg) was transferred to the zooplankton and fish, while notraces of labeled Hg deposited in the watershed (198Hg; 200Hg)were transferred to the lake biota (Harris et al., 2007). Thisobservation suggests that deposited Hg is poorly mobile: it isretained to the biota exposed to the deposition. Contrary to theatmospheric TGM concentrations, there are only few pub-lished data concerning the long-term monitoring of Hg wetdeposition in the Arctic. Most of the data are from northernEurope, where a decreasing trend has been reported. Indeed,at different stations located along the North Sea, total Hgconcentrations in precipitations varied between 10 and 25 ng/L in 1995–2002, while a decrease of 30% was recorded betweenthe periods of 1995–98 and 1999–2000 (Wängberg et al., 2007).Accordingly, total Hg concentrations of precipitations col-lected at Svalbard (Ny-Ålesund) and Pallas (Finland) were inthe range of 4–52 ng/L in 1996–97 (Berg et al., 2001). Inprecipitations, proportion of Hg under the organic formrepresents typically 1 to 5% of total Hg. At Mace Head, MeHgconcentrations between 75 and 98 pg/L have beenmeasured inrain precipitations collected in fall (Ebinghaus et al., 1999).These concentrations are in the same order of the 10–179 pg/LMeHg detected in precipitations collected at ExperimentalLakes Area (St. Louis et al., 1995).

2.5. The Hg micro biogeochemistry in snow and ice

Actual knowledge about the role of microorganisms in Hgcycling in the Arctic is very limited (Barkay and Poulain, 2007).In temperate environments, some microorganisms areinvolved in the reduction of Hg(II) to Hg0 and in MeHgdemethylation to detoxify their environments, while fortui-tous metabolism of other microbes is generating organicforms of Hg. Following AMDEs, bioavailable Hg is detected inthe snow cover and elevated Hg concentrations are alsopresent at the interface of the snow and the sea ice. It has beenproposed that these environments promoted the expressionof bacterial merA genes detected in benthic and epiphyticbiofilms samples collected on Cornwallis Island (Poulain et al.,2007b). This detection of merA transcripts suggest thatmicrobial-mediated Hg reduction is occurring in the Arcticbut further investigations are needed to establish the impor-tance of these reactions in the Hg transfer between theinterfaces.

Sulfate-reducing bacteria (SRB) are considered as the mainmethylating agents of the anoxic sediments of temperatezones. In fact, the addition of molybdate in sediments isknown to cause a significant diminution of the Hg methyla-tion activity, and several SRB have demonstrated the ability toproduce MeHg in pure culture (Compeau and Bartha, 1985).Recent investigations have however shown that iron(III)-respiring bacteria are also involved in MeHg production inanoxic sediments (Fleming et al., 2006; Kerin et al., 2006). Thisfinding is interesting and shows the importance of futurestudies dedicated to investigate the existence of other MeHgproducers in the environment. Indeed, even if physiologicallydiverse groups of bacteria, algae, fungi and yeasts havedemonstrated the ability to produce MeHg in pure culture(Vonk and Sijpesteijn, 1973; Yannai et al., 1991; Pongratz andHeumann, 1998), their role in the environment has never beenextensively investigated.

SRB have been detected in Arctic and sub-arctic ecosys-tems, but there was no correlation between their detectionand the MeHg concentrations (Loseto et al., 2004b; Constantet al., 2007). On another hand, recent investigations haverevealed the important diversity of the microorganisms thatare metabolically active within the ice and the snow coverof the Arctic ecosystems. Algae, cyanobacteria, α-, β-, γ-Pro-teobacteria as well as archaea have been detected in theseextreme environments (Vincent et al., 2000; Groudieva et al.,2004). In the ice, complex networks of saline water sustain theactivity of microorganisms that are mainly attached toparticles (Junge et al., 2004). Culture-dependant and culture-independent approaches revealed that microbial biomassexpressed by volume of melted snow is increasing duringthe snowmelt period, but further works are needed toinvestigate if growth or atmospheric deposition is responsibleof that trend (Segawa et al., 2005; Amato et al., 2007; Constantet al., 2007). Interestingly, analyses of microbial diversityrevealed the importance of the Cytophaga-Flavobacteria-Bacter-oides in ice and snow (Junge et al., 2002; Junge et al., 2004;Amato et al., 2007). These microorganisms are well-known fortheir ability to degrade complex polymers such as celluloseand chitine and tolerate extreme temperatures by synthesiz-ing exopolysaccharides favoring their agglomeration onto


particles (Kirchmann, 2002). So far, the interaction of thesemicroorganisms with Hg has not been documented, but giventhe co localization of microbes and Hg onto particles, furtherworks in that field are needed.

3. Hg concentrations in freshwater ecosystem

3.1. Hg levels in lakes and sediments

Hg concentrations in freshwater of the Arctic and sub-Arcticregion have been reported in several studies (Table 1). Semkinet al. (2005) investigated Hg levels in a high Arctic watershed(Amituk lake) located on Cornwallis Island (Nunavut, Canada).Total Hg concentrations in inflow and outflow streams were0.76 ng/L and 0.52 ng/L, respectively. Concentrations of Hg instream water were relatively stable and varied within a smallrange: 0.23–0.33 ng/L. Total Hg concentrations in the Amituklake were consistent with the data reported in Russia LakeBaikal (0.14–0.77 ng/L, Meuleman et al., 1995), Canadian Cil Lk(0.24–0.36 ng/L, Evans and Lockhart (2001). However, Hgconcentrations reported by Semkin et al. (2005) were signifi-cantly lower than levels in Swedish lakes (1.35–15 ng/L, Leeand Iverfeldt, 1991) and in Finland lakes (1.3–7.2 ng/L, Vertaet al., 1994).

The first measurements of high-time resolution of dis-solved gaseous mercury (DGM) concentrations over a large-scale distribution on the North Atlantic Ocean (latitudes from54°N to 85°N) were performed by Temme et al. (2005) insummer 2004. DGM concentration increased from 11–14 pg/Lat latitudes between 62°N and 74°N to 35 pg/L at latitudesN74°N, where the ocean was mostly covered with sea ice.These researchers hypothesized that the elevated DGMconcentrations at higher latitudes could be due to the outputof Hg from the melting sea ice to the surface ocean water.During the summer, thewater of theNorth Atlantic Oceanwashighly supersaturated in terms of DGM, which agree well withthe study by Sommar et al. (2004). These authors observed astrong diurnal cycle of DGM in the range from 12 to 70 pg/L atKongsfjorden (Ny-Alesund) during spring time, which corre-sponds to a (super) saturation of 40–480%. The diurnalvariation of DGM was clearly linked to solar radiation,suggesting photolytic reduction of oxidized Hg species insurface water (Sommar et al., 2004). In the same study, totalmercury (THg) in surface water was 2.1±1.5 ng/L (n=5), whichis significantly lower that the value in run-off of slushy snow(~30 ng/L) (Lindberg et al., 2002) and in rain precipitation (4–52 ng/L) (Berg et al., 2001).

Once absorbed into the water ecosystem from natural andanthropogenic emission sources, point and non-pointsources, and spills (Chapman et al., 1998), Hg can pass fromsolution or suspension into the sediments (Riget et al., 2000a).Microorganisms can transform inorganic Hg to organic Hg,which in turn accumulate in animals, and biomagnifies in thefood web (Riget et al., 2000a).

Many researchers have indeed used the paleolimnologicalarchives (i.e. lake sediments) to reconstruct local and regionalhistories of the evidence for temporal and spatial trends in Hgdeposition (Engstrom et al., 1994; Gubala et al., 1995; Lucotteet al., 1995; Landers et al., 1998). When lake sediments have

been carefully dated using 210Pb methodology (Robbins, 1978),it is generally accepted that total Hg mass accumulationprofiles represent a robust record, integrated over years anddecades, of the major changes in Hg deposition to awatershed. The use of Hg deposition fluxes has beensuggested as a single indicator of the degree to which thelakes and watersheds have been enriched in Hg in recenttimes. Deposition fluxes tend to minimize differences asso-ciatedwith laboratory bias and variability in “background” fluxamong individual lakes due to geological and watershedattributes. By evaluating the relationship between Hg fluxesin sediments deposited in the pre-industrial levels (e.g. b1880)and in the recent post-industrial era, enrichment factors havebeen estimated (e.g. Landers et al., 1998; Hermanson, 1998;Bindler et al., 2001).

Several papers have been published reporting Hg concen-trations in sediments from several lakes and watersheds inthe Arctic and Sub-Arctic regions. Lockart et al. (1998) reportedHg levels in sediment cores from 18 lakes from central andnorthern Canada varying between 0.036 and 0.159 µg/g (dw).Similar Hg concentrations were also reported by Outridge el al.(2007) in Amituk and DV-09 lakes in Canadian high Arctic.Higher Hg concentrations were however determined byHammerschmidt et al. (2006) for four lakes from Alaska,USA. In the Alaska lakes Hg levels varied between 0.087 and0.327 µg/g (dw). Hg concentrations ranging between 0.02 to0.110 and from 0.06 to 0.440 µg/g (dw) were determined inGreenland lake sediments by Lindeberg et al. (2006) andBindler et al. (2001), respectively.

In other study focusing on investigating Hg levels infreshwater system of Greenland, Hg in lake sediment, in soiland humus from the surrounding area were determined atfour areas in Greenland (Table 1). Hg concentrations in soilfrom the surrounding of the lakes were below or slightly abovethe detection limit in all locations. Hg levels in humus and lakesediment were also in a low range (0.02–0.12 µg/g, dw and0.018–0.05 µg/g, dw, respectively). Steinnes and Sjobakk (2005)published Hg concentrations in surface sediments from sixlakes in Sweden varying between 0.036 and 0.364 µg/g (dw). Inmost of the lakes an increase in Hg concentrations wasobserved upward suggesting an increase of Hg deposition inthese catchments areas in recent decades. Thus, Hg concen-tration variations along sediment core slices coupled to thedetermination of 210Pb levels enabled to estimate core datingas well Hg deposition rates. This information has beenparticularly important in order to relate the Hg flux/enrich-ment in each sediment slice with data from atmospheric Hgdeposition and/or time scale. Table 1 presents depositionfluxes in sediments from several lakes in Arctic and sub-Arcticregions. Hermanson (1998) determined that the deposition ofHg in Imitavik lake (Canada) increased by a factor of about 6 (2to 12.5 ng/m2y) between 1750 and 1980 and between 1750 and1867 the increase was a factor of 2 (2–4 ng/m2y). This variationwas explained by variations in Hg concentrations in morenorthern latitudes related, for example, with increase of theintensity of silver mining in Mexico from about 1700 to 1800 orthe beginning of the industrial revolution.

In Greenland lake sediments Bindler et al. (2001) observedthat the first increase in Hg accumulation rate occurred notuntil the mid-1800s (1–3 µg/m2y). From this period forward, Hg

Table 1 – Hg concentrations and Hg sediment deposition rates in freshwater, lake sediments, marine water and sediments

Substancetype

Region Location Sampleperiod

Total Hg(ng/L) or(µg/g, dw)

MeHg(ng/g, dw)

Hg deposition rates(µg/m2y)

Reference

Pos-Industrial

Pre-Industrial

Lake CornwallisIsland

Amituk Lake 1994 0.23–0.33 – – – Semkin et al. (2005)Inflow of AmitukLake

1994 0.76 – – – Semkin et al. (2005)

Outflow ofAmituk Lake

1994 0.52 – – – Semkin et al. (2005)

Lakesediment

Greenland Avanersuaq 1994–95 0.023 – – – Riget et al. (2000a)Nuuk 1994–95 0.052 – – – Riget et al. (2000a)Qaqortoq 1994–95 0.025 – – – Riget et al. (2000a)Tasillaq 1994–95 0.018 – – – Riget et al. (2000a)Avanersuaq 1994–95 0.01 – – – Riget et al. (2000a)Soil GreenlandNuuk 1994–95 0.02 – – – Riget et al. (2000a)Qaqortoq 1994–95 0.03 – – – Riget et al. (2000a)Tasillaq 1994–95 b0.01 – – – Riget et al. (2000a)Avanersuaq 1994–95 0.0367 – – – Riget et al. (2000a)Humus GreenlandNuuk 1994–95 0.117 – – – Riget et al. (2000a)Qaqortoq 1994–95 0.0751 – – – Riget et al. (2000a)Tasillaq 1994–95 0.0199 – – – Riget et al. (2000a)

Elson lagoon,Alaska, grab

71.20°N, 156°34W 2001–02 0.0654 0.012 – – Naidu et al. (2003)71.19°N, 156°32W 2001–02 0.0639 0.054 – – Naidu et al. (2003)71.18°N, 156°24W 2001–02 0.0978 0.133 – – Naidu et al. (2003)71.23°N, 156°27W 2001–02 0.0908 0.066 – – Naidu et al. (2003)71.21°N, 156°31W 2001–02 0.0541 0.047 – – Naidu et al. (2003)Core depth:0–1 cm

2001–02 0.063 0.002 – – Naidu et al. (2003)

Lagoonsediment

Elson lagoon,Alaska, core71.18°N, 156°24W

1–2 cm 2001–02 0.0519 0.008 – – Naidu et al. (2003)2–3 cm 2001–02 0.0530 0.008 – – Naidu et al. (2003)3–4 cm 2001–02 0.0529 0.001 – – Naidu et al. (2003)4–5 cm 2001–02 0.0532 0.037 – – Naidu et al. (2003)5–6 cm 2001–02 0.0544 0.068 – – Naidu et al. (2003)6–7 cm 2001–02 0.0575 0.081 – – Naidu et al. (2003)7–8 cm 2001–02 0.0525 0.035 – – Naidu et al. (2003)8–9 cm 2001–02 0.0638 0.032 – – Naidu et al. (2003)18–19 cm 2001–02 0.0735 0.013 – – Naidu et al. (2003)

Lakesediment

Norway A. Gjerstad 2000 8.3–204 – 8.5 – Steinnes and Sjobakk (2005)B. Gyland 2000 7.2–349 – 9.9 0.29 Steinnes and Sjobakk (2005)C. Fure 2000 11.3–196 – 4.9 0.72 Steinnes and Sjobakk (2005)D. Namsos 2000 7.6–106 – 3.6 0.23 Steinnes and Sjobakk (2005)E. Nordli 2000 3.6–146 – 2.1 0.22 Steinnes and Sjobakk (2005)F. Andoya 2000 15.1–364 – 11.1 0.49 Steinnes and Sjobakk (2005)

Lakesediments

Greenland Kangerlussuaq – 0.02–120 – – – Lindeberg et al. (2006)

Lakesediments

Greenland Kangerlussuaq 1997–00 0.06–0.44 – 5–8 1–5 Bindler et al. (2001)

Lakesediments

Canada Central andNorthern Canada

1995 23–155 – 2.1–114 1.1–52.7 Lockhart et al. (1998)

Canada Amituk Lake 1999 0.02–0.12 – – – Outridge et al. (2005)DV-09 1999 0.01–0.03 – 2.5–10 Outridge et al. (2005)

Lakesediments

Canada Amituk Lake 2003 0.02–0.12 – 31.7 Outridge et al. (2007)Lakesediments DV-09 2003 0.015–0.03 – 13.1 Outridge et al. (2007)

Imitavik Lake 1990 – – 12.5–18 2–12.5 Hermanson (1998)Lakesediments

Canada Annak Lake 1993 – – 250 2.5 Hermanson (1998)

Lakesediments

Finland Several – – – 2.3–49.5 1.9–27Sweden Several – – – 6.7–20.7 2.3–17.2 Landers et al. (1998)Russia Several – – – 7.1–30.2 6.6–23.3 Landers et al. (1998)W. Canada Several – – – 5–21.3 0.7–23.3 Landers et al. (1998)Québec Several – – – 5.3–33.9 2.6–24.1 Landers et al. (1998)USA-Alaska Several – – – 3.3–52.2 2.5–41.1 Landers et al. (1998)USA-Mid. Cont. Several – – – 16.5–48.8 4.5–23 Landers et al. (1998)

Lakesediments

Alaska Several – 102–239 0.33–2.38 – – Hammerschimdtet al. (2006)


Table 1 (continued)

Substancetype

Region Location Sampleperiod

Total Hg(ng/L) or(µg/g, dw)

MeHg(ng/g, dw)

Hg deposition rates(µg/m2y)

Reference

Pos-Industrial

Pre-Industrial

Marinesediments

Greenland – 1985–87 0.05–0.3 – – – Asmund and Nielsen(2000)

Marinesediments

Arctic Ocean Several 1994 0.05–0.15 – – – Gobeil et al. (1999)

Marinesediments

Russia Kara Sea 1965 195–875 – – – Galasso et al. (2000)Avanersuaq 1994–95 0.0045 – – – Riget et al. (2000c)Qeqertarsuaq 1994–95 0.018 – – – Riget et al. (2000c)

Marinesediment

Greenland Nanortalik 1994–95 0.033 – – – Riget et al. (2000c)Ittoqqortoormiit 1994–95 0.016 – – – Riget et al. (2000c)


flux to sediments of these lakes increases until about 1970s/1980s when the Hg accumulate rates peak about 5–10 µg/m2y.Landers and co-authors (1998) reportedHg deposition fluxes inpre- and post-industrialization in a larger study performed insediments from several lakes from Finland, Sweden, Russia,West Canada, Québec (Canada) and in USA Alaska and mid-continent (Table 1). These authors used post/pre industriali-zation ratios for all lakes and observed that ratios differedwidely across the arctic regions of the northern hemisphere.They determined highest flux ratios in regions associatedwithstrong regional sources of atmospheric Hg emissions such ascentral and eastern North America and Central Europe.

Lakes throughout the U.S. Arctic and Taimyr Peninsula inRussia showed only slight enrichments over backgroundfluxes. Landers et al. (1998) concluded that the higherdeposition flux ratios located about 2000 km or less fromurban/industrial sources than remote lakes are predominantlya regional phenomena while approximately 30% enrichmentof remote lakes may be viewed as the hemispheric (or at leastcircum-arctic) background enrichment factor.

In spite of all these findings, the interpretation ofsedimentary Hg profiles in most lakes as an unambiguousrecord of Hg deposition has been contested, due to thepossibility of post-burial remobilization resulting from diag-enesis: physical mixing by wind and wave action, compacta-tion, bioturbation and irrigation and, in extremely lowsedimentation environments, diffusivemigration of dissolvedHg as a result of redox-related concentration gradients(Boudreau 1999; Boyle, 2001).

Another potential problem to sediments as an Hg archivesis the largely-untested assumption that the transport pro-cesses of Hg from atmosphere to sediments via the catch-ments and water column have not changed in recent history(Outridge et al., 2005). This assumption is necessary forsedimentary metal concentration or fluxes to be interpretedas simply and directly reflecting trends in the atmosphericdeposition of Hg (Outridge et al., 2005).

Findings that global warming has caused significantincreases in the rate of spring thaw (Magnuson et al., 2000)and diatom productivity (Douglas et al., 1994; Wolfe, 2000) innorthern lakes, and the discovery of atmospheric “depletionevents” (Schroeder et al., 1998), challenge this assumption. Insummary, if processes and fluxes changed significantly andthe changes were not recognized and corrected for, conclu-

sions about atmospheric metal history based on sedimentprofilesmay be erroneous (Outridge et al., 2005). Asmentionedbefore a recent study published by Outridge et al. (2007)reported that variations in Hg concentration in cores from twolakes in the Canadian high Arctic could not be explained onlyby variations in atmospheric deposition. These authorsclaimed that climate warming in the 20th century hadprofound effects on the limnology the high Arctic lakes,including substantial increases in autochthonous primaryproductivity (APP). Through organic carbon and Hg measure-ments in sediment cores from the two lakes, Outridge and co-authors defended the hypothesis that 20th century increasesin sedimentary Hg at these latitudes were largely driven byAPP increases, via Hg scavenging by algae and/or suspendeddetrital algae matter. This finding suggest that the atmo-spheric contribution of long-range anthropogenic Hg to highArctic lakes may have been overestimated by several-foldbecause of this climate-driven process andwas responsible fornomore than 22% of the 20th century Hg increase in the studylakes. According to Outridge et al. (2007) this scavenginghypothesis has the potential to explain long-standing anoma-lies in the Arctic study.

Other process effecting Hg variations in sediment coreswas proposed by Lindeberg et al. (2006) for three lakes from theGreenland Arctic. By using deeper sediment layers Lindebergand co-authors showed that changes in Greenland climatecaused changes in lake influx of material from regionalaeolian activity that transport deflates material from thesandy outwash plants at the ice sheet margin and deposits iton lakes and their catchments. According to these authors,this material has an important influence on the Hg concen-trations especially in times before the deposition fromanthropogenic emissions of the last hundred years.

Compared to total Hg, information about MeHg in Arcticfreshwater sediments are scarce. A recent study by Ham-merschmidt et al. (2006) reported MeHg levels in Alaskan lakesediments varying between 0.33 and 2.38 ng/g, accounting toless than 1% of the total Hg present. In this work, Ham-merschmidt and co-authors also studied the multiple aspectsof the aquatic MeHg cycle, including Hg speciation and thedetermination of in situ Hg methylation potential andconcluded that loadings of Hg derived from atmosphericdeposition were a major factor affecting MeHg cycling in theArctic ecosystems. Additionally, they claimed that

Table 2 –Mercury levels in tissues of freshwater fish collected at tundra region

Species Region Tissue Length(mm)

Sample period Hg Level(µg/g, dw)

Reference

landlocked Arcticchar (Salvelinus alpinus)

Kaniqsujuaq Muscle – 1998 0.14±0.03 Muir et al. (2001a,b)Fort Good Hope Muscle 676±107 1999 0.286±0.095 Stern et al. (2001)

Burbot (Lota lota) Fort Good Hope Muscle 699±104 2000 0.345±0.097 Stern et al. (2001)Fort Good Hope Muscle 735±101 1999 0.259±0.108 Stern et al. (2001)Fort Good Hope Muscle 732±127 2000 0.364±0.14 Stern et al. (2001)Fort Good Hope Liver 676±107 1999 0.046±0.024 Stern et al. (2001)Fort Good Hope Liver 699±104 2000 0.064±0.026 Stern et al. (2001)Fort Good Hope Liver 749±77 1999 0.064±0.069 Stern et al. (2001)Fort Good Hope Liver 732±127 2000 0.094±0.056 Stern et al. (2001)Avanerauaq, Greenland Muscle 412±56.1 1994–1995 0.23±0.135 Riget et al. (2000a)Nuuk, Greenland Muscle 366±28.9 1994–1995 0.696±0.266 Riget et al. (2000a)Tasiilaq river, Greenland Muscle 235±35.2 1994–1995 0.12±0.036 Riget et al. (2000a)Tasiilaq lake, Greenland Muscle 47.8±27.5 1994–1995 0.40±0.069 Riget et al. (2000a)Qaqortoq, Lake A, Greenland Muscle 392±82.9 1994–1995 0.801±0.605 Riget et al. (2000a)Qaqortoq, Lake B, Greenland Muscle 347±70.8 1994–1995 0.566±0.34 Riget et al. (2000a)Qaqortoq, Lake C, Greenland Muscle 350±31.4 1994–1995 0.642±0.296 Riget et al. (2000a)Qaqortoq, Lake D, Greenland Muscle 335±35.9 1994–1995 0.624±0.346 Riget et al. (2000a)

Lake whitefish(Coregonus clupeaformis)

Aubry Lake Muscle 507 1999 0.048 Lockhart et al. (2000, 2001)Cli Lake Muscle 495 1996, 2000 0.078 Lockhart et al. (2000, 2001)Colville Lake Muscle 447 1993,99 0.03 Lockhart et al. (2000, 2001)Deep Lake Muscle 444 2000 0.249 Lockhart et al. (2000, 2001)Ekali Lake Muscle 470 1996 0.082 Lockhart et al. (2000, 2001)Kelly Lake Muscle 495 1998 0.165 Lockhart et al. (2000, 2001)Little Doctor Lake Muscle 407 1996 0.13 Lockhart et al. (2000, 2001)Mahoney Lake Muscle 510 1996 0.131 Lockhart et al. (2000, 2001)Manuel Lake Muscle 505 1978,89,93,95,97,98 0.111 Lockhart et al. (2000, 2001)McEwan Lake Muscle 450 2000 0.088 Lockhart et al. (2000, 2001)McGill Lake Muscle 392 2000 0.15 Lockhart et al. (2000, 2001)Mirror Lake Muscle 457 2000 0.351 Lockhart et al. (2000, 2001)Reade (Unnamed) Lake Muscle 427 2000 0.146 Lockhart et al. (2000, 2001)Sibbeston Lake Muscle 422 1997 0.071 Lockhart et al. (2000, 2001)Tagatui Lake Muscle 337 1996 0.035 Lockhart et al. (2000, 2001)Tsetso Lake Muscle 421 1997 0.075 Lockhart et al. (2000, 2001)Turton Lake Muscle 419 1996 0.113 Lockhart et al. (2000, 2001)Willow Lake Muscle 405 1999 0.085 Lockhart et al. (2000, 2001)

Broad whitefish(Coregonus nasus)

Peel River Muscle – 1999 0.08 Lockhart et al. (2000, 2001)

Inconnu Kelly Lake Muscle 760 1998 0.396 Lockhart et al. (2000, 2001)(Stenodus leucichthys nelma) Peel River Muscle – 1999 0.256 Lockhart et al. (2000, 2001)Arctic clsco(Coregonus autumnalis)

Bandy Lake Muscle 325 2000 0.229 Lockhart et al. (2000, 2001)Deep Lake Muscle 155 2000 0.26 Lockhart et al. (2000, 2001)Ekali Lake Muscle 338 1996 0.118 Lockhart et al. (2000, 2001)McEwan Lake Muscle 197 2000 0.093 Lockhart et al. (2000, 2001)Sanguez Lake Muscle 267 1996 0.158 Lockhart et al. (2000, 2001)

Lake trout(Salvelinus namaycush)

Atin Lake Muscle 577 1993, 98 0.218 Lockhart et al. (2000, 2001)Aubry Lake Muscle 576 1999 0.254 Lockhart et al. (2000, 2001)Belot Lake Muscle 618 1993, 99 0.192 Lockhart et al. (2000, 2001)Cli Lake Muscle 481 1983, 96, 2000 0.79 Lockhart et al. (2000, 2001)Coal Lake Muscle 344 1999 0.072 Lockhart et al. (2000, 2001)Colville Lake Muscle 512 1993, 99 0.204 Lockhart et al. (2000, 2001)Fox Lake Muscle 400 1998 0.397 Lockhart et al. (2000, 2001)Itsi Lake Muscle – 2001 0.086 Lockhart et al. (2000, 2001)


Kelly Lake Muscle 598 1998 0.482 Lockhart et al. (2000, 2001)Kusawa Lake Muscle 515 1999 0.573 Lockhart et al. (2000, 2001)Laberge Lake Muscle 518 1993, 96, 98, 99 0.413 Lockhart et al. (2000, 2001)Little Doctor Lake Muscle 547 1996 0.393 Lockhart et al. (2000, 2001)Loon Lake Muscle 600 1997 0.369 Lockhart et al. (2000, 2001)Mahoney Lake Muscle 672 1996 0.367 Lockhart et al. (2000, 2001)Manuel Lake Muscle 485 1997 0.299 Lockhart et al. (2000, 2001)Mirror Lake Muscle 460 2000 0.606 Lockhart et al. (2000, 2001)Nares Lake Muscle 530 1996 0.343 Lockhart et al. (2000, 2001)Quiet Lake Muscle 553 1992, 99 0.302 Lockhart et al. (2000, 2001)Saturday Night Lake Muscle 399 1998 0.221 Lockhart et al. (2000, 2001)Turton Lake Muscle 566 1996 0.6 Lockhart et al. (2000, 2001)


Table 2 (continued)

Species Region Tissue Length(mm)

Sample period Hg Level(µg/g, dw)

Reference


Watson Lake Muscle 515 1997 0.152 Lockhart et al. (2000, 2001)Willow Lake Muscle 614 1999 0.38 Lockhart et al. (2000, 2001)

North pike (Esox lucius) Bandy Lake Muscle 591 2000 0.323 Lockhart et al. (2000, 2001)Cli Lake Muscle 660 2000 0.353 Lockhart et al. (2000, 2001)Colville Lake Muscle 630 1999 0.244 Lockhart et al. (2000, 2001)Deep Lake Muscle 584 2000 0.67 Lockhart et al. (2000, 2001)Ekali Lake Muscle 578 1996 0.3 Lockhart et al. (2000, 2001)Gargan Lake Muscle 668 1996 0.587 Lockhart et al. (2000, 2001)Kelly Lake Muscle 673 1998 0.546 Lockhart et al. (2000, 2001)Little Doctor Lake Muscle 695 1996 0.772 Lockhart et al. (2000, 2001)Loon Lake Muscle 648 1997 0.505 Lockhart et al. (2000, 2001)Mahoney Lake Muscle 660 1996 0.257 Lockhart et al. (2000, 2001)Manuel Lake Muscle 637 1978, 95, 97, 98 0.441 Lockhart et al. (2000, 2001)McEwan Lake Muscle 569 2000 0.331 Lockhart et al. (2000, 2001)McGill Lake Muscle 577 2000 0.713 Lockhart et al. (2000, 2001)Nares Lake Muscle 870 1996 0.258 Lockhart et al. (2000, 2001)Reade (Unnamed) Lake Muscle 584 2000 0.43 Lockhart et al. (2000, 2001)


Sanguez Lake Muscle 683 1996 0.703 Lockhart et al. (2000, 2001)Sibbeston Lake Muscle 671 1997 0.165 Lockhart et al. (2000, 2001)Tagatui Lake Muscle 634 1996 0.17 Lockhart et al. (2000, 2001)Tsetso Lake Muscle 763 1997 0.393 Lockhart et al. (2000, 2001)Willow Lake Muscle 634 1999 0.293 Lockhart et al. (2000, 2001)

Walleye(Stizostedion vireum vitreum)

Deep Lake Muscle 448 2000 1.105 Lockhart et al. (2000, 2001)Ekali Lake Muscle 410 1996 0.256 Lockhart et al. (2000, 2001)Little Doctor Lake Muscle 474 1996 0.753 Lockhart et al. (2000, 2001)Mackenzie River Muscle 352 1997 0.191 Lockhart et al. (2000, 2001)McEwan Lake Muscle 430 2000 0.356 Lockhart et al. (2000, 2001)Sanguez Lake Muscle 480 2000 1.125 Lockhart et al. (2000, 2001)Sanguez Lake Muscle 440 1996 0.539 Lockhart et al. (2000, 2001)Sibbeston Lake Muscle 468 1997 0.327 Lockhart et al. (2000, 2001)Tathlina Lake Muscle 393 1981, 90, 93, 94, 98 0.452 Lockhart et al. (2000, 2001)Tsetso Lake Muscle 493 1997 0.485 Lockhart et al. (2000, 2001)


environmental changes associated with the warming of theArctic (increase weathering, temperature, productivity, andorganic loadings) could enhance MeHg production as thetundra and lake sediments warm.

Naidu et al. (2003) conducted a one-year study focusing ondetermining the concentrations of total Hg and MeHg insediments of Elson Lagoon (Barrow, Northwest Arctic Alaska)in 2001–2002. The sediment samples consisted of five vanVeen grab samples and one of the three gravity cores collectedat selected locations (Table 1). Total Hg levels and MeHg ingrab and core sediments were in the range of 0.052–0.098 µg/g(dw) and 0.081 ng/g (dw), respectively. Stratigraphic samplesshowed a net significant down core increase in THg (pb0.05).The down core increase in THg may be due to increaseconcentration of Hg by methylation in progressively anoxicsediment layers, as suggested by the presence of increasinglevels of MeHg from the upper 4 cm composite of the core(representing possibly oxic layers) to about 7 cmdown the core(Naidu et al., 2003).

3.2. Hg levels in lake/pond biota

Hg contamination in the biological Arctic environment hasraised substantial concern because there are indications thatHg levels have been increasing over the past few decades (Fisket al., 2003). In some Arctic biota, Hg concentrations are

present at high levels due to bio-magnification of MeHgthrough food chain.

3.2.1. Freshwater fishFresh fish are an important dietary item for indigenous peopleand in some cases are a commercial resource through sportfishing and the sale of fish for consumption (Fisk et al., 2003).In Canada, consumption of freshwater fish is generally limitedto Yukon, western Northwest Territories (NWT) and northernQuébec, but landlocked Arctic char (Salvelinus alpinus) areavailable in some Nunavut communities (Fisk et al., 2003). Themajority of Hg concentration values have been determined astotal Hg (inorganic+organic), but Hg found in freshwater fishis predominantly in the form of MeHg. MeHg bioaccumulationin fish is species specific and variable due to growth effects(Cabana et al., 1994; Greenfield et al., 2001). The frequency andspecies of fish consumed are then key determinants for properrisk assessment (Flaherty et al., 2003). A fairly large data setof Hg concentration in Arctic freshwater fish was availablein several studies. Table 2 summarizes Hg levels in tissues offreshwater fish collected at tundra region. For example, Rigetet al., (2000a) reported mean Hg levels of 0.12±0.04–0.80±0.61 µg/g in muscle tissue of freshwater Arctic char in Green-land and approximately 92% ofwhichwasMeHg. ThemeanHglevels reported by Riget et al. (2000a) seem to be in the samerange as found in Arctic Canada (ug/g, wet wt) and somewhat


higher than in Finnish Lapland (0.09–0.32 µg/g, wet wt.), Ice-land (0.02–0.03 µg/g, wet wt.), Norway (0.03–0.25 µg/g, wet wt.),Russia (0.01 µg/g, wet wt.) and Sweden (0.10 µg/g, wet wt.)(Nilsson, 1997).

A substantial survey regarding total Hg levels in northernCanadiancommercial fish for thepast 25yearswas conductedbythe Canadian Department of Fisheries andOceans (Braune et al.,1999). Results indicated that Hg levels in fish tissues in theNorthwest Territory (NWT) Yukon and northern Québec fre-quentlyexceedguideline limits for subsistence fishconsumptionor commercial sale (0.5 µg/g) (Health and Welfare Canada, 1979,1984; Brauneetal., 1999).High levelsofHg in freshwater fishwerealso documented byMuir et al. (1997) in the first Canadian ArcticContaminants Assessment Report (CACAR). Data collected atCanadian Arctic (Lockhart et al., 2000, 2001; Fisk et al., 2003)indicate that Hg concentrations in muscle were higher than inliver. Similarly, Stern et al. (2001) reported higher Hg burdens inmuscle of burbot (Lota lota) from Fort GoodHope than in liver. Onthe other hand, the highest Hg levels were observed in pike andwalleye in comparison with the levels in burbot.

The spatial coverage of Hg in freshwater fish was investi-gated in several studies (Jensen et al., 1997; Lockhart et al.,2001). For example, Lockhart et al. (2001) documented that fishfrom Yukon to northern Nunavut contain high levels of Hg,particularly in Nunavut, where lake trout have an unusuallyhigh proportion of values exceeding the limit for marketingrecommended by Health Canada. Large spatial lake-to-lakevariation regarding Hg levels in the same species of fish wasalso reported by Braune et al. (1999). These investigators foundthat Hg in Lake Whitefish from Lac Se (Yukon) were approxi-mately 40 times higher than those from Colville Lake in thewestern NWT. High or low levels of Hg in the fish were foundthroughout the Arctic and no clear geographic pattern evidentin Hg concentrations was found. This likely reflects thevariability between lake systems and food webs.

Some geographic factors, including changes in wetlandareas and reservoir establishment or instances of floodingmight have important influences on Hg concentrations in fishin some regions (St. Louis et al., 1995). However, Evans andLockhart (2001) did not reveal any strong relationshipsbetween Hg levels in fish and some factors such as lake size,drainage basin area and wetland area. Though, these authorsfound that the humic content of the water may play a role inthe uptake of Hg in fish. This is because higher humic contentfavors the transport and retention of Hg in the water phase, asorganic complexes, which may increase the availability of Hg

Table 3 –Mercury levels in waterfowl and game birds

Species Region Subpopulation Tis

Lesser snow geeseAnser c.caerulescens

Wrangel Island, Russia Southern EggLiv

Northern EggLiv

Common loonGavia immer

Canada Québec, Ontario,Atlantic

KidLivMu

Common merganserMergus merganser

KidLivMu

for uptake in the food chain. Therefore, Hg uptake in coloredstreams is higher than in clear ones (Evans and Lockhart,2001). Earlier studies of acidified lakes in Sweden and Norwayindicated that the process of acidification increased Hgaccumulation in fish, especially where the food web hadbeen damaged (Andersen et al., 1986; Björklund et al., 1984).Methylation is also known to be favored by increasing acidityand temperature (Evans and Lockhart, 2001). Therefore lowconcentrations of MeHg in cold temperatures and alkalinebedrock areas such as occur in large parts of Arctic Canada areexpected. Interestingly, no strong correlation was foundbetween the Hg concentrations detected in lake sedimentsand fish. For instance, the Hg concentrations in sediment inthe lake of Ivvavik National Park (69°26 N 139°36 W), NWYukon, were 20–40 times higher than in the other lakes in theCanadian Arctic during the Tundra North West Expedition in1999, however Hg levels in fish muscle were still not elevatedcompared to the rest of the lakes (Borg, 2001).

The influence of fish characteristics on Hg levels was in-vestigated in several studies. Lockhart et al. (2000) and Muiret al. (2000) reported effects of age and size of fish on their Hglevels. However, quite a fewstudies did not link higher levels ofHg to large fish (Amundsen et al., 1997; Evans and Lockhart,2001). For example, Stern et al. (2001) did not find significantcorrelation between length and Hg concentration in muscleand liver in burbot from Fort Good Hope. Evans and Lockhart(2001) suggested that fish had high Hg concentrations in someArctic Canadian lakes were due to their slow growth rate.Though, trophic position could also play a significant role indeterminingHg levels in freshwater fish (Details see Section 6).

3.2.2. Waterfowl and game birds

3.2.2.1. Lesser snow geese (Anser Chen caerulescens). Hglevels in eggs and livers of lesser snow geese from northernand southern subpopulations ofWrangel Island (Russia) wereexamined by Hui et al. (1998) (Table 3). The northernsubpopulation migrates to Fraser River delta (Vancouver,British Columbia) and the Skagit River delta of northernWashington in winter, while the southern subpopulationspendswinter time in the Central Valley of California. Resultsindicate that Hg levels in the tissues of lesser snow geesewere considerably low with the Hg levels in livers less thandetection limit (DL). The low Hg levels in these two sub-populations were probably due to the low Hg content in theirfood diet (bDL—0.06 µg/g dw).

sue Sampleperiod

Hg Level(µg/g, dw)

Reference

1992 0.07±0.07 Hui et al. (1998)er 1992 b limits of detection Hui et al. (1998)

1992 0.06±0.07 Hui et al. (1998)er 1992 b limits of detection Hui et al. (1998)ney 15±15 Scheuhammer et al. (1998)er 19±25 Scheuhammer et al. (1998)scle 2.9±1.1 Scheuhammer et al. (1998)ney 11±6.2 Scheuhammer et al. (1998)er 15±12 Scheuhammer et al. (1998)scle 3.0±0.2 Scheuhammer et al. (1998)


3.2.2.2. Common loon (Gavia immer) and common merganser(Mergus merganser). A study focusing on Hg levels in tissuesof two species of piscivorous birds, namely common loon andcommon merganser, was conducted by Scheuhammer et al.(1998) and thedata are summarized inTable3. For both commonloons and mergansers, liver generally had the highest total Hgconcentrations, followed by kidney, with breast muscle havingthe lowest concentrations. From inter-species comparisonbetween healthy loons and mergansers, no significant differ-enceofHgconcentrations inany three tissuesof twospecieswasreported. However, mean total Hg concentrations in tissues ofemaciated loonsweremuchhigher than those than in tissues ofapparently healthy birds. Especially, the levels in livers ofemaciated loons were more than double of healthy birds. Inboth loons and mergansers, the MeHg fraction decreased withincreasing total Hg concentrations in kidney and liver, butremained fairly constant and high in breast muscle (~94% oftotal Hg for loon breast muscle and ~78% for merganser breastmuscle). The mean total Hg concentration in emaciated loons(68.3 µg/g drywt)was similar to the data reportedbyHeinz (1976)andFinley andStendell (1978),who fedadiet containing3µg/g ofMeHg to female mallards (Anas platyrhynchos) and adult balkducks (A. rubripes) in two studies. Therefore, Hg levels in thetissue of several emaciated loons were sufficiently high to betoxic basedon theavailable toxicological literature. TheelevatedHg concentrations in emaciated loons may have contributed totheir impaired conditions. Alternatively, the wasting of themuscle and other organs in these birds may have causedapparent tissue Hg concentrations to increase.

3.2.3. Mammals

3.2.3.1. Beaver (Castor canadensisi) and raccoon (Procyon lotor).Beaver and raccoon are furbearingmammal species. In a study toinvestigate Hg concentrations in semi-aquatic mammals (Wren,1984), Hg levels in tissue samples of beaver and raccoon from awatershed in Ontario were reported (Table 4). Hg concentrationsin the different tissues of beaverwere similar. On the other hand,Hg levels in raccoon were much higher in comparison with thevalues in beaver. The highest concentrations in raccoon werefound in liver and lowest were reported in muscle. Interestingly,from concurrent measurements of Hg and selenium levels insample tissues of three semi-aquatic mammal species, signifi-cant correlation between Hg and selenium levels in liver tissueswere found in these mammals. Wren (1984) suggested thatselenium accumulation in piscivorous mammals may representa protective mechanism against MeHg toxicity to animalsexposed to high Hg levels in their diet.

3.2.3.2. River otter (Lutra canadenisis). River otter is apiscivorous mammal that occupies the highest level of thefood chain (Wise et al., 1981; Foley et al., 1988). River otter has afairly small home range, lives close to watersheds, and prefersto catch prey in rivers, streams, and ponds (Gilbert andNancekivell, 1982). In a study performed by Fortin et al. (2001),total Hg levels in tissues of river otter collected from James Bayterritory (Québec) were reported (Table 4). Hg concentrationsvaried from0.72±0.21 µg/g (w.w) in brain to 19.5±0.15 µg/g (w.w)in fur/hair, which were normally lower than the guideline of30 µg/g (wet wt.) known to cause mercurialism in experimen-

tally dosed individuals (O'Conner and Nielsen, 1981). The orderof total Hg concentrations across river otter organs/tissueswas:fur/hairNN liver=kidneyNmuscleNbrain. The highest Hg con-centrations in fur/hair suggest that fur/hair samples could beused to monitor environmental contamination. Hg levels inlivers and muscle of otters were similar to the values collectedat south central Ontario watershed (Wren, 1984). However,Fortin et al. (2001) reported much higher Hg concentrations inkidney than the levels reported by Wren (1984).

Organic Hg concentrations across tissues/organs of otterswere also investigated by Fortin et al. (2001). Organic Hgfluctuated from 0.42±0.25 to 1.78±0.17 (µg/g, w.w) with similarlevels in liver, kidney and muscle, and an exception of lowlevel in brain: kidney=liver=muscleNbrain. Organic Hg repre-sented 44–90% of total Hg level in different tissues and thehighest percentage of organic Hg was found in muscle andbrain samples. No age, sex and body mass effect on theaccumulation of total Hg levels in liver of otters was observed.

3.2.3.3. Mink (Mustela vison). Mink is a top trophic levelcarnivore found throughout the forested regions of the NWT inCanada (Poole et al., 1998; Braune et al., 1999). Small mammalsand fish form the greatest components of mink diet in mostareas (Gilbert and Nancekivell, 1982), thus mink are exposed tocontaminants derived from both terrestrial and aquatic foodwebs. Mink readily bio-accumulates environmental pollutants,thus mink maybe a sensitive indicator of ecosystem health(Poole et al., 1995; Braune et al., 1999).

In a study by Fortin et al. (2001), Hg concentrations wereanalyzed in different organs/tissues of wild minks trapped inJames Bay, Québec (Canada) during 1993–1994 and 1994–1995trapping seasons. Hg levels varied greatly across organs/tissues(0.82 to 26.9 µg/g, wet wt.). The sequence of total Hg levels inorgan/tissue samples was as follow: fur/hairNN liverNkidney=muscleNbrain (Table 4). Total Hg levels in fur/hair were ap-proximately 40 timehigher that those inbrain. Among319minkcarcasses collected in the James Bay territory, only two wildminks (1%)exceeded the lowest limit of 21µg/g (wetwt.) to causemercurialism in experimentally dosed semi-domesticatedminks (Aulerich et al., 1974; Wobeser and Swift, 1976). OrganicHg concentration in wild mink displayed a similar patternacross organs/tissues as total Hg: liver=muscleNkidneyNbrain.The percentage of organic Hg in terms of total Hg concentrationin different samples varied from 53% to 95% with the lowest inthe liver and highest in muscle and brain samples. The highpercentage found in brain samples would probably make thebrainmore vulnerable thanother organs/tissues tohigher levelsof Hg exposure. Log total Hg concentration in fur/hair samplesare positively correlated with log organic Hg levels in liver andkidney samples, suggesting that fur/hair samples could be auseful monitoring tissue for environmental contamination.Among different factors affecting total Hg levels in mink liver,such as sex, body mass and age, only age has significant in-fluence on Hg levels. An increase in total Hg levels with agesuggests a cumulative effect of Hg contamination over time.Interestingly, log liver total Hg concentration in mink wasgreatest in areas with moraine deposits and least in areas withrich clay deposits. However, the effect of soil deposits could beconfounded by uneven deposition of anthropogenic Hg. On theother hand, total Hg levels in mink livers decrease with the

Table 4 – Hg concentrations in freshwater mammals

Species Region Location Tissue Sample period Total Hg Level(µg/g, ww)

Reference

Mink Mustela vison Québec James Bay territory Fur 1993–1994 26.9±1.70 Fortin et al. (2001)Liver 1993–1994 4.36±1.82 Fortin et al. (2001)Kidney 1993–1994 2.19±1.74 Fortin et al. (2001)Muscle 1993–1994 2.14±1.66 Fortin et al. (2001)Brain 1993–1994 0.82±1.78 Fortin et al. (2001)

Northwest Territories Inuvik Liver 1991–92, 1994–95 1.45±0.11 Poole et al. (1998)Fort Good Hope Liver 1991–92, 1994–95 2.17±0.29 Poole et al. (1998)Fort Liard Liver 1991–92, 1994–95 1.00±0.16 Poole et al. (1998)Ford Providence Liver 1991–92, 1994–95 1.02±0.21 Poole et al. (1998)Fort Rae Liver 1991–92, 1994–95 3.30±0.65 Poole et al. (1998)Fort Resolution Liver 1991–92, 1994–95 0.91±0.32 Poole et al. (1998)Fort Smith Liver 1991–92, 1994–95 1.30±0.28 Poole et al. (1998)

Northwest Territories Inuvik Liver 1992 1.16 Poole et al. (1995)Inuvik Liver 1993 1.84 Poole et al. (1995)Fort Good Hope Liver 1991–1993 2.17 Poole et al. (1995)Fort Rae Liver 1991–1993 3.30 Poole et al. (1995)Fort Liard Liver 1991–1993 1.45 Poole et al. (1995)Fort Smith Liver 1991–1993 2.44 Poole et al. (1995)

Organic Hg(µg/g, w.w)

Québec James Bay territory Liver 1993–1994 2.19±1.82 Fortin et al. (2001)Kidney 1993–1994 1.29±2.09 Fortin et al. (2001)Muscle 1993–1994 1.70±1.78 Fortin et al. (2001)Brain 1993–1994 0.76±1.82 Fortin et al. (2001)

River otter Lutracanadensis

Québec James Bay territory Fur 1993–1994 19.5±1.41 Fortin et al. (2001)Liver 1993–1994 3.72±1.62 Fortin et al. (2001)Kidney 1993–1994 3.24±1.29 Fortin et al. (2001)Muscle 1993–1994 1.29±1.48 Fortin et al. (2001)Brain 1993–1994 0.72±1.62 Fortin et al. (2001)

Organic Hg(µg/g, w.w)

Liver 1993–1994 1.78±1.48 Fortin et al. (2001)Kidney 1993–1994 1.82±1.41 Fortin et al. (2001)Muscle 1993–1994 0.91±1.74 Fortin et al. (2001)Brain 1993–1994 0.42±1.78 Fortin et al. (2001)

Total Hg(µg/g, w.w)

Ontario Georgian Bay Liver – 3.0±2.2 Wren (1984)Kidney – 1.4±0.3 Wren (1984)Intestine – 0.4±0.2 Wren (1984)Muscle – 0.9±0.2 Wren (1984)

Beaver Castor canadensis Liver – 0.03±0.00 Wren (1984)Kidney – 0.03±0.00 Wren (1984)Muscle – 0.03±0.00 Wren (1984)

Raccoon Procyon lotor 0.03±0.00 Wren (1984)Liver – 4.5±3.5 Wren (1984)Kidney – 1.1±0.4 Wren (1984)Muscle – 0.3±0.1 Wren (1984)


increase of the distance from industrial centers, indicating thatlocal industrial activities could be responsible for the elevatedHg levels in livers.

Mean totalHg concentration inmink livers (3.7±3.9 µg/g,wetwt.) reportedbyFortinet al. (2001)washigher than thedata fromwestern Northwest Territories (NWT), Canada during periods of1991–1993 (community means: 0.91–3.30 µg/g wet wt.) (Pooleet al., 1995, 1998). The low Hg burden from NWT could beinfluenced by many factors, including acidity of the water,calcium and selenium levels, bedrock and soil composition,recent damming, proportion of fish in the diet, and proximity to

sources of Hg pollution. Besides, no obvious geographical trendto the pattern of Hg concentrations was observed.

4. Hg concentrations in marine ecosystem

4.1. Mercury levels in Arctic Ocean and marine sediments

Concentrations of four species of Hg in seawater: gaseouselemental mercury (GEM), dimethyl mercury (DiMeHg),methyl mercury (MeHg), and total mercury (THg) were


measured in Canadian Arctic Oceans during the Amundsenicebreaker fall 2005 cruise (Kirk et al., 2005). Seawater collectedunder the sea ice at Resolute and Churchill had highconcentrations of GEM (~150 pg/L and~100 pg/L, respectively).During spring ice out, GEM trapped under the ice fluxes off thewater surface, making the ocean a source of GEM to theatmosphere. At Resolute, concentrations of both MeHg andDiMeHg in seawater under the sea ice were also extremelyhigh at ~0.1 ng/L and ~11 pg/L, whereas on Hudson Bay,concentrations were only ~0.02 ng/L and ~3 pg/L, respectively.The Higher MeHg and DiMeHg levels observed at Resolute (inthe high Arctic Ocean) than those in Hudson Bay could beattributed to different depth of the ocean. Methylation of Hgmay occur in anaerobic zones of marine environments andthat MeHg is brought to upper layers by oceanmixing, makingit available for uptake by marine organisms at all oceandepths. On the other hand, Hudson Bay is a relatively shallowwater body and does not have large anaerobic regions. Thus,MeHg and DiMeHg are produced more efficiently in oceanswith large anaerobic zones of marine environments.

Compared to freshwater ecosystems, the informationabout Hg content in Arctic marine sediments is scarcer inliterature. Mason et al. (1994) have demonstrated that theconcentration of Hg in the ocean has increased by a factor of 3as a result of anthropogenic emissions. These and otherobservations concerning Hg levels in biota and ice from theArctic suggest also an increase of the Hg content in marinesediments from this earth region (Asmund and Nielson, 2000).

A few works have been published about Hg content insediment cores from different Arctic regions (e.g. Gobeilet al.,1999; Galasso et al., 2000; Asmund and Nielson, 2000).As observed for freshwater sediments the majority of theworks reported an increase of Hg levels in surface sedi-ments, decreasing linearly or exponentially with depth. VanDalen and Nolting (1992) reported Hg concentrations varyingbetween 0.03 and 0.14 ng/g (dw) in sediments from theLaptev Sea (Russia); Gobeil and MacDonald (1997) deter-mined Hg levels varying between 0.05 and 0.15 ng/g (dw) inthe North Pole and; Gobeil et al. (1999) reported levels bet-ween 10 and 113 ng/g (dw) for several cores collect in deepsea sediments from the Arctic ocean. More recently, Hgsediment levels varying between 195–875 ng/g from RussianArctic (Kara Sea) and from 0.05 to 0.3 ng/g from Greenlandmarine sediments have been reported by Asmund and Niel-son (2000); Galasso et al. (2000), respectively. Riget et al.(2000c) also reported Hg levels in the Greenland marinesediments collected at four regions in 1994–95 (Table 1).Geometric mean concentrations of Hg varied from 0.045 to0.033 µg/g (dw), with the highest concentration in Nanortalikand the lowest level in Avanersuaq. Hg concentrationsreported by Riget et al. (2000c) were in a good agreementwith the level by Asmund and Nielson (2000). Additionally,these researchers investigated Hg profiles in twenty marinesediment cores from Greenland in 1985–1994. Hg levels insediment cores collected at different locations varied from0.003 to 0.15 µg/g, with higher Hg concentrations in theupper centimeters of most cores. From linear regression oflnHg vs. age of the sediment determined by the 210Pbmethod, Asmund and Nielson (2000) found that Hg levelsdecreased with depth in the sediment with various degrees

of significance. The increase of the Hg flux during the last100 years is roughly a doubling. The increase may be ofanthropogenic origin as it is restricted to the last 100 years.

In spite of the importance of these findings, Gobeil et al.(1999) make clear that marine sediments, particularly fromthe deep ocean, are poor locations to evaluate moderncontaminant trends because of the very low sedimentationrates. For Hg enrichments in surface sediments to beattributed to modern contamination, it is necessary that Hghave a strong affinity for particles and that, once sedimented,it does not undergo migration by remobilization duringdiagenesis (Young et al., 1973; Gobeil et al., 1999). By analyz-ing sediments from seven cores from the Arctic Ocean, Gobeilet al. (1999) have pointed that Hg profiles in these cores havebeen produced by Hg redistribution during diagenesis. In allseven cores the authors observed similarities between Hgand the reactive Fe profiles, implying that a portion of thetotal Hg deposit is recycling along with Fe during redoxchanges.

4.2. Hg concentrations in marine biota

4.2.1. InvertebratesMarine invertebrates provide a link between phytoplanktonand fish, seabirds and mammals in Arctic marine food webs.These invertebrates are important in the transfer of carbonand nutrients but also contaminants to organisms at uppertrophic levels (Fig. 3). Knowledge of the trends and dynamicsof Hg levels in marine invertebrates and fish are important forunderstanding the trends of contaminants in Arctic marineecosystems (Fisk et al., 2003).

Low concentrations of Hg (0.03 µg/g, wet wt.) were found insamples of blue mussels (Mytilus edulis) from six communitiesin the Hudson Bay, Hudson Strait andUngava Bay area (Doidgeet al., 1993), which agreed well with the results by Muir et al.(2000). These authors reported total Hg levels in the range of0.01–0.03 µg/g (wet wt.) from Labrador and Nunavik during asite survey in 1998–1999 (Table 5). Organic Hg consistsapproximately 54% of total Hg, which could be all in theform ofMeHg. No geographic trend of Hg levels inmusselswasobserved. Similarly, lowHg levels were also reported in tissuesfrom 12 species of invertebrates and particulate organicmatter (POM) from Lancaster Sound, Northwest Territories(1998–1990) in a food web survey by Atwell et al. (1998). TotalHg in invertebrates varied from 0.03 to 0.18 µg/g (dw) with thelowest level in sea star (Crossaster papposus) and the highestlevel in sea urchin (Stronglocentrotus sp.). Non-detectable levelof Hg in POM was reported (b0.02 µg/g dw).

In a more recent paper, Campbell et al. (2005) reported totalHg and MeHg levels in the tissues of primary and zooplanktonsamples collected in 1998 in the North water Polynya (Table 5).Total Hg in ice algaewas fairly low (0.003 µg/gwetwt., n=1) andno MeHg was detected. On the other hand, total Hg and MeHgin four species of zooplankton varied in the range of 0.011±0.003–0.025±0.017 µg/g wet wt. and 2.4±1.3–28.3±0.8 ng/gwet wt., respectively. The percentage of MeHg in terms oftotal Hg varied widely among species, with the lowest per-centage of 7.5% in copepod zooplankton C. hyperboreus and thehighest levels of 100% in the predatory zooplankton Themistolibellula and Mysis oculata.


4.2.2. Marine fish and anadromous fishArctic char is one of the most studied fish species for moni-toring purposes because of it is a top predator in lakes/oceans.

Fig. 3 –Arctic's terrestrial (A), tundra ponds (B), lakes (C) and marbioaccumulation of Hg (adapted from AMAP, 1997, 2002).

They occur both as lake residents living their whole life infreshwater and in an anadromous formmigrating intomarinewaters during summer for eating (Riget et al., 2000a). Hg levels

ine (D) food web structure which is influencing the fate and

Table 5 –Mercury in marine or anadromous fish and marine invertebrates from Arctic

Species Region Location Tissue Age Sampleperiod

Hg Level(µg/g, dw)

Reference

PrimaryParticulateorganic matter

Lanscaster Sound NWT Muscle – 1988–1990 b0.02 Atwell et al. (1998)

Ice age Baffin Bay Northwater Polynya Whole – 1998 0.003 Campbell et al. (2005)

InvertebratesBlue mussel(Mytilus edulis)

N. Labrador Makkovik Whole body – 1998–1999 0.020 Muir et al. (2000)N. Labrador Nain Whole body – 1998–1999 0.010 Muir et al. (2000)Nunavik (QC) Kangiqsualujjuaq Whole body – 1998–1999 0.030 Muir et al. (2000)Nunavik (QC) Quaqtaq Whole body – 1998–1999 0.010 Muir et al. (2000)Nunavik (QC) Kuujjuaq Whole body – 1998–1999 0.010 Muir et al. (2000)Nunavik (QC) Deception Bay Whole body – 1998–1999 0.020 Muir et al. (2000)Nunavik (QC) Nain Whole body – 1998–1999 0.010 Muir et al. (2000)

Scallops(Placopecten magellanicus)

Labrador Nain Muscle 19 (9–30) 1999 0.03 (0.01–0.05) Muir et al. (2000)Labrador Nain Gut 19 (9–30) 1999 0.05 (0.03–0.07) Muir et al. (2000)Labrador Nain Gonad 19 (9–30) 1999 0.02 (0.00–0.06) Muir et al. (2000)

Mya truncata Lanscaster Sound NWT Muscle – 1988–1990 0.07±0.01 Atwell et al. (1998)Hiatella arctica Lanscaster Sound NWT Muscle – 1988–1990 0.15±0.02 Atwell et al. (1998)Unidentified sea cucumber Lanscaster Sound NWT Muscle – 1988–1990 0.04±0.00 Atwell et al. (1998)Strongylocentrotus sp. Lanscaster Sound NWT Muscle – 1988–1990 0.18 Atwell et al. (1998)Calanus hyperboreus Lanscaster Sound NWT Muscle – 1988–1990 0.06±0.01 Atwell et al. (1998)Macoma calcarea Lanscaster Sound NWT Muscle – 1988–1990 0.09 Atwell et al. (1998)Leptastarias sp. Lanscaster Sound NWT Muscle – 1988–1990 0.05±0.01 Atwell et al. (1998)Onisimus glacialis Lanscaster Sound NWT Muscle – 1988–1990 0.10 Atwell et al. (1998)Parathemisto libellula Lanscaster Sound NWT Muscle – 1988–1990 0.06 Atwell et al. (1998)Mertensia ovum Lanscaster Sound NWT Muscle – 1988–1990 0.07 Atwell et al. (1998)Anenome urticina Lanscaster Sound NWT Muscle – 1988–1990 0.06 Atwell et al. (1998)Crossaster papposus Lanscaster Sound NWT Muscle – 1988–1990 0.03 Atwell et al. (1998)Clione limacina Greenland Northwest Greenland Whole – 1987 b0.005 (w.w) Dietz et al. (1996)Iceland scallop(Chlamys islandica)

Greenland Northwest Greenland Soft tissue −/b8 g 1984 0.016 (w.w) Dietz et al. (1996)

Iceland scallop(Chlamys islandica)

Greenland Northwest Greenland Soft tissue −/N8 g 1984 0.020 (w.w) Dietz et al. (1996)

Green crenella(Musculus discors)

Greenland Northwest Greenland Soft tissue – 1984 0.013 (w.w) Dietz et al. (1996)

Cockle(Serripes groenlandicus)

Greenland Northwest Greenland Soft tissue – 1984 0.011 (w.w) Dietz et al. (1996)

Calanus finmarchicus Greenland Central WestGreenland

Whole – 1987 b0.005 (w.w) Dietz et al. (1996)

Parathemisto libellua Greenland Northwest Greenland Whole – 1987 b0.005 (w.w) Dietz et al. (1996)Fuphausiaces sp. Greenland Central West

Greenland NorthWhole – 1987 b0.005 (w.w) Dietz et al. (1996)

Deep sea prawn(Pandalus borealis)

Greenland Northwest Greenland Whole −/N5 g 1987 0.119 (w.w) Dietz et al. (1996)

Sabinea sp. Greenland Central WestGreenland North

Whole −/b5 g 1987 0.056 (w.w) Dietz et al. (1996)

Greenland Central WestGreenland North

Whole −/N5 g 1987 0.0146 (w.w) Dietz et al. (1996)

Greenland Central WestGreenland South

Whole −/b5 g 1987 0.032 (w.w) Dietz et al. (1996)

Greenland Central WestGreenland South

Whole −/N5 g 1987 0.070 (w.w) Dietz et al. (1996)

Greenland Southwest Greenland Whole −/b5 g 1985 0.078 (w.w) Dietz et al. (1996)Greenland Southwest Greenland Whole −/N5 g 1985 0.189 (w.w) Dietz et al. (1996)Greenland Southwest Greenland Whole −/b5 g 1985 0.149 (w.w) Dietz et al. (1996)Greenland Central West

Greenland NorthWhole −/N5 g 1987 0.055 (w.w) Dietz et al. (1996)

Greenland Central East Greenland Whole −/N5 g 1985 0.327 (w.w) Dietz et al. (1996)Copepod(Calanus hyperboreus)

Baffin Bay Northwater Polynya Whole – 1998 0.025±0.017 (w.w) Campbell et al. (2005)

Mixed zooplankton Baffin Bay Northwater Polynya Whole – 1998 0.006±0.002 (w.w) Campbell et al. (2005)Amphipod(Themisto libellula)


(continued on next page)(continued on next page)


Table 5 (continued)


Hg Level(µg/g, dw)

Reference

InvertebratesMysid (Mysis oculata) Baffin Bay Northwater Polynya Whole – 1998 0.011±0.003 (w.w) Campbell et al. (2005)

MeHg(ng/g, w.w)

Copepod(Calanus hyperboreus)


Mixed zooplankton Baffin Bay Northwater Polynya Whole – 1998 4.0±3.0 (w.w) Campbell et al. (2005)Amphipod(Themisto libellula)


Mysid Baffin Bay Northwater Polynya Whole – 1998 14.3±4.8 (w.w) Campbell et al. (2005)Marine fish Labrador Hopedale Muscle 7±1 1998 0.033±0.005 Muir et al. (1999a,b)Arctic charSalvelinus alpinus

Hopedale Muscle 6.7±1.9 1999 0.03±0.01 Muir et al. (2000)Makkovik Muscle 6.6±1.6 1999 0.03±0.01 Muir et al. (2000)Nain Muscle – 1997 0.027±0.007 Muir et al. (2000)Nain Muscle 7.8±1.5 1998 0.03±0.01 Muir et al. (1999a,b)

N. Québec Kangirsuk Muscle 7.9±1.5 1998 0.032±0.026 Muir et al. (1999a,b)Tasuijaq Muscle 8.5±1.5 1998 0.040±0.011 Muir et al. (1999a,b)Quaqtaq Muscle 8.9±2.4 1998 0.072±0.035 Muir et al. (1999a,b)Povungnituk Muscle 8.1±0.94 1998 0.044±0.009 Muir et al. (1999a,b)Kaniqsujuaq Muscle 11.3±1.53 1998–99 0.04±0.00 Muir et al. (1999a,b)

Greenland Qaqortoq, River I Muscle 6.1 1994–1995 0.05±0.012 Riget et al. (2000a)Qaqortoq, River I Muscle 6.2 1994–1995 0.04±0.012 Riget et al. (2000a)Qaqortoq, River I Muscle 6.7 1994–1995 0.045±0.014 Riget et al. (2000a)

Boreogadus saida NWT Lancaster Sound Muscle – 0.19±0.03 Atwell et al. (1998)Icelus bicornis NWT Lancaster Sound Muscle – 0.24±0.04 Atwell et al. (1998)Polar cod Northwater

PolnyaBaffin Bay Whole – 1998 0.04 (w.w) Campbell et al. (2005)

Boreogadus saida Liver – 1998 0.015–0.002 (w.w) Campbell et al. (2005)Greenland Northwest Greenland Liver −/14.9 cm 1987 0.016 (w.w) Dietz et al. (1996)

Muscle −/14.8 cm 1987 0.039 (w.w) Dietz et al. (1996)Southwest Greenland Liver −/12.6 cm 1985 b0.005 (w.w.) Dietz et al. (1996)

Muscle −/12.6 cm 1985 0.011 (w.w) Dietz et al. (1996)Arctic codArctogadus glacialis

Greenland Northwest Greenland Liver −/21 cm 1987 0.187 (w.w.) Dietz et al. (1996)Muscle −/21 cm 1987 0.278 (w.w.) Dietz et al. (1996)

Arctic eelpout(Lycodes reticulates)

Central EastGreenland

Liver −/19.4 cm 1985 0.569 (w.w.) Dietz et al. (1996)

Twohorn sculpinIcilus bicornis


Muscle −/6.8 cm 1985 0.082 (w.w.) Dietz et al. (1996)

Shorthorn sculpinMyoxycephalussorpius

Northwest Greenland Liver −/27.1 cm 1985 0.024 (w.w.) Dietz et al. (1996)Muscle −/27.1 cm 1985 0.065 (w.w) Dietz et al. (1996)

Central WestGreenland South

Liver −/25.6 cm 1986–87 0.014 (w.w) Dietz et al. (1996)Muscle −/25.6 cm 1986–87 0.030 (w.w.) Dietz et al. (1996)

Southwest Greenland Muscle −/26.1 cm 1985 0.081 (w.w) Dietz et al. (1996)Central EastGreenland

Liver 16.9–23 cm 1985 0.031 (w.w) Dietz et al. (1996)Muscle 22.6 cm 1985 0.059 (w.w.) Dietz et al. (1996)Kidney 25.2 cm 1985 0.027 (w.w.) Dietz et al. (1996)

Greenland sharks Baffin Island Cumberland Sound Liver – 1999–2000 0.52±0.11 Fisk et al. (2003)Somniosus microcephalus Cumberland Sound Liver – 1999–2000 0.47±0.06 Fisk et al. (2003)


inmuscle tissue of sea-run char inGreenlandwere determinedina studybyRiget et al. (2000a) (Table 5). Themeasured totalHglevels were in the range of 0.04–0.05 µg/g (wet. wt.). Approxi-mately 72% of total Hgwas in the form ofMeHg. Total Hg levelsreported by Riget et al. (2000a) agree with the concentrationsfound in sea-run char fromeight communities in Labrador andthe Ungava/Hudson Strait region of Nunavik (0.032–0.040 µg/g,wet. wt.) (Muir et al., 2000). In a food web study, Atwell et al.(1998) measured total Hg in muscle tissue of Arctic cod(Boregadus saida) and twohorn sculpin (Icelus biocornis) fromthe Lancaster Sound region. Concentrations of Hg in the Arcticcod and two horn sculpin (0.19±0.03 µg/g drywt.; 0.24±0.04 µg/g dw) are similar to those observed in Arctic char if concentra-tions are converted to wet weight.

Recently, data on trace metals have been produced forlivers of Greenland shark (Somniosus microceplalus), the onlyshark species that routinely inhabits Arctic waters (Fisk et al.,2003). Greenland shark samples (n=24) were collected inCumberland Sound in 1999 and 2000 (Table 5). Concentrationsof total Hg in Greenland shark liver (0.49±0.06 ug/g) weremorethan an order of magnitude greater than sea-run Arctic charmuscle collected in the same region, possibly due to its hightrophic level in the food web (Fisk et al., 2002).

The possible factors influencing total Hg levels in Arcticchar were investigated using Pearson's correlation analysis byRiget et al. (2000a). These researchers found that fish lengthand dry weight rather than age were the most important co-variables in determining Hg levels in muscle of Arctic char.


4.2.3. Sea ducks and seabirdsSea ducks (Mergini) have widely used as bio-monitors ofinshore marine pollution in North America and Europe(Wayland et al., 2005). King eiders (Somateria spectabilis) andcommon eiders (S. mollissima) are sea ducks found principallyin arctic and sub arctic coastal areas (Wayland et al., 2001).Population declines in several species of sea ducks have raisedconsiderable concerns on the influence of environmentalcontamination, particularly heavy metals (Trust et al., 2000;Wayland et al., 2001, 2005). Several studies have examined Hglevels in sea ducks for risk assessments and bio-monitoringanalyses (Trust et al., 2000; Wayland et al., 1999, 2000, 2001,2005) (Table 6). In a study byWayland et al. (2001), total Hg andorganic Hg concentrations were measured in liver of commonand king eiders collected at three locations in the CanadianArctic. Total and organic Hg levels ranged from 0.7 to 4.5 and0.5–3.8 µg/g (dw), respectively. A clear species differences withhigher total and organic Hg levels in king eider males thanfemale sea ducks was observed. Foraging habitat and dietarydifferences between these two sea ducksmay partially explaindifferent Hg levels (Wayland et al., 2001). Both common andking eiders feed heavily on mussels. However, king eidersconsume a more diversified diet including not only musselsbut also echinoderms and other benthic invertebrates(Bustnes and Erikstad, 1988; Frimer, 1997). Furthermore, kingeiders are found to prefer feeding in relatively deep water overbare, cobble, or coralline algal substrates, whereas commoneiders preferred shallower water dominated by kelp beds orsandy substrates interspersed with rock (Bustnes and Lønne,1997). Hg concentrations in marine invertebrates vary amongspecies, water depth and substrate characteristics (Phillips,1980). Thus, it is possible that benthic organisms accumulateslightly higher concentrations of Hg in habitats preferred byking eiders compared to those preferred by common eiders.Moreover, king eiders, being smaller-bodied than commoneiders (Dunning, 1993), may have higher energy requirements,and hence energy intake rates, per unit body mass (King andFarner, 1961). This may result in greater dietary contaminantintake per unit body mass in king eiders. Organic Hg accountsfor 69–88% of total Hg levels in liver and significant correla-tions were reported between these two Hg species. Total Hglevels in the study by Wayland et al. (2001) were higher thanthose data measured in tissues of spectacled eiders (Somateriafischeri) collected near St. Lawrence Island, Alaska (Trust et al.,2000). Hg levels in liver and kidney of spectacled eiders were1.13±0.05 (range: 0.80–1.69 µg/g dw) and 0.73±0.2 (range: 0.57–1.14 µg/g dw.), respectively. Levels of Hg in liver wereapproximately 55% higher than in kidney.

Hg levels in Arctic seabirds have been examined in severalstudies and the results are summarized in Table 6. Hg concen-trations in the tissues of seabirds from Arctic seabirds col-lected during 1989–1999 sampling period in Northern BaffinBay (Canada)were inter-species compared (Braune, 2000, 2001,Fisk et al., 2003). Among eight species of Arctic seabirds, totalHg levels in the liver of Northern fulmars (Fulmarus glacialis)exhibited the highest concentrations (2.90±0.84–3.92±0.59 µg/g dw) in comparison with others (0.26±0.02–1.94 µg/g dw).Northern fulmars are a very long-lived bird and high Hg levelsin liver could be related to Hg accumulation with age. Thehighest Hg burdens in muscle of seabirds were recorded in

Glaucous gull (Larus hyperboreus, 0.79±0.16–0.81±0.09 µg/gdw), which could be attributed to higher trophic position ofGlaucous gull. On the other hand, total Hg levels in the liverandmuscle of Dovekie or Little auk (Alle alle) were significantlylower than other seabirds (0.26±0.02–0.28±0.05 µg/g dw inliver, 0.076±0.012–0.078±0.009 µg/g dw in muscle). In general,Hg concentrations in liver were greater than muscle across allspecies of sea birds and concentrations of total Hg in liver andmuscle varied between genders. Similar to the pattern of totalHg, MeHg concentrations in glaucous gull were highest and indovekie were lowest, and levels of MeHg strongly correlated totrophic position.

Hg levels have also been monitored in Canadian seabirdeggs. Interestingly, Hg concentrations in glaucous gull eggswere significantly higher than the levels measured in otherspecies of seabirds, a difference attributed to the highertrophic level of glaucous gull (Braune, 2001, Fisk et al., 2003).On the other hand, Fisk et al. (2003) found that Hg concentra-tions in eggs of black guillemots were higher than the levelsdetected in eggs of black-legged kittiwakes. Migration differ-ence in geographic locations might be a good explanation forthe different levels of Hg in eggs of seabirds. It is know thatblack guillemots are most likely to remain resident in theArctic, while black-legged kittiwakes migrate south along theeastern seaboard as far as Florida (Fisk et al., 2003). Hg has atendency to accumulate at higher latitudes (Barrie et al., 1997);therefore, higher levels of Hg in eggs of black guillemots thanblack-legged kittiwakes are expected. However, no influenceof migration difference on Hg levels in the tissues of adultseabirds from northern Baffin Bay was observed. It is possiblethat Hg concentrations in eggs may not reflect levels in adulttissues.

Hg concentrations in a top-level predator, bald eagle(Haliaeetus leucocephalus), were investigated by Stout andTrust (2002). Bald eagles prey primarily on fish and birds(Anthony et al., 1993) and are believed to be especiallyvulnerable to many contaminants, thus can be used assentinel species for contaminated areas (Holl and Cairns,1995). Stout and Trust (2002) reported Hg levels in tissues of 26bald eagles collected between 1993 and 1998 fromAdak Island,Alaska. Hg concentrations in livers of bald eagle fluctuatedfrom 1.70 to 17.5 with a geometric mean of 7.10 µg/g (dw). Thegeometric level of Hg in kidney was nearly two times as thevalue in liver (geometric mean: 14.6 µg/g, dw, range: 3.19–68.4 µg/g, dw). Mean Hg concentrations observed in the studyby Stout and Trust (2002) were similar to those detected in baldeagle carcasses from Florida (Wood et al., 1996). However,these values were significantly higher than other seabirds atlower trophic positions (Table 6), suggesting bio-magnificationat higher trophic levels (details see Section 6). No effect of ageor sex onHg levels in the tissues of bald eagleswas detected byStout and Trust (2002).

Seabirds generally exhibit higher Hg concentrations thanterrestrial birds because of the higher Hg burdens encounteredin marine ecosystems (Thompson, 1996; Fisk et al., 2003).Recent studies have shown that the predominant form of Hgfound in seabirds is inorganic, suggesting that biotransforma-tion of ingested MeHg is an important mechanism by whichlong-lived and slow-moulting seabirds avoid the toxic effectsof accumulating large quantities of MeHg (Thompson, 1996;

Table 6 – Total Hg concentrations in sea ducks and seabirds

Species Region Location Tissue Sex Sampleperiod

Hg Level(µg/g, dw)

Reference

Sea ducksCommon Eider(Somateria mollissima)

Northwest Territories Lancaster Sound Muscle – 1988–1990 0.38±0.02 Atwell et al. (1998)Southampton, NT East Bay Migratory

Bird SanctuaryLiver F 1997 1.6 (0.9–3.7)/ 1.4 (0.9–3.2)⁎ Wayland et al. (2001)

Victoria Island, NWT Village of Holman Liver F 1997 1.5 (1.0–2.2)/ 1.3 (0.8–1.9)⁎ Wayland et al. (2001)Nauavut Belcher Islands Liver F 1997 1.3 (0.7–2.8)/ 0.9 (0.5–1.7)⁎ Wayland et al. (2001)Greenland Central West

GreenlandLiver – 1982–88 0.046 (w.w) Dietz et al. (1996)

North Kidney – 1982–88 0.105 (w.w) Dietz et al. (1996)Muscle – 1982–88 0.016 (w.w.) Dietz et al. (1996)Liver – 1983–91 0.644 (w.w) Dietz et al. (1996)Kidney – 1983–91 0.218 (w.w) Dietz et al. (1996)Muscle – 1983–91 0.100 (w.w) Dietz et al. (1996)

SouthWest Greenland Liver – 1984–88 0.784 (w.w) Dietz et al. (1996)Kidney – 1984–88 0.302 (w.w) Dietz et al. (1996)Muscle – 1984–88 0.166 (w.w) Dietz et al. (1996)

King eiders(Somateria spectabilis)

Southampton, NT East Bay Migratory Liver F 1997 2.1 (1.5–2.5)/ 1.7 (1.2–2.2)⁎ Wayland et al. (2001)Bird Sanctuary M 1997 3.8 (3.4–4.4) /3.0 (2.6–3.8)⁎ Wayland et al. (2001)

Victoria Island, NWT Village of Holman Liver F 1997 1.7 (1.0–2.5)/1.4 (90.8–2.0)⁎

Wayland et al. (2001)

Spectacled Eider(Somateria fischeri)

Alaska, USA St. Lawrence Island Liver M 1995 1.13±0.05 (0.80–1.69) Trust and al. (2000)Alaska, USA St. Lawrence Island Kidney M 1995 0.73±0.2 (0.57–1.14) Trust et al. (2000)Greenland Central West Liver – 1983–86 0.44 (w.w.) Dietz et al. (1996)

Greenland North Kidney – 1983–86 0.276 (w.w.) Dietz et al. (1996)Muscle – 1983–86 0.109 (w.w.) Dietz et al. (1996)

SeabirdsNothern fulmars(Fulmarus glacialis)

Northern Baffin Bay Nothwater polynaya Liver F 1998–1999 2.9±0.84 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Liver M 1998–1999 3.92±0.59 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle F 1998–1999 0.35±0.07 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle M 1998–1999 0.44±0.057 Fisk et al. (2003)Northwest Territories Lancaster Sound Muscle – 1988–1990 0.75±0.23 Atwell et al. (1998)Greenland Northwest Greenland Liver – 1984–85 2.25 (w.w) Dietz et al. (1996)

Kidney – 0.791 (w.w.) Dietz et al. (1996)Muscle – 0.313 (w.w.) Dietz et al. (1996)

Black-leggedkittiwakes(Rissa tridactyla)

Lancaster Sound Prince Leopold Island Eggs – 1998 0.6±0.1 Braune (2000, 2001)Northwest Territories Lancaster Sound Muscle – 1988–1990 1.93±1.14 Atwell et al. (1998)Northern Baffin Bay Northwater polynya Liver F 1998–1999 1.08±0.25 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Liver M 1998–1999 1.02±0.14 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle F 1998–1999 0.27±0.03 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle M 1998–1999 0.33±0.047 Fisk et al. (2003)Greenland Northwest Greenland Liver – 1984–85 0.967 (w.w.) Dietz et al. (1996)

Muscle – 1984–85 0.246 (w.w.) Dietz et al. (1996)Thick-billed murres(Uria lomvia)

Lancaster Sound Prince Leopold Island Eggs – 1998 1.2±0.1 Braune (2000, 2001)Northwest Territories Lancaster Sound Muscle – 1988–1990 1.79±0.78 Atwell et al. (1998)Northern Baffin Bay Northwater polynya Liver F 1998–1999 0.99±0.06 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Liver M 1998–1999 1.16±0.15 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle F 1998–1999 0.33±0.04 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle M 1998–1999 0.27±0.03 Fisk et al. (2003)

Black or Commonguillemot(Cepphus grille)

Lancaster Sound Prince Leopold Island Eggs – 1998 1.8±0.1 Braune (2000, 2001)Northwest Territories Lancaster Sound Muscle – 1988–1990 1.57±0.65 Atwell et al. (1998)Northern Baffin Bay Northwater polynya Liver F 1998–1999 0.85±0.067 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Liver M 1998–1999 1.49±0.19 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle F 1998–1999 0.27±0.036 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle M 1998–1999 0.41±0.022 Fisk et al. (2003)


Greenland Northwest Greenland Liver – 1984–85 0.637 (w.w.) Dietz et al. (1996)Kidney – 1984–85 0.417 (w.w.) Dietz et al. (1996)Muscle – 1984–85 0.227 (w.w.) Dietz et al. (1996)

Central west Liver – 1985–91 0.595 (w.w.) Dietz et al. (1996)Greenland North Kidney – 1985–91 0.438 (w.w.) Dietz et al. (1996)

Muscle – 1985–91 0.170 (w.w.) Dietz et al. (1996)Southwest Greenland Liver – 1986 0.497 (w.w.) Dietz et al. (1996)

Kidney – 1986 0.402 (w.w.) Dietz et al. (1996)Muscle – 1986 0.150 (w.w.) Dietz et al. (1996)


Table 6 (continued)

Species Region Location Tissue Sex Sampleperiod

Hg Level(µg/g, dw)

Reference


Central East Greenland Liver – 1986 0.505 (w.w.) Dietz et al. (1996)Muscle – 1986 0.123 (w.w.) Dietz et al. (1996)

Dovekie or Littleauk (Alle alle)

Northwest Territories Lancaster Sound Muscle – 1988–1990 0.33 Atwell et al. (1998)Northern Baffin Bay Northwater polynya Liver F 1998–1999 0.26±0.018 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Liver M 1998–1999 0.28±0.048 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle F 1998–1999 0.078±0.009 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle M 1998–1999 0.076±0.012 Fisk et al. (2003)Greenland Central East Greenland Liver – 1986 0.552 (w.w.) Dietz et al. (1996)

Kidney – 1986 0.314 (w.w.) Dietz et al. (1996)Muscle – 1986 0.135 (w.w.) Dietz et al. (1996)

Glaucous gull(Larus hyperboreus)

Lancaster Sound Prince Leopold Island Eggs – 1998 3.4±0.1 Braune (2000, 2001)Northwest Territories Lancaster Sound Muscle – 1988–1990 1.86±0.68 Atwell et al. (1998)Northern Baffin Bay Northwater polynya Muscle – 1998–1999 0.79±0.16 Fisk et al. (2003)Northern Baffin Bay Nothwater polynaya Muscle – 1998–1999 0.81±0.087 Fisk et al. (2003)Greenland Southwest Greenland Liver – 1986 0.680 (w.w) Dietz et al. (1996)

Kidney – 1986 0.613 (w.w) Dietz et al. (1996)Muscle – 1986 0.148 (w.w) Dietz et al. (1996)

Northwest Greenland Liver – 1986 2.67 (w.w) Dietz et al. (1996)Kidney – 1986 2.05 (w.w) Dietz et al. (1996)Muscle – 1986 0.665 (w.w) Dietz et al. (1996)

Ivory gull(Pagophila eburnean)

Northern Baffin Bay Northwater polynya Muscle F 1998–1999 0.20±0.03 Fisk et al. (2003)Northern Baffin Bay Northwater polynaya Muscle M 1998–1999 0.22±0.045 Fisk et al. (2003)Greenland Northwest Greenland Liver – 1984–85 0.739 Dietz et al. (1996)

Muscle – 1984–85 0.172 Dietz et al. (1996)Thyler gulls(Larus thayeri)

Northern Baffin Bay Northwater polynya Liver M 1998–1999 1.94 Fisk et al. (2003)Northern Baffin Bay Northwater polynaya Muscle M 1998–1999 0.48 Fisk et al. (2003)

Arctic tern(Sterna paradisaea)

Northwest Territories Lancaster Sound Muscle – 1988–1990 0.83±0.21 Atwell et al. (1998)

Brunnichs guillemot Greenland Northwest Greenland Liver – 1984–85 0.637 (w.w.) Dietz et al. (1996)Uria lomvia Kidney – 1984–85 0.417 (w.w.) Dietz et al. (1996)

Muscle – 1984–85 0.227 (w.w.) Dietz et al. (1996)Bald eagle (Haliaeetusleucocephalus)

Alaska Adak Island Liver – 1993–1998 7.10 (1.70–17.5) Stout and Trust (2002)Kidney – 1993–1998 14.6 (3.19–68.4) Stout and Trust (2002)


Fisk et al., 2003). This survival mechanism allowing the birdsto cope with the naturally high Hg levels encountered inmarine environments (Thompson and Furness, 1989).

4.2.4. Marine mammalsMarine mammals represent species in the highest trophiclevel of the food chain. Thus, themeasurements of Hg levels inthe tissues of marine mammals could provide some insightsinto the potential effects of trace metal due to consumption ofHg containing substances.

4.2.4.1. Walrus (Odobenus rosmarus). Walruses are bottomfeeders (LentferandGalster, 1987). InNunavik (Canada), reportedHg concentrations in tissues of walruswere in the range of 0.04–2.64 µg/g (Table 7). The highest level was found in livers and thelowest level was reported inmuscle (Muir et al., 2000). Similarly,Dietz et al. (1996) reported much higher Hg concentrations inlivers of walruses than in muscle. Moreover, they found that Hglevels in livers of adult walrus were 5–6 times higher than innewborn ones. On the other hand, Hg levels in muscle of adultwalrus were only 40% higher than the newborn walrus,indicating stronger accumulation of Hg in livers than in muscle.

4.2.4.2. Belugas (Delphinaperus leucas). In Canada, Belugasinhabit the eastern and western Arctic (Sergeant and Brodie,

1975;Wagemann et al., 1996). Thesemarinemammals are oneof the major diets for Inuit people. In 1990, Federal Govern-ment of Canada launched a multi-year scientific researchactivity named ‘‘The Northern Contaminants Program of theArctic Environmental Strategy’’ as part of the ‘‘Arctic GreenPlan’’. Partial of the results related to this program werereported by Wagemann et al. (1996). These researchersreported Hg levels in tissues of belugas sampled between1981 and 1994 across Arctic and in the St. Lawrence Riverregion (Table 7). The sequence of Hg concentrations in tissuesof belugas is: liverNNkidneyNmuscleNmuktuk (traditionalmeal of whale skin and blubber). Though levels of Hgconcentrations in muktuk were one or two orders lower thanin liver, still approximately 50–60% of the animals had Hglevels in muktuk in excess of 0.5 (µg/g, wet wt., the guidelinefor commercial sale of fish, Table 1). The skin of somecetaceans consists of four distinct layers: the outermostlayer (degenerative epidermis), underlain by the stratumexternum, which in turn is underlain by the stratum inter-medium and finally the innermost layer, the dermis, which isunderlain by blubber (St. Aubin et al., 1990; Wagemamn andKozlowska, 2005). Wagemamn and Kozlowska (2005) reportedHg concentrations in different skin layers of belugas (Table 7).They found that Hg levels increased progressively outwardwith a mean concentration of 1.5 µg/g (wet wt.) in the outmost

Table 7 –Mercury concentrations in marine mammals


Hg Level(µg/g, dw)

Reference

Walrus(Odobenusrosmarus)

Nunavik,Canada

Inukjuaq Liver – 1999 2.64 (1.46–3.81) Muir et al. (2000)Kidney – 1999 0.31 Muir et al. (2000)Muscle – 1999 0.04 (0.03–0.04) Muir et al. (2000)

Lancaster Sound Muscle – 1988–1990 0.41±0.08 (w.w) Atwell et al. (1998)Greenland Northwest Greenland Liver Newborn 1975–1977 0.31 (w.w) Dietz et al. (1996)

Kidney Newborn 1975–1977 0.094 (w.w) Dietz et al. (1996)Muscle Newborn 1975–1977 0.057 (w.w) Dietz et al. (1996)Liver 11 1975–1977 1.78 (w.w) Dietz et al. (1996)Muscle 11 1975–1977 0.08 (w.w) Dietz et al. (1996)

Beluga whales(Delphinapterusleucas)

NWT Meckenzie delta Liver 12.2±4.46 1981 10.7±10.7 Lockhart et al. (2001)Meckenzie delta Liver 21.1±9.36 1984 17.8±16.5 Lockhart et al. (2001)Meckenzie delta Liver 22.2±6.27 1993 34.5±27.4 Lockhart et al. (2001)Meckenzie delta Liver 17.7±6.21 1994 28.4±29.3 Lockhart et al. (2001)Meckenzie delta Liver 15.7±5.43 1995 44.0±35.1 Lockhart et al. (2001)Meckenzie delta Liver 14.6±5.68 1996 43.8±31.3 Lockhart et al. (2001)Meckenzie delta Liver 15.5±5.24 2001 38.9±41.7 Lockhart et al. (2001)Lancaster Sound Muscle – 1988–1990 2.25±0.97 Atwell et al. (1998)

Nunavut Arviat Liver 11.8±6.29 1984 7.25±6.94 Lockhart et al. (2001)Arviat Liver 11.1±4.84 1999 12.5±10.0 Lockhart et al. (2001)Coral Harbour Liver 16.1±3.22 1993 6.54±2.97 Lockhart et al. (2001)Coral Harbour Liver 13.1±7.31 1997 13.8±29.3 Lockhart et al. (2001)Coral Harbour Liver 8.92±5.90 2000 4.10±2.40 Lockhart et al. (2001)Grise Fiord Liver 5.65±4.83 1984 2.0±1.72 Lockhart et al. (2001)


Nunavut Iqaluit Liver 12.9±3.48 1993 7.57±4.86 Lockhart et al. (2001)Iquluit Liver 12.9±6.34 1994 16.3±8.93 Lockhart et al. (2001)Lake Harbour,Kimmirut

Liver 10.8±6.78 1994 8.8±6.16 Lockhart et al. (2001)

Nastapoka River liver 13.2±7.67 1984 11.5±13.9 Lockhart et al. (2001)Pangnirtung Liver 11.1±3.81 1984 5.05±4.43 Lockhart et al. (2001)Pangnirtung Liver 8.23±3.77 1993 8.45±7.00 Lockhart et al. (2001)Pangnirtung Liver 8.41±6.32 1994 10.7±13.4 Lockhart et al. (2001)Pangnirtung Liver 13.0±4.19 1997 8.73±4.67 Lockhart et al. (2001)Paulatuk Liver 14.3±5.13 1993 8.58±10.0 Lockhart et al. (2001)Repulse Bay Liver 8.00±8.49 1993 3.42±3.09 Lockhart et al. (2001)Sanikiluaq Liver 13.7±5.52 1994 12.9±9.53 Lockhart et al. (2001)Sanikiluaq Liver 13.0±5.37 1998 21.1±25.3 Lockhart et al. (2001)Kangiqsujuaq &Puvirnituq

Liver 6–28 1999 11.6 Lockhart et al. (2001)

WesternArctic

69oN–69o38N124oW–137o27'W

Bubbler – 1993 0.12±0.10 (wet. wt) Wagemann andKozlowska (2005)

Canada Western Arctic Muscle 13.9±5.5 1981–1984 1.07±1.47 (w.w.) Wagemann et al. (1990)Liver 13.9±5.5 1981–1984 11.8±12.1 (w.w) Wagemann et al. (1990)Kidney 13.9±5.5 1981–1984 2.83±1.71 (w.w.) Wagemann et al. (1990)Muktuk 19.3±6.6 1993–1994 0.78±0.41 (w.w.) Wagemann et al. (1996)Muscle 19.3±6.6 1993–1994 1.34±0.67 (w.w.) Wagemann et al. (1996)Liver 19.3±6.6 1993–1994 27.1±24.7 (w.w.) Wagemann et al. (1996)Kidney 19.3±6.6 1993–1994 4.91±2.84 (w.w.) Wagemann et al. (1996)

Eastern Arctic Muktuk 11.9±6.0 1984–1994 0.59±0.22 (w.w.) Wagemann et al. (1996)Muscle 11.9±6.0 1984–1994 0.94±0.44 (w.w.) Wagemann et al. (1996)Liver 11.9±6.0 1984–1994 8.40±8.25 (w.w.) Wagemann et al. (1996)Kidney 11.9±6.0 1984–1994 3.10±1.71 (w.w.) Wagemann et al. (1996)


WesternArctic

Eastern Arctic Muscle 10.2±6.6 1984 0.82±0.46 (w.w.) Wagemann et al. (1996)

CanadaLiver 10.2±6.6 1984 6.1±8.37 (w.w.) Wagemann et al. (1996)Kidney 10.2±6.6 1984 2.38±1.59 (w.w.) Wagemann et al. (1996)Muktuk 13.5±4.87 1993–1994 0.59±0.22 (w.w.) Wagemann et al. (1996)Muscle 13.5±4.87 1993–1994 1.04±0.43 (w.w.) Wagemann et al. (1996)Liver 13.5±4.87 1993–1994 10.2±8.00 (w.w.) Wagemann et al. (1996)Kidney 13.5±4.87 1984–1994 3.73±1.95 (w.w.) Wagemann et al. (1996)

St. Lawrence Muscle 18.1±9.1 1982–1987 2.46±1.46 (w.w.) Wagemann et al. (1990)Liver 18.1±9.1 1982–1987 33.6±43.0 (w.w.) Wagemann et al. (1990)Kidney 18.1±9.1 1982–1987 6.37±9.16 (w.w.) Wagemann et al. (1990)Muscle – 1993 1.4 (wet. wt) Dietz et al. (1996)


Table 7 (continued)


Hg Level(µg/g, dw)

Reference


Greenland Northwest Greenland Liver 0–6 1985 0.799 (w.w.) Dietz et al. (1996)Kidney 0–6 1985 0.442 (w.w.) Dietz et al. (1996)Muscle 0–6 1985 0.343 (w.w.) Dietz et al. (1996)Liver 7–13 1985 2.38 (w.w.) Dietz et al. (1996)Kidney 7–13 1985 1.44 (w.w.) Dietz et al. (1996)Muscle 7–13 1985 0.281 (w.w) Dietz et al. (1996)


WesternArctic

69oN–69o38'N124oW–137o27’W

Skin: stratumexternum 1

0.2±0.11 1993 1.5±0.93 Wagemann andKozlowska (2005)


0.58±0.53 1993 1.3±0.73

Layers 1 and 2 0.78±0.44 1993 1.40±0.84Skin: Stratumintermed

7.4±1.4 1993 1.0±0.54

Dermis 4.4±1.3 1993 0.29±0.15Layers 3 and 4 11.8±1.35 1993 0.65±0.40Skin as a whole(Muktuk)

12.5±2.7 1993 0.84±0.43

Narwhals(Monodonmonoceros)

EasternArctic

Lancaster Sound Muscle – 1988–1990 2.32 Atwell et al. (1998)63o20'N, 86o2'W66o44'N, 66o3'W

Bubbler – 1993 0.03±0.03 (wetwt.) Wagemann andKozlowska (2005)Muscle – 1993 0.81±0.19 (w. w)

Lancaster Sound Muscle – 1988–1990 2.32 Atwell et al. (1998)Narwhals(Monodonmonoceros)

EasternArctic

63o20'N, 86o2'W66o44'N, 66o3'W


0.15±0.05 1993 1.4±0.50 Wagemann andKozlowska (2005)


0.38±0.18 1993 0.82±0.23

Layers 1 and 2 0.53±0.15 1993 1.11±0.39Skin: Stratumintermed

8.5±0.98 1993 0.68±0.18

Dermis 4.6±1.9 1993 0.16±0.06Layers 3 and 4 13.1±1.5 1993 0.42±0.13Skin as a whole(Muktuk)

13.6±2.2 1993 0.59±0.13

Ringed seals(Phoca hispida)

Canada Arctic Bay Liver 8.1 1998–2000 7.83±3.55 Muir et al. (1999a,b)Nunavut Kidney 8.1 1998–2000 1.56±0.26 Muir et al. (1999a,b)

Liver 0–5 2000 7.42±7.11 (w.w) Riget et al. (2005)Liver N=6 2000 12.4±14.4 (w.w) Riget et al. (2005)Kidney 0–5 2000 1.38±0.89 (w.w) Riget et al. (2005)Kidney N=6 2000 1.78±0.95 (w.w) Riget et al. (2005)

Resolute Liver 5.8 1998–2000 3.08±2.50 Muir et al. (1999a,b)Kidney 6.0 1998–2000 1.39±0.27 Muir et al. (1999a,b)

Pond inlet Liver 3.3 1998–2000 4.00±2.58 Muir et al. (1999a,b)Kidney 3.1 1998–2000 1.23±0.26 Muir et al. (1999a,b)Liver 0–5 2000 6.84±16.9 (w.w) Riget et al. (2005)Liver N=6 2000 14.9±9.8 (w.w) Riget et al. (2005)Kidney 0–5 2000 1.04±0.45 (w.w) Riget et al. (2005)Kidney N=6 2000 2.54±0.73 (w.w) Riget et al. (2005)

Pangnirtung Liver 5.3 1998–2000 6.04±4.34 Muir et al. (1999a,b)Ringed seals(Phoca hispida)

Arviat Liver 18.7 1998–2000 14.0±8.79 Muir et al. (1999a,b)Grise Fiord Liver 18.7 1998–2000 18.4±11.0 Muir et al. (1999a,b)

Kidney 18.7 1998–2000 2.75±0.59 Muir et al. (1999a,b)Liver 0–5 2000 19.6±10.0 (w.w) Riget et al. (2005)Liver N=6 2000 29.9±26.5 (w.w) Riget et al. (2005)Kidney 0–5 2000 2.31±0.87 (w.w) Riget et al. (2005)Kidney N=6 2000 3.19±1.48 (w.w) Riget et al. (2005)

Labrador Nain Makkovik Liver 5.9 1998–2000 6.31±3.71 Muir et al. (2001b)Kidney 5.9 1998–2000 1.04±0.13 Muir et al. (2001b)Muscle – 1998–2000 0.328±0.085 Muir et al. (2001b)

Nunavik,QC

Ungava Bay Liver 5.1 1998–2000 10.8±6.29 Muir et al. (2001b)Kidney 4.9 1998–2000 1.08±0.21 Muir et al. (2001b)Muscle – 1998–2000 0.251±0.078 Muir et al. (2001b)

Hudson Stait Liver 5.2 1998–2000 3.5±3.24 Muir et al. (2001b)Kidney 5.2 1998–2000 0.69±0.14 Muir et al. (2001b)Muscle – 1998–2000 0.182±0.061 Muir et al. (2001b)

(continued on next page)(continued on next page)


Table 7 (continued)


Hg Level(µg/g, dw)

Reference

NWT Lancaster Sound Muscle – 1988–1990 1.07±0.11 Atwell et al. (1998)Holman Liver 0–5 2001 19.6±10.0 (w.w) Riget et al. (2005)Holman Liver N=6 2001 29.9±26.5 (w.w) Riget et al. (2005)Sachs Harbour Liver 0–5 2001 19.2±23.6 (w.w) Riget et al. (2005)Sachs Harbour Liver N=6 2001 19.1±9.6 (w.w) Riget et al. (2005)Sachs Harbour Kidney 0–5 2001 2.23±1.24 (w.w) Riget et al. (2005)Sachs Harbour Kidney N=6 2001 2.04±0.81 (w.w) Riget et al. (2005)


U.S.A. Barrow Liver 0–5 1995–1997 0.734±0.653 (w.w) Riget et al. (2005)Alaska Liver N=6 1995–1997 4.78±5.72 (w.w) Riget et al. (2005)

Kidney 0–5 1995–1997 0.379±0.285 (w.w) Riget et al. (2005)Kidney N=6 1995–1997 0.55±0.27 (w.w) Riget et al. (2005)

Greenland Avanersuaq Liver 0–5 1998 4.59±4.20 (w.w) Riget et al. (2005)Liver N=6 1998 9.15±6.58 (w.w) Riget et al. (2005)

Qeqertarsuaq Liver 0–5 1999–2000 1.23±2.66 (w.w) Riget et al. (2005)Liver N=6 1999–2000 2.91±2.03 (w.w) Riget et al. (2005)

Ittoqqortoormitt Liver 0–5 1999–2000 4.78±4.35 (w.w) Riget et al. (2005)Liver N=6 1999–2000 7.96±5.08 (w.w) Riget et al. (2005)

NorthwestGreenland

Liver 1 1984–85 0.714 (w.w) Dietz et al. (1996)Kidney 1 1984–85 0.585 (w.w) Dietz et al. (1996)Muscle 1 1984–85 0.164 (w.w) Dietz et al. (1996)Liver 2–4 1984–85 1.58 (w.w) Dietz et al. (1996)Kidney 2–4 1984–85 0.83 (w.w) Dietz et al. (1996)Muscle 2–4 1984–85 0.225 (w.w) Dietz et al. (1996)Liver 5–10 1984–85 2.96 (w.w) Dietz et al. (1996)Kidney 5–10 1984–85 1.03 (w.w) Dietz et al. (1996)Muscle 5–10 1984–85 0.229 (w.w) Dietz et al. (1996)

Central WestGreenland

Liver 1 1978–87 0.514 (w.w) Dietz et al. (1996)Kidney 1 1978–87 0.367 (w.w) Dietz et al. (1996)Muscle 1 1978–87 0.068 (w.w) Dietz et al. (1996)

SouthwestGreenland

Liver 1 1986 0.948 (w.w) Dietz et al. (1996)Kidney 1 1986 0.465 (w.w) Dietz et al. (1996)Muscle 1 1986 0.084 (w.w) Dietz et al. (1996)


Liver 1 1985–91 1.40 (w.w) Dietz et al. (1996)Kidney 1 1985–91 0.834 (w.w) Dietz et al. (1996)Muscle 1 1985–91 0.162 (w.w) Dietz et al. (1996)


Liver 2–4 1985–91 2.79 (w.w) Dietz et al. (1996)Kidney 2–4 1985–91 1.05 (w.w) Dietz et al. (1996)Muscle 2–4 1985–91 0.251 (w.w) Dietz et al. (1996)Liver 5–10 1985–91 7.57 (w.w) Dietz et al. (1996)Kidney 5–10 1985–91 1.86 (w.w) Dietz et al. (1996)muscle 5–10 1985–91 0.333 (w.w) Dietz et al. (1996)Liver N15 1985–91 19.9 (w.w) Dietz et al. (1996)Kidney N15 1985–91 3.29 (w.w) Dietz et al. (1996)Muscle N15 1985–91 0.514 (w.w) Dietz et al. (1996)

Norway Svalbard Liver 0–5 1996 0.64±0.32 (w.w) Riget et al. (2005)Liver N=6 1996 1.19±0.74 (w.w) Riget et al. (2005)Kidney 0–5 1996 2.87±2.89 (w.w) Riget et al. (2005)Kidney N=6 1996 4.62±3.99 (w.w) Riget et al. (2005)

Russia White sea Liver 0–5 2001 1.87±1.44 (w.w) Riget et al. (2005)Liver N=6 2001 2.41±2.78 (w.w) Riget et al. (2005)Kidney 0–5 2001 0.43±0.106 (w.w) Riget et al. (2005)Kidney N=6 2001 0.79±0.40 (w.w) Riget et al. (2005)

Hooded seal(Cystophoracristata)

Greenland NorthwestGreenland

Liver 1 1986 7.85 (w.w.) Dietz et al. (1996)Kidney 1 1986 1.40 (w.w.) Dietz et al. (1996)Muscle 1 1986 0.244 (w.w.) Dietz et al. (1996)

Polar bear(Ursusmaritimus)

Greenland Disko Bay region Muscle – 1995–1996 0.086±0.204 (w.w) Johansen et al. (2000)Liver – 1995–1996 10.3±7.62 (w.w) Johansen et al. (2000)

Greenland NorthwestGreenland

Liver 2–6 1988–90 12.6 (w.w) Dietz et al. (1996)N=7 1988–90 21.6 (w.w) Dietz et al. (1996)

Kidney 2–6 1988–90 10.8 (w.w) Dietz et al. (1996)N=7 1988–90 20.9 (w.w) Dietz et al. (1996)

Muscle 2–6 1988–90 0.056 (w.w) Dietz et al. (1996)N=7 1988–90 0.057 (w.w) Dietz et al. (1996)



Table 7 (continued)


Hg Level(µg/g, dw)

Reference

Polar bear(Ursus maritimus)





Avanersuaq,North




Avanersuaq,South

Muscle 1 1988–90 0.082 (w.w) Dietz et al. (2000)2–6 1988–90 0.034 (w.w) Dietz et al. (2000)N=7 1988–90 0.034 (w.w) Dietz et al. (2000)

Liver 1 1988–90 4.25 (w.w) Dietz et al. (2000)2–6 1988–90 12.9 (w.w) Dietz et al. (2000)N=7 1988–90 22.0 (w.w) Dietz et al. (2000)

Kidney 1 1988–90 4.22 (w.w) Dietz et al. (2000)2–6 1988–90 12.6 (w.w) Dietz et al. (2000)N=7 1988–90 28.1 (w.w) Dietz et al. (2000)

Ittoqqortoormiit,North




Polar bear(Ursusmaritimus)

Ittoqqortoormiit,South




Alaska Western area Muscle 2–5 1972 0.04±0.004 (w.w) Lentfer and Galster (1987)N5 1972 0.04±0.013 (w.w) Lentfer and Galster (1987)

Liver 2–5 1972 3.92±0.319 (w.w.) Lentfer and Galster (1987)N5 1972 4.8±0.487 (w.w) Lentfer and Galster (1987)

Northern area Muscle 2–5 1972 0.15±0.004 (w.w) Lentfer and Galster (1987)N5 1972 0.19±0.030 (w.w) Lentfer and Galster (1987)

Liver 2–5 1972 22.35±4.69 (w.w.) Lentfer and Galster (1987)N5 1972 38.08±5.194 (w.w) Lentfer and Galster (1987)

NWT Lancaster Sound Muscle – 1988–1990 0.84±0.17 Atwell et al. (1998)


layer. By peeling off this layer, consumers of muktuk canreduce somewhat the dietary intake of Hg. During the processof moulting, this and the underlying layer are shed andapproximately 20% of the total Hg in the skin is lost annually(Wagemann et al., 1996).

Mean Hg concentrations in each tissue of belugas in thewestern Arctic were much higher than in Eastern Arctic (by afactor of 3 in liver). The large difference in Hg levels of belugasbetween the eastern and western Arctic could be linked togeologically affected food diet. Environmental background ofHg levels in western Arctic are higher than those in easternpart (Wagemann et al., 1995). The concentration range of Hg innear-shore surficial coastal marine bottom sediments in the

western Arctic is 68–243 ng/g dw (Thomas et al., 1982), and 40–60 ng/g (dw), in the eastern Arctic (Loring, 1984). In the water, arange of 11–29 ng/L has been reported for the Beaufort Sea(Weiss et al., 1974) and 3.7 ng/L for the Arctic Ocean (Schmidtand Freimann, 1984). In rocks, Hg levels are approximately onemagnitude higher in the western Arctic than in the easternArctic (40–400 ng/g vs. 4–40 ng/g) (Wedephol et al., 1978;Cannon et al., 1978; Wagemann et al., 1995). Therefore, theenvironmental background concentration of Hg is clearlyhigher in the western than the eastern Arctic and this appearsto reflect in the tissues of belugas (Wagemann et al., 1996) andthe Hg discharge by the Makenzie River described by Leitchet al., 2007. On the other hand, the highest levels of Hg in all


tissues of belugas from the St. Lawrence River could be due tohighly polluted environment in the St. Lawrence River regionthan in the Arctic area.

4.2.4.3. Ringed seals (Phoca hispida). Ringed seals (P. hispida)are among the top predators of the marine food chain andhave a broad circumpolar distribution (Fisk et al., 2003; Rigetet al., 2005). Therefore, ringed seals are suitable monitoringspecies for both spatial and temporal trends (Riget et al., 2005).The diet of ringed seals consists of fish (mainly schoolinggadids, Arctic cod) and crustaceans (amphipods, mysids, andeuphausids). Ringed seals are a key component of the diet ofInuit (Fisk et al., 2003), thus they are one of themost frequentlystudied species in Arctic area. For example, Riget et al. (2005)reported an extensive study regarding spatial distribution ofHg concentrations in tissues of ringed seals in Polar region.Hg concentrations in liver and kidney of ringed seal weremeasured from 11 locations across the Arctic, from Alaska,Canada, Greenland, Svalbard to the White Sea. Results of thisstudy are summarized in Table 7. Mean Hg levels in liverand kidney of ringed seals varied with in a range of 0.64±0.32–29.9±26.5 and 0.38±0.29–3.19±1.48 µg/g (wet wt.), respectively.Total Hg levels in liver of ringed seals were approximately onemagnitude higher than in kidney with the largest differencesin Canadian Arctic. High levels of Hg in livers of ringed seals(27.5 µg/g wet wt.) were also reported by Smith and Armstrong(1978) from West Victoria Island in the Canadian Arctic.Interestingly, no evidence of Hg intoxication in these sealswas observed. It appears likely that marine mammals ingeneral have developed mechanisms enabling them to copewith the substantial amount of Hg that occurs naturally intheir diets (Smith and Armstrong, 1978).

A clear geographic pattern of Hg levels in liver of adult wasreported by Riget et al. (2005) (Table 7). Total Hg concentrationsincrease westward from the White Sea and Salbard toIttoqqortoormiit (East Greenland) and Avanersuaq (WestGreenland), across the Canadian Arctic with the highest levelsin Holman and Sachs Harbour. The lowest Hg levels in liverand kidney of ringed seals were found in Svalbard (Norway)and the highest value was reported in Western CanadianArctic. The most likely explanation for the geographicalpattern of Hg in ringed seals could be due to differentmineralogy of the Arctic and diet (Muir et al., 1999a; Rigetet al., 2005). On the other hand, age influence on Hg concen-trations in liver and kidney was documented by Riget et al.(2005). They found significantly higher levels in adultringed seals than in sub adults, which agreed well with thestudy by Fisk et al. (2003). The strong correlations between Hglevels and age indicated accumulation of Hgwith age in ringedseals.

4.2.4.4. Polar bears (Ursus maritimus). Polar bear is the prin-cipal mammalian predator at the top of the Arctic food chain.The diet of the polar bear consists mainly of the ringed seal andin some areas the bearded seal. Ringed seals eat mainlycrustaceans andArctic cod, thereby forming a simple and directfood chain between polar bears and the marine environment(Braune et al., 1991). The exposure to high concentrations of anumber of bio-accumulating contaminants makes polar bearimportant to study (Dietz et al., 2000).

Hg concentrations in the tissues of polar bears have beenextensively examined in Arctic regions, such as Greenland,Alaska and Canada (Lentfer and Galster, 1987; Braune et al.,1991; Dietz et al., 1996, 2000) (Table 7). Dietz et al. (2000) inves-tigated Hg levels from four areas in Greenland between 1983and 1990. Levels ofHgwerewithin the rangeof 0.034–0.191 µg/g(wet wt.) in muscle, 2.13–22 µg/g (wet wt.) in liver and 2.87–32.0 µg/g (wetwt.) in kidney, respectively. Hg concentrations inliver andkidneywerequite similar,while these levelswere oneor two magnitudes larger than in muscle. The highest meanconcentration was found in kidney of adults from Ittoqqor-toormiit (32.0 µg/g) and the largest liver means were found inadult bears from Southern Avanersuaq (22 µg/g).

Hg levels measured in the tissues of Greenland bears werewithin the same range of the values from Alaskan bears(muscle: 0.04–0.19 µg/g wet wt., liver: 3.92–33.08 µg/g wet wt.)(Lentfer and Galster, 1987). Levels of Hg in muscle tissue werenot different for young and adults animals, which wereconsistent with the observations by Dietz et al. (2000). On theother hand, both Lentfer and Galster (1987) and Dietz et al.(2000) found accumulation of Hg in liver with age. Interest-ingly, Lentfer and Galster (1987) found that Hg levels collectedin Alaska from tissues in northern area were much higherthan those from western area. These authors suggested thatdifferences in origin and composition of the water mass,which in turn affect diets of Hg levels in bears, might becontribute to different Hg levels in these two areas.

A comprehensive comparison regarding Hg distributionover a broader geographical scope based on published data(Braune et al., 1991; Norheim et al., 1992; Dietz et al., 2000) wasreported by Dietz et al. (2000). These researchers found that Hgconcentrations tended to increase from Svalbard over eastGreenland and northwest Greenland, peaking in polar bearsfrom southwest Melville Island. Hg levels in Melville wereapproximately 10 times higher than in east Greenland, and upto 30 times higher than in bears from Svalbard. A rapiddecrease of Hg from west of Melville Island with the lowestvalues appeared in the Chukchi Sea. No data was availablefrom the western and central part of the Russian Arctic forfurther comparison.

5. Hg concentrations in terrestrial biota

5.1. Hg concentrations in plants

Vegetation is sustaining the terrestrial food web of the tundra;hence it is a valuable indicator of contaminants entering thetundra ecosystem (Braune et al., 1999).

5.1.1. Hg levels in rootless plants

5.1.1.1. Lichen (Cetratia nivalis). Someperennial plants, suchas mosses and lichens, are the main diet for terrestrial animalsin tundra. For example, in northern Québec, lichens representup to 77% of the caribou diet during the fall and winter seasons(Gauthier et al., 1989). Lichens lack root system and grow upmainly on absorbing nutrients in the atmosphere. Meanwhile,toxic substances, such as Hg, could be carried along withnutrients, resulting in an accumulation of contaminants in


these plants (Crête et al., 1992; Thomas et al., 1992; Macdonaldetal., 1996; Larter andNagy, 2000;Gambergetal., 2005a). PrimaryHg uptake by plants and secondary Hg intake via food byanimals could lead to Hg accumulation in biota. Therefore, thecontaminant burden in lichens and other plant foods areimportant in affecting toxic levels in biota (Robillard et al., 2002).

Crête et al. (1992) reportedHg levels in lichens collected overa 640000 km2 fromnorthern Québec. Themeasured Hg concen-trations in lichens varied from 0.01 to 0.27 (µg/g, dry wt.) with amean value of 0.09±0.01 µg/g (dw) (Table 8). A significantinteraction between altitude and biome for Hg concentrationswas observed. Hg levels in the tundra below 400mwere almostdouble as those in the forest tundra and the boreal forest. Aspatial variation with progressive increase of Hg eastward andalongwith distance from seawaterswas documented. Hg levelsin Northern Québec were higher than in Greenland and inAlaska. Riget et al. (2000b) reported a median value of 0.034–0.061 µg/g (dw) in lichensat four locationsofGreenland (Table8).Meanwhile, no influence of substrate on Hg concentrations inlichen was reported, indicating other sources rather than soildust have strong influence on Hg levels in lichen.

5.1.1.2. Moss (Rhacomitrium lanuginosum). Moss is a com-mon species used to track environmental concentration ofatmospheric contaminants. Besides, ground squirrels arebelieved to incorporate small amount of mosses in their diet(Batzli and Sobaski, 1980). Ford et al. (1995) reported an averageHg level of 0.044 µg/g (dw) in Hylocomium splendens (a mosscommonly used inmonitoring studies) in Alaska (Table 8). Thislevel was approximately half value of the Hg concentrations inGreenland (0.099–0.106 µg/gdw, Riget et al., 2000b). Interestingly,Hg concentrations in mosses were about one magnitude largerthan in soil, suggesting that other sources influenced theelevated Hg levels in moss tissues. Meanwhile, from theconcurrent measurements of Hg concentrations in moss andlichen, these authors discovered that lichen ismore suitable formonitoring atmospheric deposition of Hg than the moss. Thedifferent monitoring abilities might be attributed to dissimilarphysical appearance and micro habitat.

Table 8 – Hg concentrations in vegetation

Species Region Location T

Lichen(C. nivalis)

Canada Northern QuébecGreenland Avanersuaq

NuukQaqortoqTasiilaq

Cosmopolitan lichen(Cetraria cucullata)

Alaska Schrader/Peters Lake

Moss(Rhacomitrium lanuginosum)

Greenland AvanersuaqNuukQaqortoqTasiilaq

Moss(Hylocomium splendens)


Blueberry(Vaccinium uliginosum)


Fruit(Dryas octapetala)


5.1.2. Vascular plantsHgconcentrations intissuesof vascularplants reflect rootuptakeand within-plant partitioning via vascular transport as well(Adriano, 1986). In a study performed by Ford et al. (1995), Hglevels in the fruits of two vascular plants, namely blueberry(Vaccinium uliginosum) and Dryas octapetala collected from Alaskawere reported (Table 8). The blueberry is a forage food and fruitsformanyherbivores and omnivores of arctic tundra.D. octapetalafruits, on the other hand, were the kind of diet that the groundsquirrelswere feedingduring their studyperiod. Bothblueberriesand D. octapetala fruits have very low concentrations of Hg withHg levels in blueberries near detection value (0.003 µg/g, dw).However, these reported Hg levels in two vascular plants werebased on small sample size, their specific toxicological applica-tions; quantitative estimates of potential exposure to contami-nants from food web sources must be carefully evaluated basedon the known seasonal roundof ingested foods (Ford et al., 1995).

5.2. Hg concentrations in terrestrial animals

5.2.1. Moose (Alces alces)Consumption of organs from some wildlife species, such asmoose (A. alces) and caribou (Rangifer tarandus) as food iscommon for indigenous populations of the regions in Canada(Gamberg, 2000). Since 1994, the Yukon Contaminants Com-mittee has conducted a hunter survey for themeasurement ofheavy metals in the tissues of moose (Gamberg et al., 2005b).Measured mean Hg level in the kidney of moose was 0.02±0.02 µg/g (wet wt.) (Table 11). Hg concentrations in moosekidneys were uniformly low, thus did not approach concen-trations that would be considered of toxicological concern.

5.2.2. Caribou (R. tarandus)Caribou (R. tarandus) is one of the dominant terrestrial animalsin theCanadianArctic and is considered as amajor food sourceby local people. Caribou are strict herbivores with a winter dietconsisting primarily of lichens. The air–plant–animal contam-ination pathway represented by the lichen–caribou food webmake caribou a potentially useful species for monitoring

issue Sampleperiod

Hg Level(µg/g, dw)

Reference

1985 0.09±0.01 (0.01–0.27) Crete et al. (1992)0.061 (0.049–0.089) Riget et al. (2000b)0.034 (0.033–0.056) Riget et al. (2000b)0.038 (0.029–0.0430 Riget et al. (2000b)0.038 (0.032–0.044) Riget et al. (2000b)

1990–1992 0.054 (0.042–0.072) Ford et al. (1995)

0.106 (0.094–0.132) Riget et al. (2000b)0.091 (0.088–0.137) Riget et al. (2000b)0.104 (0.059–0.172) Riget et al. (2000b)0.101 (0.074–0.196) Riget et al. (2000b)0.044 (0.035–0.054) Ford et al. (1995)

b0.01 Ford et al. (1995)

0.016 (0.01–0.018) Ford et al. (1995)


changes inArctic terrestrial ecosystem (Elkin andBethke, 1995;Larter and Nagy, 2000).

Total Hg levels in tissues of barren-ground caribou havebeen reported in several studies (Table 11). For example,Robillard et al. (2002) conducted a survey regardingHg levels insamples of muscle, kidney and liver from caribou harvested intwo regions of northern Québec between 1994 and 1996. Hglevels in tissues of caribou varied in a range of 0.02–1.39 µg/g(wet wt.). The distribution of Hg levels in tissues of caribouwas: kidneyN liverNNmuscle. Liver and kidney represent theprincipal target organs for heavy metal accumulation inanimals (Medvedev, 1999). Therefore, both liver and kidneyare the main site of Hg accumulation in caribou. Similarly, aconsistently higher Hg concentration in kidney than in liverwas reported in other studies (Elkin and Bethke, 1995; Fisk etal., 2003).

The influence of spatial variation on Hg levels in cariboutissues was documented by Robillard et al. (2002). Theseresearchers found that levels of Hg in muscle, kidneys andliver in Leaf River (LR) regionwere higher than in George River–Torngat Mountains (GRTM) area (Table 11): the former beinglocated on this west coast of Nunavik whereas the latter beinglocated on the East cost of Nunavik. Additionally LR regionwere more contaminated than GRTM area with the lichenbiomass of the north-western part nearly twice than in theeastern area. Therefore, the LR region probably provides alarger supply of more contaminated lichens to browsingcaribou, resulting in an elevated Hg levels in caribou tissuesfrom LR than in GRTM area. The results from Robillard et al.(2002) were consistentwith other studies that foraging ecologycould be an important factor influencing Hg levels in caribou.For example, Larter and Nagy (2000) found that Hg levels in thekidney tissue of Banks Island Peary caribou (10.45±0.85 µg/g(dw), n=20) were significantly higher than in Bluenose caribou(5.43±0.31 µg/g (dw), n=20) in Northwest Territories of Canada.The high Hg levels in Bluenose than in Peary caribou could beattributed to considerable higher dietary lichen component inBluenose than in Banks Island Peary caribou (85% vs. 1%)(Larter and Nagy, 1997, 2000). Caribou that feed substantiallyon lichen may accumulate significant levels of contaminants,especially in their kidneys and livers, over a lifetime (Larterand Nagy, 2000).

On the other hand, little geology effect on Hg concentrationwas reported by Larter and Nagy (2000) when comparing Hglevels in the kidney tissues of caribouwith other caribou herdsfrom five caribou herds located in Nunavut and NorthwestTerritories (Elkin and Bethke, 1995) (Table 11). Larter and Nagy(2000) conclude that dietary lichen rather than geographicvariation has substantial effects on Hg accumulation incaribou kidneys. Temporal variation of Hg concentrations incaribou tissues was reported by Robillard et al. (2002), whofound that Hg concentrations in kidneys and muscledecreased from December to May in the LR region, but itincreased or remained stable in the GRTM region. Theysuggest that foraging ecology and physiology of caribou inthe two regions could provide some explanations for Hgfluctuations. Sex related differences in metal concentrationshave been reported in some studies. Robillard et al. (2002)found that the effects of sex proved limited and marginal: Hgconcentration decreased with time of year in males, but not in

females. However, no age effect on Hg levels was reported intheir study.

5.2.3. Reindeer (R. tarandus fennica)Hg levels in hair samples of reindeer in Karelian region werereported by Medvedev (1999). Measured Hg concentrations inhair varied from0.08 to 0.44 µg/g (dw)with amean level of 0.15±0.02 µg/g (n=19). No sex- and age-dependent differences werefound.

5.2.4. Wolf (Canis lupus)Arcticwolves are top trophic level terrestrial carnivores, feedingon snowshoe hare (Lepus americanus),moose, and other animals(Gamberg and Braune, 1999). In Arctic areas of the NorthwestTerritories (Canada), caribou has been shown to be thepredominant food item of wolves (Braune et al., 1999). Thus,an experimental wolf reduction program in the Yukon wasdesigned to enhance a declining sympatric caribou population(Heyes, 1992). This program provided a unique opportunity toobtain tissue samples for contaminant analysis (Gamberg andBraune, 1999). Hg levels in kidney and liver of wolves aged fromb18months to N36months were within the range of 0.18±0.06–1.06±0.61 µg/g (wet wt.) and 0.03±0.02–1.45±2.24 µg/g (wet wt.),respectively (Table 11). Average Hg concentrations reported forwolf tissues are similar to those found in NWT wolves. Hgconcentrations in the tissues of wolves collected in Yukon andNWT regions are generally low and do not approach levels thatareknown topotentially cause adverse effects in animals (30 µg/g, wet wt., Thompson, 1996). Therefore, Hg concentrationsfound in this study should be considered baseline levels. Anincrease in renal Hg with age was observed by Gamberg andBraune (1999). On the other hand, no correlation between Hg inliver and age was reported (Table 9).

6. Trophic transfer of mercury

Asdiscussedabove, once released to theenvironment,Hgmightbe transferred and remained in plants and animals via marine,freshwater and terrestrial food chains (Fig. 3) (AMAP, 1997). Thefood webs in arctic lake ecosystem usually include four well-defined levels, with phytoplankton at the bottom, herbivoresthat include zooplanktons, detrital feeders such as benthicinsect larvae and crustacean, and carnivores feeding on benthicorganisms (Braune et al., 1999). Fish inArctic freshwater, suchaschar and lake whitefish (Coregonus clupeaformis) are the mostcommon first-order carnivores. Arctic predatory fish includeinconnu (Coregonus leucichthys), lake trout, and burbot (L. lota)(Braune et al., 1999). Arctic terrestrial food webs, on the otherhand, are generally simple, consisting of plants or lichens at theprimary producer level, a few herbivores, and one or two mainpredators (Braune et al., 1999). The most abundant livingbiomass in Arctic is found in the marine ecosystem. The highproductivity of biota inmarine ecosystem is a result of seasonalcycle of growing andmelting sea ice,which allowsnutrient-richwater to reach the sunlit surface of the ocean (AMAP, 1997).

Transfer of Hg through food ingestion is the dominantpathway for uptake of Hg in food webs (Fig. 3). Studies havedemonstrated that concentrations of Hg increase with trophicpositioning in the food web, with top predators having the

Table 9 – Hg concentrations in terrestrial animals

Species Region Location or herd Tissue Age Sampleperiod

Hg Level(µg/g, dw)

Reference

Snowshoe hare Northwest Territories Inuvik Liver – 1992–93 0.08±0.01 (w.w) Poole et al. (1998)Lepus americanus Fort Good Hope Liver – 1992–93 0.04±0.01 Poole et al. (1998)Red-backed vole Northwest Territories Inuvik Liver – 1992–93 0.08 Poole et al. (1998)Clethrionomysrutilus

Fort Good Hope Liver – 1992–93 0.25 Poole et al. (1998)Fort Liard Liver – 1992–93 0.11 Poole et al. (1998)Fort Smith Liver – 1992–93 0.25 Poole et al. (1998)

Caribou(Rangifer tarandus)

Nunavut Beverly herd Kidney – 2000 6.15 (5.64–8.16) Fisk et al. (2003)Liver – 0.80 (0.46–1.38) Fisk et al. (2003)

Bluenose herd Kidney – 1998 1.92 (1.03–3.59) Fisk et al. (2003)Liver – 0.46 (0.31–0.69) Fisk et al. (2003)

South Baffin population Kidney 5.4±2.8 1999 3.13 (2.28–4.30) Fisk et al. (2003)Liver 5.4±2.8 0.75 (0.54–1.06) Fisk et al. (2003)

Porcupine herd Kidney – 1996 2.12 (1.58–2.84) Fisk et al. (2003)Kidney 4.0±1.9 1997 2.03 (1.49–2.76) Fisk et al. (2003)Kidney 3.2±2.9 1998 1.39 (0.42–4.57) Fisk et al. (2003)

Tay herd Kidney 3.2±2.2 1998 0.87 Fisk et al. (2003)NWT Banks Island Kidney – 1995 5.43±0.31 Larter and Nagy (2000)

Bluenose Kidney – 1995 10.45±0.85 Larter and Nagy (2000)Bathurst Kidney – 1991, 92 0.52±0.04 (w.w.) Elkin and Bethke (1995)

Liver – 1991, 92 0.16±0.03 (w.w) Elkin and Bethke (1995)Arviat Kidney – 1991, 92 2.93±0.21 (w.w.) Elkin and Bethke (1995)

Liver – 1991, 92 0.92±0.08 (w.w) Elkin and Bethke (1995)Southampton Is. Kidney – 1991, 92 2.22±0.13 (w.w.) Elkin and Bethke (1995)Cape Dorset Kidney – 1991, 92 1.25±0.05 (w.w.) Elkin and Bethke (1995)

Liver – 1991, 92 0.38±0.11 (w.w) Elkin and Bethke (1995)Lake Harbour Kidney – 1991, 92 2.56±0.25 (w.w.) Elkin and Bethke (1995)

Liver – 1991, 92 0.58±0.08 (w.w.) Elkin and Bethke (1995)

Table 10 – Trophic level of marine food web (Source:Campbell et al., 2005)

Species Common name Scientificname

Trophiclevel

Primary Ice alae – 1.3Invertebrate Copepod Calanus

hyperboreus2.0

Invertebrate Mixed zooplankton – 2.2Invertebrate Amphipod Themisto libellula 2.5Invertebrate Mysid Mysis oculata 2.7Fish Arctic cod Boreogadus saida 3.7Sea bird Dovekie Alle alle 3.2Sea bird Black-legged

kittiwakeRissa tridactyla 3.8

Sea bird Black guillemot Ceppus grylle 3.9Sea bird Thick-billed murre Uria lomvia 3.9Sea bird Ivory gull Pagophilia

eburnea4.0

Sea bird Northern fulmar Fulmarus glacialis 4.1Sea bird Glaucous gull Larus

hyperboreus4.7

Sea bird Thayer's gull Larus thayeri 4.9Mammal Ringed seal Phoca hispida 4.6


highest Hg burdens (Gamberg et al., 2005a; Bodaly et al., 1993;Muir et al., 2005).

High levels of Hg have been discovered in Arctic fish andmarine mammals (AMAP, 1997). Therefore, Hg accumulationin the biota at the Arctic regionsmay present a serious risk forindigenous populations of the regions. In particular, Hg can bebio-accumulated through food chain as methyl Hg (MeHg),which is a neurotoxin that may cause irreversible damage tothe central nervous system of fetuses and may lead to severeneurological impairment or death in adults (National ResearchCouncil, 2000; Grandjean et al., 1997, 1999). Moreover, MeHgaccumulates in fish muscle tissue cannot be reduced byselective cooking methods, such as removal fat or skin.(Flaherty et al., 2003). MeHg has been found in some types ofbiota at concentrations that may be harmful to human health.Due to the health concerns about people who tend to beexposed to high Hg levels, some advisory information/recommendations regarding Hg levels were issued (Table 10).For example, fish consumption advisories use a conservativereference dose for pregnant women, women of child-bearingage, and children (National Research Council, 2000; Grandjeanet al., 1997, 1999). The current reference dose is 0.1 µg/kg/day,which is the maximum level of MeHg exposure recordedwithout deleterious effects on fetuses (National ResearchCouncil, 2000; Marsh et al., 1987). In Canada, Hg level in wildnon-marine mammal tissues (liver or kidney) considered tocause toxic effects is 30 µg/g (wet wt.) (Thompson, 1996). Theguideline for human consumption and sale of fish is restrictedif levels of Hg in muscle of fish exceed 0.5 µg/g. The recom-mended maximum is even lower (0.2 µg/g) for people who

engage in subsistence fisheries and consumer larger quanti-ties of fish (Health and Welfare Canada, 1979, 1984). Addition-ally, the U.S. Environmental Protection Agency (USEPA, 1995)proposed guideline of Hg levels in fish tissue b57 ng/g (wet wt.)for protection of fish-eating wildlife. The National Academy ofScience recommend keeping the whole Hg level in humanblood b5.0 µg/L or in hair b1.0 µg/g corresponding to a refer-


ence dose (RfD) of 0.1 µg/kg body weight per day (NationalAcademy of Science (NAS), 2000; Hightower and Moore, 2003).On the other hand, the half-life time of ingested or absorbedHg varies with different Hg species and biota (Table 11). Inhuman body, the average half-life of MeHg in blood of adults is70 days, in children 90 days, and in lactating women 46 days(Swartout and Rice, 2000).

It is known that Hg can be absorbed and accumulated in thebody of biota through three pathways: bioconcentration,bioaccumulation, and bio-magnification (Fisk et al., 2003).Bioconcentration refers to the process by which plants andanimals take up contaminants directly from air, water, or soil.The bioavailability of contaminant to aquatic and terrestrialanimals will depend on the properties of the chemical and onthe physical, chemical and biological environment into whichit is released (Fisk et al., 2003). The food requirements of anorganism are controlled by metabolic rate and production(growth, lipid deposition, reproduction). Thusmetabolic rate islinked to the rate of uptake of contaminant (Fisk et al., 2003).Bioaccumulation includes bioconcentration as well as theuptake of contaminants from the food an animal eats. Thebody burden due to bioaccumulation will thus depend oncontaminant concentrations in air, water, soil, animal's diet,as well as on an organism's ability to rid itself of contaminants(AMAP, 1997). Bio-magnification occurs when a contaminantis not broken down or excreted, but accumulates as it passesup the food web. Bio-magnification is the major reason whypersistent environmental contaminants reach high concen-trations in top predators even when levels in air, soil, andwater are low (AMAP, 1997). Therefore, animals on higher

Table 11 – Threshold levels of biological effects for mercury lev

Group Tissue Concentration Gu

Fish Muscle 0.2 µg/g (w.w) Frequent Consumers ofMuscle 0.5 µg/g (w.w.) Commercial Sale of Fish

Birds Liver 30 mg/kg (w.w) Lethal level in free rangKidney 30 mg/kg (w.w) Lethal level in free rangEgg 30 mg/kg (w.w) Detrimental effect upon

Marinemammals

Liver 60 mg/kg (w.w) Liver damage

Terrestrialmammals

Liver 25 mg/kg (w.w) Laboratory succumbed aKidney 25 mg/kg (w.w) Laboratory succumbed a

Human Bloodhair

5.0 µg/L1.0 µg/g

Tolerable levelTolerable level

Body 0.1 µg/kg Reference dose (RfD)bodyweight/day

Body 1.6 µg/kg Tolerable weekly intakebodyweight/week

(TWI) levels

Birds andmammals

Muscle 0.05 µg /g (w.w) Danish food standard li

Birds andmammals

Liver and kidney 0.1 µg/g (w.w) Danish food standard li

Fish andcrustaceans

Muscle 0.3 µg/g (w.w) Danish food standard li

Fish Muscle 0.7–1.5 µg/g (w.w) Consumption and expoline in European Contri

Fish Muscle 0.057 ng/g (w.w) USEPA guideline for proPeregrinefalon

Eggs 1 µg/g (w.w) Critical value for populareproductive effects

Food intakes 300 µg Provisional tolerable we

trophic levels tend to accumulate higher Hg levels than thoseof lower trophic levels (Wagemann, 1989; Chan et al., 1995).

Utilization of stable isotopes of nitrogen (δ15 N/δ14 N) can beuseful for tracing biocontaminants in food webs (Cabana andRasmussen, 1994; Atwell et al., 1998). Concentrations of δ15 N areprogressively enriched from prey to predator on the order of 3–5‰ (Peterson and Fry, 1987). A continuous enrichment of δ15 Nfrom filter-feeding clam to polar bear at the topof the arctic foodchain was documented by Atwell et al. (1998) during a food webstudy. Campbell et al. (2005) determined the structure of foodweb based on nitrogen and carbon stable isotope (Table 12).Similar as the pattern reported by Atwell et al. (1998), anincreasing δ15 N values was observed from primary, to inver-tebrate, to fish, to seabirdsand ringedseals. The trophicpositionof sea birds varied from 3.2 for dovekie (a species mainly feedson zooplankton and fish) to a tropic level of 4.7 for glaucous gull(a species whose main diet is fish, carrion and seabird chicks).Significant trends of increasing total Hg concentrations withincreasing trophic level through the food web were reported inboth studies (Atwell et al., 1998; Campbell et al., 2005). On theother hand, the accumulation of Hg through trophic food chainmay vary among different group of species. For example, nocorrelation between Hg concentrations and trophic level withinthe invertebrates as a single group was found by two researchgroups (Atwell et al., 1998; Campbell et al., 2005). Atwell et al.(1998) speculate that there may be different transfer mechan-isms occurring at different levels of the food web. On the otherhand, no evidence of bioaccumulation of Hg with age in themuscle tissue of clams (Mya truncate) was observed, despite theage range covering 42 years.

els birds, mammal and humans

ideline/Effect Reference

Fish Guideline Health Canada (1979 )Guideline Shilts and Coker (1995);

Stephens (1995)ing birds Thompson (1996)ing birds Thompson (1996)free ranging bird hatching Thompson (1996)

Law et al. (1996)

nimals due tomercury intoxication Thompson (1996)nimals due tomercury intoxication Thompson (1996)

Mahaffey and Rice (1998)Mahaffey and Rice (1998)NAS (2000)

WHO; Booth and Zeller (2005)

mits Dietz et al. (1996)



rt guild Clark (1989)estection of fish-eating birds USEPA (1995); Braune et al. (1999)tion level Peakall et al. (1990)

ekly intake by people AMAP (2002)

Table 12 – Half-life of mercury in the tissues of organisms

Mercury species Organism Half-life (time it takes for the tissueconcentration to be reduced by half)

Reference

Organic Hg Fish 323 days from diet AMAP (1997)Inorganic Hg 45–61 from water or diet AMAP (1997)MeHg Seals and dolphins 500–100 days AMAP (1997)MeHg Whole body of human 52–93 days AMAP (1997)Inorganic Hg Whole body of human 40 days AMAP (1997)MeHg Blood of adults 70 days Swartout and Rice (2000)MeHg Blood of children 90 days Swartout and Rice (2000)MeHg Lactating women 46 days Swartout and Rice (2000)


Solid evidences of bio-magnification of Hg with tropic levelwithin fish were found in several studies (Bruce and Spencer,1979; Kidd et al., 1995). For instance, Lockhart et al. (2001) re-ported that predatory fish like northern pike and walleye(Stizostedion vitreum) have generally higher Hg concentrationsthan species feeding on lower trophic levels, like whitefish andArctic char, which mainly feed on zooplankton and benthos.Significant correlation between Hg and δ15 N in landlocked charfrom was observed by Muir et al. (2001a,b). However, no otherheavy metals were correlated with δ15 N, implying that bio-magnifications of Hg is occurring within the char populationdue to the presence of piscivorous char.

Similarly, a survey regarding trophic transfer of Hg in water-fowl and game birds was conducted in the Canadian Arctic from1988 to 1995 (Coad, 1994; Braune et al., 1999) (Table 13). Thesebirds represent a total of 35 species that collected from 29 areas,13 of which were located in the western Arctic (north of 60°N,west of 95°W)and16ofwhichwere in theeasternArctic (Northof55°N, east of 95°W). The birds were grouped into 5 classesaccording to feeding habit: Browsers—ground dwellers such asgrouse and ptarmigan that feedmainly on terrestrial vegetation;Grazers—geese that graze mainly on aquatic and terrestrialvegetation;Omnivores—surface-feedingduckswith a varieddietconsisting mainly of aquatic vegetation; Molluscivores—divingducks feedingmainly on invertebrates; Piscivores—diving ducksfeedingmainlyon fish. Ingeneral, totalHg levels inbreastmuscleof these birds varied within a range of no detectable—1930 ng/g(wet wt.) as shown in Table 13. The highest Hg concentrationswere found in the top trophic level of birds (piscivores) and thelowest was found in the first and second trophic levels of birds(browsers and grazers). The pattern of Hg transfer with trophic

Table 13 – Hg transfer through waterfowl and game birds

Waterfowl andgame birds

Region Location

Browsers North Canada East BreNorth of 55oN, East of 95oW West Bre

Grazers East BreWest Bre

Omnivores East BreWest Bre

Molluscivores East BreWest Bre

Piscivores East BreWest Bre

level in birds was in a good agreement with Atwell et al. (1998).They found that lower trophic level seabirds (e.g. dovekie andeider)hadHgconcentrationsalmostanorderofmagnitude lower(0.33, 0.38 µg/g, respectively) than the two highest trophicposition species, thick-billed murre and glaucous gull (1.79 and1.86 µg/g, respectively). The bio-magnification of Hg in the toptrophic level of birds indicates transfer ofHg through vegetation–invertebrate–fish–bird food chain.

An overall bio-magnification power of 0.20 in arctic marinefoodweb (i.e. invertebrates, fish, birds, mammals) was reportedby Atwell et al. (1998), which was of similar magnitude to thatobserved in freshwater fishes (0.20–0.30) from several lakes byKidd et al. (1995). On the other hand, the total bio-magnificationreported by Atwell et al. (1998) does not reflect the potential ofHg to biomagnify in specific subsets of marine food web. Thetransfer of Hg in polar bears was a notable exception, having alower mean Hg concentration than their prey, ringed seals.Atwell et al. (1998) suggested that the low bio-magnification ofHg in polar bears could be attributed to their food dietary habits.Polar bears preferentially consume seal skin and fat (Norstromet al., 1986), which, in many marine mammals, are low in Hgcontent relative to other tissues (Sergeant and Armstrong, 1973;Muir et al., 1988). Polar bears are therefore exposed to lessdietary Hg than other animals that consume whole bodies, ormostly muscle, of their prey.

7. Human health

From time to time, various agencies have provided referenceexposure limits (REL) of GEM for risk assessment purposes;

Tissue Sampleperiod

Hg Level(µg/g, w.w.)

Reference

ast muscle 1988–1995 b100 Braune et al. (1999)ast muscle 1988–1995 b60 Braune et al. (1999)ast muscle 1988–1995 b130 Braune et al. (1999)ast muscle 1988–1995 b30–156 Braune et al. (1999)ast muscle 1988–1995 74–218 Braune et al. (1999)ast muscle 1988–1995 b30–665 Braune et al. (1999)ast muscle 1988–1995 b40–455 Braune et al. (1999)ast muscle 1988–1995 b20–369 Braune et al. (1999)ast muscle 1988–1995 74–1230 Braune et al. (1999)ast muscle 1988–1995 258–1930 Braune et al. (1999)


300 ng/m3 (US EPA, 1995), 200 ng/m3 (ASTDR, 1999), 1000 ng/m3

(European office of the World Health organization (WHO),2000), and 90 ng/m3 (California environmental protectionagencies (CalEPA), 2005). As a continuing effort to improvesuch criteria, the contaminated sites division of HealthCanada proposed a chronic REL of GEM as 80 ng/m3 (HealthCanada, 2007) (Table 14). According to various GEM concentra-tions measured in the Arctic (Steffen et al., 2005; Steffen et al.,2007) no risk appears in regard to REL in the Arctic.

However, chronic metal toxicity is a concern in theCanadian Arctic because of the findings of high metal levelsin wildlife animals and the fact that indigenous peoplesdependent onwild animals and plants as their traditional food(Chan et al., 1995). It is estimated that the average annual percapita consumption of traditional food among Inuit is morethan twice the estimated Canadian average annual consump-tion of meat and fish (Wong, 1985). The major source ofhuman exposure to trace metals from the environment isfrom food. The high bio-magnification of Hg through foodchains implies that small changes in concentrations in thediet could result in large changes in Hg in top predators (Muiret al., 2005), thus imposing high health risk to indigenouspeople (Goyer, 1991). Based on a comprehensive dietary surveycollected in Baffin Inuit community, Chan et al. (1995) reportedthat Hg levels in traditional food ranged from 0 to 797 µg/100 gwith a mean value of 38 µg/100 g, which were higher thanthose in corresponding Canadianmarket food. Concentrationsof Hg in organ meats of traditional food were higher than theaction level of 50 µg/100 g or 0.5 ppm set by AgricultureCanada. For all individual food studied, ringed seal (P. hispida)meat was the most frequently consumed item. It constitutedabout one-third by weight of all traditional food eaten by bothadults (N20 years old) and children (3–12 years old). Ringed sealmeat and liver and narwhal mattack (Monodon monoceros)together contributed about 75, 71, and 70% of Hg to women,men and children, respectively. The average daily intakelevels of total Hgin traditional food for both adults (65 µg forwomen and 97 µg for men) and children (38 µg) were higherthan the Canadian average value (16 µg). Based on daily intakeof Hg level, the calculated average weekly intake for all agegroups (6.6, 8.0 and 6.3 µg/kg body weight/week for adultwomen, men and children, respectively) exceed the intakeguideline (5.0 µg/kg bodyweight/week) established by the JointFood and Agriculture Organization/World Health Organization

Table 14 – Reference exposure limits (REL) of GEM for riskassessment purposes

Year Organisation Reference exposurelimits (REL) (ng/m3)

1995 US EPA 3001999 ASTDR _Agency for toxic substance

and disease registry—U.S.Department health and humanservice

200

2000 European office of theWorld Healthorganization

1000

2005 California environmentalprotection agencies

90

2007 Health Canada 80

Expert Committee on Food Additives and Contaminants(WHO, 1989). On the other hand, by assuming 80% of thetotal Hg is in the form of MeHg (Muir et al., 1992), Chan et al.(1995) determined that the average daily intake of MeHg were0.75, 0.91 and 1.5 µg/g body weight, respectively. These levelswere higher than the no-adverse-effect level of 0.48 µg/g bodyweight. Therefore, potential health effects on Inuit peoplemaybe of concern due to consumption of high levels of con-taminants in their traditional food.

A project conducted by an AMAP's Human Health groupestimated that the mean weekly intakes of Hg by Inuit peoplefrom Canada and Greenland (850 µg) were far beyond theCanadian dietary guidelines for exposure to Hg (300 µg), whichcould be attributed to their high intake meat or tissues of bothmarine and terrestrial animals (AMAP, 2002). Since 1994,AMAP initiated a circumpolar study of contaminant levels inblood. In general, levels of Hg are higher in people who relyheavily on marine food, such as the Inuit of Greenland andArctic Canada, as well as the Yup'ik in western Alaska. ForInuit, the Hg comes mainly from the muscle of marinemammals. In western Alaska, the levels can probably beexplained by high intake of northern pike. In the Faroe Islandsthe major source of Hg is pilot whale.

8. Conclusions

It is estimated that 90 to 450 metric tons of Hg are depositedannually in the Arctic due to AMDEs. However, it is debatewhether springtime AMDEs increase the Hg concentration insurface snow and how long this input may be preserved.Although snow may act as a sink during AMDEs, Hg can bereemitted back to the atmosphere and then act as a source tothe atmosphere. It was shown that Hg is highly mobile insnow, particularly in exposed snow to solar radiation.

Many recent findings suggest that the atmospheric con-tribution of long-range anthropogenic Hg to High Arctic lakesmay have been overestimated by several-fold because ofclimate-driven process, and were responsible for no morethan 22% of the 20th century Hg concentrations increase in thestudy lakes; or highlighted the importance of the formationand postdeposition crytollographic history of the snow and icecrystals in determining the fate and concentration of mercuryin the cryosphere in addition to AMDEs.

Nevertheless, oneof themost important issues is emerging:assessing the Hgmethylation process in the Arctic and tundra:Is Hg methylation effective in the Arctic tundra? Where thesources of MeHg are? What is its fate? Is this stimulated byhumanmade? And; how international policy can support andhelp to reduce Hg impact on the whole arctic ecosystem.

More than two hundred papers published over the lastdecade showed the interest of the scientific communities tothe Hg issue in the Arctic, especially the tundra. Despite, Hgfate in the Arctic is still under investigation and of concern.

Acknowledgements

The authors would like to thank the ArcticNet (Centre ofexcellence of Canada) theme 2.6 for funding.


R E F E R E N C E S

AdrianoDC.Traceelements in the terrestrial environment.NewYork:Springer; 1986, p. 533.

Agency for Toxic Substances and Desease Registry. Toxicologicalprofile for mercury. Atlanta, GA: US Department of Health andHuman Services, Public Health Service; 1999. Available athttp://www.atsdr.cdc.gov/tfacts46.html.

AMAP. Arctic pollution issues: a state of the Arctic environmentreport. Oslo, Norway: Arctic monitoring and AssessmentProgramme (AMAP); 1997. p. 188.

AMAP. Arctic pollution 2002 (Persistent organic pollutants, heavymetals, radioactivity, human health, changing pathways).Oslo, Norway: Arctic Monitoring and Assessment Programme(AMAP); 2002. xii+112pp.

Amato P, Hennebelle R, Magand O, Sancelme M, Delort AM,Barbante C, et al. Bacterial characterization of the snow coverat Spitzberg, Svalbard. FEMS Microbiol Ecol 2007;59:255–64.

Amundsen PA, Staldvik FJ, Lukin AA, Kashulin NA, Popova OA,Reshetnikov YS. Heavymetal contamination in freshwater fishfrom the border region between Norway and Russia. Sci TotalEnviron 1997;201:211–24.

AndersenR, Skurdal J, LinloekkenA.Mercury accumulation inArcticchar (Salvelinus alpinus) and brown trout (Salmo trutta) in anacidified coastal area, S.W. Norway. Vatten/Wate 1986;42:222–6.

Anthony RG, Garrett MG, Schuler CA. Environmental contaminantsin bald eagles in the Columbia River Estuary. J Wildlife Manage1993;57:10–9.

Ariya PA, Dastoor AP, Amyot M, Schroeder WH, Barrie L, Anlauf K,et al. The Arctic: a sink for mercury. Tellus 2004;56B:397–403.

Asmund, G., Nielsen, S.P., Mercury in dated Greenland marinesediments. Sci Total Environ 2000, 245, 61–72.

Aspmo K, Gauchard PA, Steffen A, Temme C, Berg T, Bahlmann E,et al. Measurements of atmospheric mercury species during aninternational study of mercury depletion events at Ny-Ålesund, Svalbard, spring 2003. How reproducible are ourpresent methods? Atmos Environ 2005;39:7607–19.

Aspmo K, Temme C, Berg T, Ferrari C, Gauchard PA, Fain X, et al.Mercury in the atmosphere, snow and melt water ponds in theNorth Atlantic Ocean during Arctic summer. Environ SciTechnol 2006;40(3):4083–9.

Atwell L, Hobson KA, Welch HE. Biomagnification andbioaccumulation of mercury in an arctic marine food web:insights from stable nitrogen isotope analysis. Can J Fish AquatSci 1998;55:1114–21.

Aulerich RJ, Ringer RK, Iwamoto S. Effects of dietary mercury onmink. Arch Environ Contam Toxicol 1974;2:43–51.

Bargagli R, Agnorelli C, Borghini F, Monaci F. Enhanced depositionand bioaccumulation of mercury in Antarctic terrestrialecosystems facing a coastal polynya. Environ Sci Technol2005;39(21):8150–5.

Barkay T, Poulain AJ. Mercury (micro)biogeochemistry in polarenvironments. FEMS Microbiol Rev 2007;59:232–41.

Barrie L,MacdonaldR, BidlemanT,DiamondM,GregorD, SemkinW,et al. Sources, occurrence and pathways, Chapter 2. Ecosystemuptake and effects. In: Jensen J, Adare K, Shearer R, editors.Canadian Arctic contaminants assessment report. Ottawa:Indian and Northern Affairs Canada; 1997. p. 25–182.

Batzli GO, Sobaski ST. Distribution, abundance, and foragingpatterns of ground squirrels near Atkasook, Alaska. Arct AlpRes 1980;12:501–10.

Berg T, Bartnicki J, Munthe J, Lattila H, Hrehoruk J, Mazur A.Atmospheric mercury species in the European Arctic:measurements andmodelling. Atmos Environ 2001;35:2569–82.

Berg T, Sekkesæter S, Steinnes E, Valdal AK,Wibetoe G. Springtimedepletion of mercury in the European Arctic as observed atSvalbard. Sci Total Environ 2003;304(1–3):43–51.

Bindler R, Renberg I, Appleby PG, Anderson NJ, Rose NL. Mercuryaccumulation rates and spatial patterns in lake sedimentsfrom West Greenland: a coast to ice margin transect. EnvironSci Technol 2001;35:1736–41.

Björklund I, Borg H, Johansson K. Mercury in Swedish lakes—regional distribution and causes. Ambio 1984;13:118–21.

Bodaly RA, Rudd JWM, Fudge RJP, Kelly CA.Mercury concentrationsin fish related to size of remoteCanadian shield lakes, 1993. CanJ Fish Aquat Sci 1993;50:980–7.

Booth S, Zeller D. Mercury, food webs, and marine mammals:implications of diet and climate change for human health.Environ Health Perspect 2005;113:521–6.

Borg H. Distribution of trace elements in Arctic lakes—somepreliminary results from the Tundra North West Expedition.McLean, VA, USA, 8–22 June 2001; 2001.

Boudreau BP. Metals and models: diagenetic modelling infreshwater lacustrine sediments. J Paleolimnol 1999;22:227–51.

Boyle J. Redox remobilization and the heavy metal record in lakesediments: amodelling approach. J Paleolimnol 2001;26:423–31.

Braune B. Retrospective survey of organochlorines andmercury inArctic seabird eggs. In: Ka lhok S, editor. Synopsis of researchconducted under the 1999–2000 Northern ContaminantsProgram. Ottawa: Indian and Northern AffairsCanada; 2000. p. 155–9.

Braune B. Retrospective survey of organochlorines and mercuryin Arctic seabird eggs. In: Ka lhok S, editor. Synopsis ofresearch conducted under the 2000–2001NorthernContaminants Program. Ottawa: Indian and NorthernAffairs Canada; 2001. p. 160–5.

Braune BM, Norstrom RJ, Wong MP, Collins BT, Lee J. Geographicaldistribution ofmetals in livers of polar bears from the northwestterritories, Canada. Sci Total Environ 1991;100:283–99.

Braune B, Muir D, DeMarch B, Gamberg M, Poole K, Currie R, et al.Spatial and temporal trends of contaminants in CanadianArctic freshwater and terrestrial ecosystems: a review. SciTotal Environ 1999:145–207.

BrauneBM, Outridge PM, FiskAT,Muir DCG, HelmPA, Hobbs K, et al.Persistent organic pollutants andmercury inmarine biota of theCanadian Arctic: an overview of spatial and temporal trends. SciTotal Environ 2005;351–352:4–56.

Brooks SB, Saiz-Lopez A, Skov H, Lindberg SE, Plane JMC,Goodsite ME. The mass balance of mercury in the springtimearctic environment. Geophys Res Lett 2006;33:L13812._________________________doi:10.1029/2005GL025525.

Bruce WJ, Spencer KD. Mercury levels in Labrador fish, 1977–78.Can Ind Rep Fish Aquat Sci 1979;111 12 pp.

Bustnes JO, ErikstadKE. The diet of sympatricwintering populationsof common eider, Somateria mollissima and king eider,S. spectabilis in northern Norway. Ornis Fenn 1988;65:163–8.

Bustnes JO, Lønne OJ. Habitat partitioning among sympatricwintering common eiders Somateria mollissima and king eiders,Somateria spectabilis. Ibis 1997;139:549–54.

Cabana G, Rasmussen JB. Modelling food chain structure andcontaminant bioaccumulation using stable nitrogen isotopes.Nature 1994;372:255–7.

Cabana G, Tremblay A, Kalff A, Rasmussen JB. Pelagic food chainstructure in Ontario lakes: a determinant of mercury levels inlake trout (Salvelinus namaycush). Can J Fish Aquat Sci1994;51:381–9.

California Environmental Protection Agency (CalEPA). Mercury,inorganic-chronic reference level and chronic toxicitysummary. California: Office of Environmental Health hazardAssessment; 2005.

Campbell LM,NorstromR,HobsonKA,MuirDCG, Backus S, FiskAT.Mercury and other trace elements in a pelagic Arctic marinefood web (Northwater Polynya, Baffin Bay). Sci Total Environ2005;351–352:247–63.

Cannon HL, Connally GG, Epstein JB, Parker JG, Thornton I,Wixson G. Rocks: geological sources of most trace elements.

http://www.atsdr.cdc.gov/tfacts46.html

http://dx.doi.org/10.1029/2005GL025525


Report to the workshop at south seas plantation, CaptivaIsland, FL, US. Geochem Environ 1978;3:17–31.

Clark RB. Marine pollution. 2nd edn. Oxford, England: ClarendonPress; 1989.

Coad, S. Consumption of fish and wildlife by Canadian nativepeoples: A quantitative assessment from the published andunpublished literature. Unpublished Report, Health Canada,1994.

Chapman PM, Wang F, Janssen C, Persoone G, Allen HE.Ecotoxicology of metals in aquatic sediments: binding andrelease, bioavailability, risk assessment, and remedation. Can JFish Aquat Sci 1998:2221–43.

Chan HM, Kim C, Khoday K, Receveur O, Kuhnlein HV. Assessmentof dietary exposure to trace metals in Baffin Inuit food. EnvironHealth Perspect 1995;103:740–6.

Compeau GC, Bartha R. Sulfate-reducing bacteria: principalmethylators of mercury in anoxic estuarine sediments. ApplEnviron Microbiol 1985;50(2):498–502.

Constant P, Poissant L, Villemur R, Yumvihoze E, Lean D. Fate ofinorganic mercury and methyl mercury within the snow coverin the low arctic tundra on the shore of Hudson Bay (Québec,Canada). J Geophys Res 2007;vol. 112:D08309.

_____________________________doi:10.1029/2006JD0079612007, 2007.Crête M, Lefebvere L, Walsh P. Cadmium, lead, mercury, and

137cesium in fruticose lichens of northern Quebec. Sci TotalEnviron 1992;121:217–30.

Dietz R, Riget F, Johansen P. Lead, cadmium, mercury andselenium in Greenland marine animals. Sci Total Environ1996;186:67–93.

Dietz R, Riget F, Born EW. Geographical differences of zinc,cadmium, mercury and selenium in polar bears (Ursusmaritimus) from Greenland. Sci Total Environ 2000;245:25–47.

Dietz R, Riget FF, Boertmann D, Sonne C, Olsen MT, Fjeldså J, et al.Time trends of mercury in feathers of West Greenland birds ofprey during 1851–2003. Environ Sci Technol 2006a;40:5911–6.

Dietz R, Riget FF, Born EW, Sonne C, Grandjean P, KirkegaardM, et al.Trends in mercury in hair of Greenlandic polar bears (Ursusmaritimus) during 1892–2001. Environ Sci Technol2006b;40(4):1120–5.

Doidge DW, Kaminski G, Mesher C0. Bluemussels (Mytilus edulis) inNunavik: distribution, demography and commercial potential.Report to fisheries and aquaculture testing andexperimentation program of Dept. of Fisheries and Oceans.Kuujjuaq, PQ: Makivik Research Centre; 1993.

Dommergue A, Ferrari CP, Gauchard PA, Boutron CF, Poissant L,Pilote M, et al. The fate of mercury species in a sub-arcticsnowpack during the snowmelt. Geophys Res Lett 2003a;30(12):1621. _________________________doi:10.1029/2003GL017308.

Dommergue A, Ferrari CP, Poissant L, Gauchard PA, Boutron CF.Diurnal cycles of gaseous mercury within the snowpack atKuujjuarapik/Whapmagoostui, Québec, Canada. Environ SciTechnol 2003b;37:3289–97.

Douglas MSV, Smol JP, Blake Jr W. Marked post-18th centuryenvironmental change in high-Arctic ecosystems. Science1994;266:416–9.

Douglas TA, Sturm M, Simpson WR, Brooks S, Lindberg SE,Perovich DK. Elevated mercury measured in snow and frostflowers near Arctic sea ice leads. Geophys Res Lett 2005;32:L04502. _________________________doi:10.1029/2004GL022132.

Douglas TA, Sturm M, Simpson WR, Blum JD, Alvare-Aviles L,Keeler GJ, et al. Influence of snow and ice crytals formation andaccumulation on mercury deposition to the Arctic. Environ SciTechnol 2008;42:1542–52.

Downs SG, Macleod CL, Lester JN. Mercury in precipitations and itsrelation to bioaccumulation in fish: a literature review. WaterAir Soil Pollut 1998;108:149–87.

Dunning Jr JB. Body masses of birds of the world. In: Dunning Jr JB,editor. CRC handbook of avian body masses. Boca Raton, FL:CRC Press; 1993. p. 3–312.

Ebinghaus R, Jennings SG, Schroeder WH, Berg T, Donaghy T,Guentzel J, et al. International field intercomparisonmeasurements of atmospheric mercury species at Mace Head,Ireland. Atmos Environ 1999;33:3063–73.

Engstrom DR, Swain EB, Henning TA, Brigham ME, Brezonik PL.Environmental chemistry of lakes and reservoirs. In: Baker LA,editor. ACS advances in chemistry series N° 237. AmericanChemical Society; 1994. p. 33–66.

Elkin BT, Bethke RW. Environmental contaminants in caribou inthe Northwest Territories, Canada. Sci Total Environ 1995;160/161:307–21.

EvansM,LockhartL.An investigationof factorsaffectinghighmercuryconcentrations in predatory fish in the Mackenzie River Basin.In: Kalhok S, editor. Synopsis of research conducted under the2000–2001Northern Contaminants Program. Ottawa: Indian andNorthern Affairs Canada; 2001. p. 174–84.

Ferrari CP, Dommergue A, Boutron CF, Jitaru P, Adams FC. Profilesofmercury in the snow pack at Station Nord, Greenland shortlyafter polar sunrise. Geophys Res Lett 2004;31:L03401.

_________________________doi:10.1029/2003GL018961.Ferrari CP, Gauchard PA, Aspmo K, Dommergue A, Magand O,

BahlmannE, et al. Snow-to-air exchanges ofmercury in anArcticseasonal snow pack in Ny-Ålesund, Svalbard. Atmos Environ2005;39:7633–45.

Finley MT, Stendell RC. Survival and reproductive success of blackducks fed methylmercury. Environ Pollut 1978;16:51–64.

Fisk AT, Tittlemier SA, Pranschke JL, Norstrom RJ. Usinganthropogenic contaminants and stable isotopes toassess the feeding ecology of Greenland shark. Ecology2002;83:2162–72.

Contaminant levels and trends in the biological environment. In:Fisk AT, Hobbs KE, Muir DCG, editors. Canadian arcticcontaminants assessment report, II. Canada: Indian andNorthern Affairs; 2003. 202 pp.

Fitzgerald WF, Engstrom DR, Lamborg CH, Tseng CM, Balcom PH,Hammerschmidt CR. Modern and historic mercury fluxes innorthern Alaska: global sources and Arctic depletion. EnvironSci Technol 2005;39:557–68.

Flaherty CM, Sass GG, Stiles KE. Human mercury toxicity and iceangler fish consumption: are people eating enough to causehealth problems? Risk Anal 2003;23:497–504.

Fleming EJ, Mack EE, Green PG, Nelson DC. Mecury methylationfrom unexpected sources: molybdate-inhibited freshwatersediments and an iron-reducing bacterium. Appl EnvironMicrobiol 2006;72(1):457–64.

Foley RE, Jackling SJ, Sloan RJ, Brown MK. Organochlorine andmercury residues in wildmink and otter: comparisonwith fish.Environ Toxicol Chem 1988;7:363–74.

Ford J, Landers D, Kugler D, Lasorsa B, Allen0Gil S, Crecelius E, et al.Inorganic contaminants in Arctic Alaskan ecosystems: long-rangeatmospheric transport or local point sources? Sci Total Environ1995;160/161:323–35.

Fortin C, Beauchamp G, Dansereau M, Lariviere N, Bélanger D.Spatial variation in mercury concentrations in wild mink andriver otter carcasses from the James Bay Territory, Québec,Canada. Arch Environ Contam Toxicol 2001;40:121–7.

Frimer O. Diet of moulting king eiders Somateria spectabilis at DiskoIsland, West Greenland. Ornis Fenn 1997;74:187–94.

Galasso JL, Siegel FR, Kravitz JH. Heavy metals in eight 1965 coresfrom the Novaya Zemlya Trough, Kara Sea, Russian Arctic. MarPollut Bull 2000;40:839–52.

Gamberg, M., Contaminants in Yukon Country Foods,Unpublished report prepared for the Department of Indian andNorthern Affairs, Whitehorse, Yukon, 2000. p95.

Gamberg M, Braune BM. Contaminant residue levels in arcticwolves (Canis lupus) from the Yukon Territory, Canada. SciTotal Environ 1999;243/244:329–38.

Gamberg M, Braune B, Davey E, Elkin B, Hoekstra PF, Kennedy D,et al. Spatial and temporal trends of contaminants in

http://dx.doi.org/10.1029/2003GL017308

http://dx.doi.org/10.1029/2004GL022132

http://dx.doi.org/10.1029/2003GL018961


terrestrial biota from the Canadian Arctic. Sci Total Environ2005a:148–64.

Gamberg M, Palmer M, Roach P. Temporal and geographic trendsin trace element concentrations in moose from Yukon,Canada. Sci Total Environ 2005b:530–8.

Gauchard PA, AspmoK, Temme C, Steffen A, Ferrari C, Berg T, et al.Study of the origin of atmospheric mercury depletion eventsrecorded in Ny-Ålesund, Svalbard, spring 2003. Atmos Environ2005a;39:7620–32.

Gauchard PA, Ferrarim CP, Dommergue A, Poissant L, Pilote M,Guehenneux G, et al. Atmospheric particle evolution during anighttime atmospheric mercury depletion event in sub-Arcticat Kuujjuarapik/Whapmagoostui, Québec, Canada. Sci TotalEnviron 2005b;336(1–3):215–24.

Gauthier L, Nault R, Créte M. Variations saisonnières du régimealimentaire des caribous du troupeau de la rivière George,Québec, nordique. Nat Can 1989;116:101–12.

Gilbert FF, Nancekivell EG. Food habits of mink (Mustela vison) andotter (Lutra Canadensis) in north eastern Alberta. Can J Zool1982;60:1282–8.

Gobeil C, MacDonald RW. A state of the Arctic Environment. AMAPReport ISBN 82-7655-060-6 Artic monitoring and assessmentprogram; 1997. AMAP, P.O. Boks 8100 Dep. N-0032 Oslo,Norway, 107 pp.

Gobeil C, McDonald RW, Smith JN. Mercury profiles in sedimentsof the Arctic Ocean basis. Environ Sci Technol 1999;33:4194–8.

Goyer RA.Toxic effects ofmetals. In: AmdurMO,Doull J, KlaasenCD,editors. Casarett and Doull's toxicology. 4th ed. New York:Pergamon Press; 1991. p. 623–80.

Grandjean P, Weihe P, White RF, Debes F, Araki S, Yokoyama K, etal. Cognitive deficit in 7-year-old children with prenatalexposure to methylmercury. Neurotoxicol Teratol1997;19:417–28.

Grandjean P, Budtz-Jorgensen E, White RF, Jorgensen PJ, Weihe P,Debes F, et al. Methyl mercury exposure biomarkers asindicators of neurotoxicity in children aged 7 years. Am JEpidemiol 1999;150:301–5.

Greenfield BK, Hrabik TR, Harvey CJ, Carpenter SR. Predictingmercury levels in yellow perch: use of water chemistry, trophicecology, and spatial traits. Can J Fish Aquat Sci 2001;58:1419–29.

Groudieva T, Kambourova M, Yusef H, Royter M, Grote R, Trinks H,et al. Diversity and cold-active hydrolytic enzymes ofculturable bacteria associated with Arctic sea ice, Spitzbergen.Extremophiles 2004;8:475–88.

Gubala CP, Landers DH, Monetti M, Heit M,Wade T, Lasorsa B, et al.The rates of accumulation and chronologies ofatmospherically derived pollutants in Arctic Alaska, USA. SciTotal Environ 1995;160/161:347–61.

Hammerschimidt CR, Fitzgerald WF, Lamborg CH, Balcom PH,Tseng CM. Biogeochemical cycle of methylmercury in lakesand tundra watersheds of Artic Alaska. Environ Sci Technol2006;40:1204–11.

Harris RC, Rudd JWM,AmyotM, Babiarz CL, Beaty KG, Blanchfield PJ.Whole-ecosystem study shows rapid fish–mercury response tochanges in mercury deposition. Proc Nat Acad Sci2007;104:16586–91.

Helth andWelfare Canada. Methylmercury in Canada. Exposure ofIndian and Inuit residents to methylmercury in the Canadianenvironment. Ottawa: Health and Welfare Canada, MedicalServices Branch; 1979. p. 200.

Helth andWelfare Canada. Methylmercury in Canada. Exposure ofIndian and Inuit residents to methylmercury in the Canadianenvironment. Ottawa: Health and Welfare Canada, MedicalServices Branch; 1984. p. 164.

Health Canada. Scientific criteria document for mercury vapour(HgO): Establishing refernce exposure level (REL) for riskassessment of HgO exposures in Canada. Contaminated sitesdivision, Bureau of risk and impact assessment, SelfEnvironments Programme June; 2007.

Heinz G. Methyl mercury: second-year feeding effects on mallardreproduction and duckling behaviour. J Wildl Manage1976;40:82–90.

Hermanson MH. Anthropogenic mercury deposition to Arctic lakesediments. Water Air Soil Pollut 1998;101:309–21.

Heyes RD. An experimental design to test wolf regulation ofungulates in the Aishihik area, southwest Yukon, TechnicalReport TR-92-6. White-horse, Yukon: Dept. of RenewableResources; 1992. p. 53.

Holl KD, Cairns Jr J. Landscape indicators in ecotoxicology. In:Hoffman DJ, Rattner BA, Burton Jr GA, Cairns Jr J, editors.Handbook of ecotoxicology. Boca Raton, Florida: LewisPublishers; 1995. p. 185–97.

Hoënninger, G. (2002), Halogen oxide studies in the boundary layerby¨multi axis differential optical absorption spectroscopy andactive longpath DOAS, Ph.D. thesis, Inst. fü r Umweltphys.,Univ. of Heidelberg, Heidelberg, Germany.

Hightower JM, Moore D. Mercury levels in high-end consumers offish. Environ Health Perspect 2003;111:604–8.

Hui A, Takekwa JY, Baranyuk VV, Litvin KV. Trace elementconcentrations in two subpopulations of lesser snow geesefrom Wrangel Island, Russia. Arch Environ Contam Toxicol1998;34:197–203.

Jensen J, Adare K, Shearer R, editors. Canadian Arcticcontaminants assessment report. Ottawa: Indian andNorthernAffairs Canada; 1997. 460 pp.

Johansen P, Pars T, Bjerregaard P. Lead, cadmium, mercury andselenium intake by Greenlanders from local marine food. SciTotal Environ 2000;245:187–94.

Junge K, Eicken H, Deming JD. Bacterial activity at − 2 to − 20 °C inArctic wintertime sea ice. Appl Environ Microbiol2004;70(1):550–7.

Junge K, Imhoff F, Staley T, Deming JW. Phylogenetic diversity ofnumerically important Arctic sea-ice bacteria cultured atsubzero temperature. Microb Ecol 2002;43:315–28.

Kerin EJ, Gilmour CC, Roden E, Suzuki MT, Coates JD, Mason RP.Mercury methylation by dissimilatory iron-reducing bacteria.Appl Environ Microbiol 2006;72(12):7919–21.

Kidd KA, Schindler DW, Muir DCG, Lockhart WL, Hesslein RH. Highconcentrations of toxaphene in fishes from a subarctic lake.Science 1995;269:240–2.

Kim KH, Ebinghaus R, Schroeder WH, Blanchard P, Kock HH,Steffen S, et al. Atmospheric mercury concentrations fromseveral observatory sites in the northern hemisphere. J AtmosChem 2005;50:1–24.

King JR, Farner DS. Energy metabolism, thermoregulation andbody temperature. In: Marshall AJ, editor. Biology andcomparative physiology of birds. New York: Academic Press;1961. p. 215–88.

Kirchmann DL. The ecology of Cytophaga-Flavobacteria in aquaticenvironments. FEMS Microbiol Ecol 2002;39:91–100.

Kirk JL, St. Louis VL, Hintelmann H. What is the source of methylmercury in Canadian Arctic oceans? ArcticNet AnnualScientific Meeting; 2005. Banff, Alberta.

Kirk JL, St. Louis VL, Sharp ML. Rapid reduction and reemission ofmercury deposited into the snowpacks during atmosphericmercury depletion events at Churchill, Manitoba, Canada.Environ Sci Technol 2006;40:7590–6.

Lahoutifard N, Poissant L, Scott SL. Scavenging of gaseousmercuryby acidic snow at Kuujjuarapik, Northern Québec. ScienceTotal Environ 2006;355(1–3):118–26.

Lahoutifard N, Sparling M, Lean D. Total and methyl mercurypatterns in Arctic snow during springtime at Resolute,Nunavut, Canada. Sci Total Environ 2005;39:7597–606.

Lalonde JD, Poulain AJ, Amyot M. The role of mercury redoxreactions in snow on snow-to-air mercury transfer. Environ SciTechnol 2002;36:174–8.

Landers DH, Gubala C, Verta M, Lucotte M, Johansson K, Vlasova T,et al. Using lake sediment mercury flux ratios to evaluate the


regional and continental dimensions of the mercurydeposition in Arctic and boreal ecosystems. Atmos Environ1998;32(5):919–28.

Larter NC, Nagy JA. Peary caribou, muskoxen and Banks Islandforage: assessing seasonal diet similarities. Rangifer1997;17:9–16.

Larter NC, Nagy JA. A comparison of heavymetals in the kidneys ofHighArctic andmainland cariboupopulations in theNorthwestTerritories of Canada. Sci Total Environ 2000;246:109–19.

Lee YH, Iverfeldt A. Measurement of methylmercury and mercuryin run-off, lake and rain waters. Water, Air, Soil Pollut1991;56:309–21.

Leitch DR, Carrie J, Lean D, Macdonald RW, Stern GA, Wang F. Thedelivery of mercury to the Beaufort Sea of the Arctic Ocean bythe Mackenzie River. Sci Total Environ 2007;373(1):178–95.

Lentfer JW, Galster A. Mercury in polar bears from Alaska. J WildlDis 1987;23:338–41.

Lin CJ, Pehkonen SO. The chemistry of atmospheric mercury: areview. Atmos Environ 1999;33:2067–79.

Lindberg SE, Brooks S, Lin CJ, Scott K, Meyers T, Chambers L, et al.Formation of reactive gaseous mercury in the Arctic: evidenceof oxidation of Hg0 to gas-phase Hg-II compounds after arcticsunrise. Water Air Soil Pollut: Focus 2001;1:295–302.

Lindberg S, Brook S, LinCJ, LandisMA, StevensRK, GoodsiteM, et al.Dynamic oxidation of gaseous mercury in the Arctic tropo-sphere at Polar Sunrise. Environ Sci Technol 2002;36:1245–56.

Lindeberg C, Bindler R, Renberg I, Emteryd O, Karlsson E,Anderson NJ. Natural fluctuations of mercury and lead inGreenland lake sediments. Environ Sci Technol 2006;40:90–5.

LockartWL,WilkinsonP,BilleckBN,DannelRA,HuntRV,BrunskillGJ,et al. Fluxes ofmercury to lake sediments in central and northernCanada inferred from dated sediment cores. Biogeochemistry1998;40:163–73.

Lockhart WL, Low G, Evans M. Mercury in fish from stock surveys inthe Western Northwest Territories: Investigations in the factorsaffectingmercury levels. In: Kalhok S, editor. Synopsis of researchconducted under the 1999–2000Northern Contaminants Program.Ottawa: Indian and Northern Affairs Canada; 2000. p. 181–90.

Lockhart WL, Stern G, Low G, Evans M. Mercury in fish from stocksurveys in theWestern Northwest Territories: investigations inthe factors affecting mercury levels. In: Kalhok S, editor.Synopsis of Research Conducted Under the 2000/01 NorthernContaminants Program. Ottawa: Indian and Northern AffairsCanada; 2001. p. 194–201.

Loring DH. Trace-metal geochemistry of sediments from BaffinBay. Can J Earth Sci 1984;21(12):1368–78.

Loseto L, Lean DRS, Siciliano SD. Snowmelt sources ofmethylmercury to high arctic ecosystems. Environ Sci Technol2004a;38:3004–10.

Loseto LL, Siciliano SD, Lean DRS. Methylmercury production inhigh Arctic wetlands. Environ Toxicol Chem 2004b;23:17–23.

Lucotte M, Mucci A, Hillare-Marcel C, Pichet P, Grodin A.Anthropogenicmercury enrichment in remote lakes of northernQuébec (Canada). Water Air Soil Pollut 1995;80:467–76.

Macdonald CR, Ewing LL, Elkin BT, Wiewel AM. Regional variationin radionuclide concentrations and radiation dose in caribou(Rangifer tarandus) in the Canadian Arctic; 1992–1994. Sci TotalEnviron 1996;182:53–73.

Macdonald RW, Barrie LA, Bidleman TF, Diamond ML, Gregor DJ,Semkin RG, et al. Contaminants in the Canadian Arctic: 5 yearsof progress in understanding sources, occurrence andpathways. Sci Total Environ 2000;254:93–234.

Macdonald RW, Harner T, Fyfe J. Recent climate change in theArctic and its impact on contaminant pathways andinterpretation of temporal trend data. Sci Total Environ2005;342:5–86.

Magnusson JJ, RobertsonDM,BensonBJ,WynneRH, LivingstoneDM,Arai T, et al. Historical trends in lake and river ice cover of theNorthern Hemisphere. Science 2000;289:1743–6.

Marsh DO, Clarkson TW, Cox C, Myers GJ, Amin-Zaki L, Al-Tikriti S.Fetal methyl mercury poisoning: relationship betweenconcentration in single strands of maternal hair and childeffects. Arch Neurol 1987;44:1017–22.

Mason RP, Fitzgerald WF, Morel FMM. The biogeochemical cyclingof elemental mercury: anthropogenic influence. GeochimCosmochim Acta 1994;58:3191–8.

Medvedev N. Levels of heavy metals in Karelian wildlife, 1989–91.Environ Monit Assess 1999;56:177–93.

Meuleman C, Leermakers M, Baeyens W. Mercury speciation inLake Baikal. Water Air Soil Pollut 1995;80:539–51.

Muir DCG, Wagemann R, Grift N, Norstrom RJ, Simon M, Lien J.Organochlorine chemical and heavy metal contaminantsin white-beaked dolphins (Lagenorhynchus albirostis) andpilot whales (Globicephala melaena) from the coast ofNewfoundland, Canada. Arch Envrion Contam Toxicol1988;17:613–29.

Muir DCG, Wagemann R, Hargrave BT, Thomas DJ, Peakall DB,NorstromRJ. Arcticmarine ecosystem contamination. Sci TotalEnviron 1992;122:75–134.

MuirD, BrauneB, deMarchR,NorstromR,WagemannM,GambergK,et al. Chapter 3. Ecosystem uptake and effects. In: Shearer R,editor. Canadian Arctic contaminants assessment report.Ottawa: Indian and Northern Affairs Canada; 1997. p. 191–294.

Muir D, KwanM, Lampe J. Spatial trends and pathways of POPs andmetals in fish, shellfish and marine mammals of northernLabrador and Nunavik. In: Kalhok S, editor. Synopsis ofresearch conducted under the 1998/99 Northern ContaminantsProgram. Ottawa: Indian and Northern Affairs Canada; 1999b.p. 165–71. Environmental Studies Report.

Muir D, Braune B, DeMarch B, NorstromR,WagemannR, Lockhart L,et al. Spatial and temporal trends and effects of contaminants inthe Canadian Arctic marine ecosystem: a review. Sci TotalEnviron 1999a;230:83–144.

Muir D, Kwan M, Lampe J. Spatial trends and pathways ofpersistent organic pollutants and metals in fish, shellfish andmarine mammals of northern Labrador and Nunavik. In:Kalhok S, editor. Synopsis of research conducted under the1999–2000 Northern Contaminants Program. Ottawa: Indianand Northern Affairs Canada; 2000. p. 191–201.

Muir DCG, Köch G, Reist J, Bright D. Temporal trends of persistentorganic pollutants and metals in landlocked char. In: Kalhok S,editor. Synopsis of research conducted under the 2000/01Northern Contaminants Program. Ottawa: Indian andNorthernAffairs Canada; 2001a. p. 202–7.

Muir D, Fisk A, Kwan M. Temporal trends of persistent organicpollutants andmetals in ringed seals from the Canadian Arctic.In: Kalhok S, editor. Synopsis of research conducted under the2000/01 Northern Contaminants Program. Ottawa: Indian andNorthern Affairs Canada; 2001b. p. 208–14.

Muir D, Wang X, Bright D, Lockhart L, Kock G. Spatial and temporaltrends of mercury and other metals in landlocked char fromlakes in the Canadian Arctic archipelago. Sci Total Environ.2005:464–78.

Naidu AS, Kelley JJ, Goering JJ, Venkatesan MI. Trace metals andhydrocarbons in Sediments of Elson Lagoon (Barrow, NorthwestArctic Alaska) as related to the Prudhoe Bay industrial region,OCS Study MMS 2003-057, Final report; 2003.

National Research Council. Toxicological effects of methylmercury. Washington, DC: National Academy Press; 2000.

NAS. Toxicological effects of methylmercury. Washington DC:National Academy of Science; 2000. Available: http://nap.edu.books/03097142/html (accessed February 2006).

NilssonA. Arctic pollution issues: a state of theArctic environmentreport. Oslo: Arctic Monitoring and Assessment Programme,AMAP; 1997. 188.

Norheim G, Skaare JU, Wiig Ø. Some heavy metals, essentialelements, and chlorinated hydrocarbons in polar bear (Ursusmaritimus) at Svalbard. Environ. Pollut 1992;77:51–7.

http://nap.edu.books/03097142/html

http://nap.edu.books/03097142/html


Norstrom RJ, Schweinsberg RE, Collins BT. Heavy metals andessential elements in livers of the polar bear (Ursus maritimus)in the Canadian Arctic. Sci Total Environ 1986;48:195–212.

O'Conner DJ, Nielsen SW. Environmental survey of methylmercurylevels in wild mink (Mustela vison) and otter (Lutra Canadensis)from the north-eastern United States and experimentalpathology ofmethylmercurialism in the otter. In: Champman JA,PursleyD,editors.Worldwide furbearers conferenceproceedings.Frostburg, M.D: Worldwide Furbearer Conference; 1981. p. 1728.

Outridge PM, Stern GA, Hamilton PB, Percival JB, McNeely R,Lockhart WL. Trace metal profiles in varved sediment of anArctic lake. Geochim Cosmochim Acta 2005;69:4881–94.

Outridge PM, Sanei H, Stern GA, Hamilton PB, Goodarzi F. Evidencefor control of mercury accumulation rates in Canadian highArctic lake sediments by variations of aquatic primaryproductivity. Environ SciTechnol 2007;41:5259–65.

Pacyna EG, Pacyna JM, Steenhuissen F, Wilson S. Globalanthropogenic mercury emission inventory for 2000. AtmosEnviron 2006;40:4048–63.

Peakall DB, Noble DG, Elliott JE, Sommers JD, Erickson G.Environmental contaminants in Canadian Peregrine Falcons,Falco peregrinus: a toxicological assessment. Can Field-Nat1990;104:244–54.

Peterson BJ, Fry B. Stable isotopes in ecosystem studies. Ann RevEcol Syst 1987;18:293–320.

Phillips DH. Quantitative biological indicators: their use tomonitortrace metal and organochlorine pollution. London: AppliedScience Publishers; 1980.

Poissant L. Atmospheric mercury transport, oxidation and falloutin northern Québec (Nunavik): an important potential route ofcontamination. Northern Contaminants Program, synopsis ofresearch 1999–2000. Montréal, Canada: Indian and NorthernAffairs Canada; 2000. p. 132–6. 0-662-29320-7.

Poissant L, Hoenninger G. Atmospheric mercury and ozonedepletion events observed at the Hudson Bay in northernQuebec along with BrO (DOAS) measurements. RMZ-MaterGeoenviron 2004;51:1722–4.

Poissant L, Dommergue A, Ferrari CP. Mercury as a globalpollutant. J Phys, IV France 2002;12:143–60.

Poissant L, Pilote M, Xu XH, Zhang H, Beauvais C. Atmosphericmercury speciation and deposition in the Bay St. Francoiswetlands. J Geophys Res 2004;109(D11):D11301.

_________________________doi:10.1029/2003JD004364.Poissant L, Pilote M, Beauvais C, Constant P, Zhang HH. A year of

continuous measurements of three atmospheric mercuryspecies (GEM, RGM and pHg) in southern Québec, Canada.Atmos Environ 2005;39:1275–87.

Poissant L, PiloteM, Constant P, Aulagnier F, ZhangH.Mercury in air,water and sediments across the Canadian Arctic: preliminaryresults fromArcticNet'sAmundsen icebreaker cruise in 2005. The8th International Conference on Mercury as a Global Pollutant,August 2006, Wisconsin, USA; 2006.

Poissant, L., Deloos, A.C., Beauvais, C., Pilote, Multi-yearObservations of Total Gaseous Mercury Along a South to NorthTransect (parallels: 45°; 50°; 55°) in Québec, Canada. 2008.Atmospheric Envrionment (in preparation).

Pongratz R, Heumann KG. Production of methylated mercury andlead by polar macroalgae—a significant natural source foratmospheric heavy metals in clean room compartments.Chemosphere 1998;36(9):1935–46.

Poole KG, Elkin BT, Bethke RW. Environmental contaminants inwild mink in the Northwest Territories, Canada. Sci TotalEnviron 1995;160/161:473–86.

Poole KG, Elkin BT, Bethke RW. Organochlorine and heavy metalcontaminants in wild mink in western northwest territories,Canada. Arch Environ Contam Toxicol 1998;34:406–13.

Poulain AJ, Garcia E, Amyot M, Campbell PGC, Ariya PA. Mercurydistribution, partitioning and speciation in coastal vs. inlandHigh Arctic snow. Geochim Cosmochim Acta 2007a;71:3419–31.

Poulain AJ, Ni Chadhain SM, Ariya PA, Amyot M, Garcia E,Campbell PGC, et al. Potential for mercury reduction bymicrobes in the high Arctic. Appl. Environ. Microbiol.2007b;73(7):2230–8.

Poulain AJ, Roy V, Amyot M. Influence of temperate mixed anddeciduous tree covers on Hg concentrations and photoredoxtransformations in snow. Geochim Cosmochim Acta2007c;71:2448–62.

Riget F, Asmund G, Aastrup P. Mercury in Arctic char (Salvelinusalpinus) populations from Greenland. Sci Total Environ2000a;245:161–72.

Riget F, Asmund G, Astrup P. The use of lichen (Cetraria nivalis) andmoss (Rhacomitrium lanuginosum) as monitors for atmosphericdeposition in Greenland. Sci Total Environ 2000b;245:137–48.

Riget F, Dietz R, Johansen P, Asmund G. Lead, cadmium, mercuryand selenium in Greenlandmarine biota and sediments duringAMAP phase 1. Sci Total Environ 2000c;245:3–14.

Riget F, Muir D, Kwan M, Savinova T, Nyman M, Woshner V, et al.Circumpolar pattern of mercury and cadmium in ringed seals.Sci Total Environ 2005;351–352:312–22.

Robbins JA. In: Nriagu J, editor. Biogeochemistry of lead in theenvironment. Amsterdam: Elsevier; 1978. p. 285–393.

Robillard S, Beauchamp G, Paillard G, Bélanger D. Levels ofcadmium, lead, mercury and 137cesium in caribou (Rangifertarandus) tissues from Northern Québec. Arctic 2002;55:1–9.

Scheuhammer AM, Wong AlHK, Bond D. Mercury and seleniumaccumulation in common loons (Gavia immer) and commonmergansers (mergus merganser) from eastern Canada. EnvironToxicol Chem 1998;17:197–201.

Schmidt D, Freimann P. Determination of very low levels ofmercury in sea water (North Sea, Baltic Sea, northern NorthAtlantic Ocean). Freseniu's Z Anal Chem 1984;317:385–7.

Schroeder WH, Anlauf KG, Barrie LA, Lu JY, Steffen A,Schneeberger DR, et al. Arctic springtime depletion of mercury.Nature 1998;394:331–2.

Scott KJ. Bioavailable mercury in Arctic snow determined by alight-emitting mer-lux bioreporter. Arctic 2001;54(1):92–101.

Segawa T,Miyamoto K, Ushida K, AgatamK, OkadaN, Kohshima S.Seasonal change in bacterial flora and biomass in mountainsnow from the Tateyama Mountains, Japan, analyzed by 16SrRNA gene sequencing and real-time PCR. Appl EnvironMicrobiol 2005;71(1):123–30.

Selifonova O, Burlage R, Barkay T. Bioluminescent sensors fordetection of bioavailable Hg(II) in the environment. ApplEnviron Microbiol 1993;59(9):3083–90.

Semkin RG, Mierle G, Neureuther RJ. Hydrochemistry andmercurycycling in a high Arctic watershed. Sci Total Environ2005;342:199–221.

Sergeant DE, Armstrong FAJ. Mercury in seals from easternCanada. J Fish Res Board Can 1973;30:843–6.

Sergeant DE, Brodie PF. Identity, abundance, and present status ofpopulations of white whales, Delphinapterus leucas, in NorthAmerica. J Fish Res Board Can 1975;32:1047–54.

Shilts WW, Coker WB. Mercury anomalies in lake water and incommercially harvested fish, Kaminak Lake area, district ofKeewatin, Canada. Water Air Soil Pollut 1995;80:881–4.

SkovH,Christensen JH,GoodsiteME,HeidamNZ, JensenB,WahlinP,et al. Fate of elementalmercury in the Arctic during atmosphericmercury depletion episodes and the load of atmosphericmercury to the Arctic. Environ Sci Technol 2004;38(8):2373–82.

SkovH, Brooks SB, GoodsiteME, Lindberg SE,Meyers TP, LandisMS,et al. Fluxes of reactive gaseous mercury measured with anewly developed method using relaxed eddy accumulation.Atmos Environ 2006;40:5452–63.

Slemr F, Brunke EG, Ebinghaus R, Temme C, Munthe J, Wängberg I,et al. Worldwide trend of atmospheric mercury since 1977.Geophys Res Lett 2003;30(10):1516. _________________________doi:10.1029/2003GL016954.

Smith TG, Armstrong FA. Mercury and selenium in ringed andbearded seal tissues from arctic. Canada 1978;31:75–84.

http://dx.doi.org/10.1029/2003GL016954


Sommar J,Wänberg I, Berg T, Gårdfeldt K, Munthe J, Richter A, et al.Circumpolar transport and air-surface exchange ofatmospheric mercury at Ny-Ålesund (79oN), Svalbard, spring2002. Atmos Chem Phys Discuss 2004;4:1727–71.

St. Aubin DJ, Smith TG, Geraci JR. Seasonal epidermal molt inbeluga whales, Delphinapterus leucas. Can J Zool 1990;69:359–67.

Steffen A, Schroeder W, Bottenheim J, Narayan J, Fuentes JD.Atmospheric mercury concentrations: measurements andprofiles near snow and ice surfaces in the Canadian Arcticduring Alert 2000. Atmos Environ 2002;36:2653–61.

Steffen A, Schroeder W, Macdonald R, Poissant L, Konoplev A.Mercury in the Arctic atmosphere: an analysis of eight years ofmeasurements of GEM at Alert (Canada) and a comparisonwith observations at Amderma (Russia) and Kuujjuarapik(Canada). Sci Total Environ 2005;342:185–98.

Steffen A, Douglas T, Amyot M, Ariya P, Aspmo K, Bergm T, et al. Asynthesis of atmospheric mercury depletion event chemistrylinking atmosphere, snow and water. Atmos Chem PhysDiscuss 2007;7:10837–931.

Steinnes E, Sjobakk TE. Order-of-magnitude increase of Hg inNorwegian peat profiles since the outset of industrial activityin Europe. Environ Pollut 2005;137:365–70.

Stephens GR. Mercury concentrations in fish in a remote CanadianArctic Lake. Water Air Soil Pollut 1995;80:633–6.

Stern GA, Lockhart WL, Ikonomou M. Temporal trends oforganochlorine, organobromine and heavy metal contaminantsin burbot from Fort Good Hope, Northwest Territories. In:Kalhok S, editor. Synopsis of research conducted under the 2000/01 Northern Contaminants Program. Ottawa: Indian andNorthern Affairs Canada; 2001. p. 230–6.

St. Louis VL, Rudd JWM, Kelly CA, Barrie LA. Wet deposition ofmethyl mercury in northwestern Ontario compared to othergeographic locations. Water Air Soil Pollut 1995;80:405–14.

St. Louis VL, Sharp MJ, Steffen A, May A, Barker J, Kirk JL, et al.Some sources and sinks ofmonomethyl and inorganicmercuryon Ellesmere Island in the Canadian high Arctic. Environ SciTechnol 2005;39(8):2686–701.

St. Louis VL, Hintelmann H, Graydon JA, Kirk JL, Barker J, Dimock B,et al. Methylated mercury species in Canadian high Arcticmarine surface waters and snowpacks. Environ Sci Technol2007;41(18):6433–41.

Stout JH, Trust TA. Elemental and organochlorine residues in baldeagles from Adak Island. Alaska J Wildlife Dis 2002;38:511–7.

Swartout J, Rice G. Uncertainty analysis of the estimated ingestionrates used to derive the methylmercury reference dose. DrugChem Toxicol 2000;23:293–306.

TemmeC,Baukau J, Schneider B,AspmoK,FainX, FerrariC, et al. Air/water exchange of mercury in the North Atlantic Ocean duringArctic summer. Extended abstract for the XIII International onHeavy Metals in the Environment, Rio de Janeiro; 2005.

Thomas DJ, Macdonald RW, Cornford AB. Canadian data report.Hydrogr Ocean Sci 1982;5:243.

Thomas DJ, Tracey B, Marshall H, Norstrom RJ. Arctic terrestrialecosystem contamination. Sci Total Environ 1992;122:135–64.

Thompson DR. Mercury in birds and terrestrial mammals. In:Beyer WN, Heinz GH, Redmon-Norweed AW, editors.Environmental contaminants in wildlife: interpreting tissueconcentrations. SETAC Special Publication SeriesCRC Press,Inc.; 1996. p. 341–56.

Thompson DR, Furness RW. Comparison of the levels of total andorganic mercury in seabird feathers. Mar Pollut Bull1989;20(11):577–9.

Trust KA, Rummel KT, Scheuhammer AM, Brisbin IL, Hooper MJ.Contaminant exposure and biomarker responses in spectacledelders (Somateria fischeri) from St. Lawrence Island, Alaska. ArchEnviron Contam Toxicol 2000;38:107–13.

UNEP. United Nations Environment Programme (UNEP). Globalenvironment outlook. (GEO)-1, Chapter 2: regionalperspectives. Oxford University Press; 1997.

USEPA. Great LakesWater Quality Initiative Criteria documents forthe Protection of Wildlife. EAP-820-B-95-008. Washington DC:US Environmental Protection Agency, Office of Science andTechnology; 1995.

VertaM,Matilainen T, Porvari P, NiemiM, Uusi-RauvaA, BloomNS.Methlymercury sources in boreal lake ecosystems. In: WatrasCJ, Huckabee JW, editors. Mercury pollution—integration andsynthesis. Ann Arbor: Lewis; 1994. p. 119–36.

VincentWF,Gibson JAE, PienitzR,VilleneuveV,BroadyPA,HamiltonPB, et al. Ice shelf microbial ecosystems in the High Arctic andimplications for life on snowball Earth. Naturwissenschaften2000;87:137–41.

Vonk JW, Sijpesteinjn AK. Studies on the methylation of mercuricchloride by pure cultures of bacteria and fungi. Antonie vanLeeuwenhoek 1973;39:505–13.

Wagemann R. Comparison of trace metals in two groups of ringedseal (Phoca hispida) from the Canadian Arctic. Can J Fish AquatSci 1989;46:1558–65.

Wagemann R, Kozlowska H. Mercury distribution in the skin ofbeluga (Delphinapterus leucas) and narwhal (Monodon monoceros)from the Canadian Arctic and mercury burdens and excretionby moulting. Sci Total Environ 2005;351–352:333–43 Volumes.

Wagemann R, Stewart REA, Béland P, Desjardins C. Heavy metalsand selenium in tissues of beluga whales, Delphinapterus leucas,from the Canadian Arctic and the St. Lawrence estuary. In:Smith TG, St. Aubin DJ, Geraci JR, editors. Advances in researchon the beluga whale, Delphinapterus leucas, vol. 224. Can BullFish Aquat Sci; 1990. p. 191–206.

Wagemann R, Lockhart LW, Welch H, Innes S. Arctic marinemammals as integrators and indicators of mercury in theArctic. Water Air Soil Pollut 1995;80:683–93.

Wagemann R, Innes S, Richard PR. Overview and regional andtemporal differences of heavy metals in Arctic whales andringed seals in the Canadian Arctic. Sci Total Environ1996;186:41–66.

Wängberg I,Munthe J, BergT, EbinghausR, KockHH, TemmeC, et al.Trends in air concentration and deposition of mercury in thecoastal environment of the North Sea Area. Atmos Environ2007;41:2612-1619.

Wayland M. Contaminants in Arctic seaducks. In: Kalhok S, editor.Synopsis of research conducted under the 1998–1999 NorthernContaminants Program. Ottawa: Indian and Northern AffairsCanada; 1999. p. 213–22.

Wayland M. Contaminants in Arctic seaducks. In: Kalhok S, editor.Synopsis of research conducted under the 1999–2000 NorthernContaminants Program. Ottawa: Indian and Northern AffairsCanada; 2000. p. 237–44.

Wayland M, Gilchrist HG, Dickson DL, Bollinger T, James C,Carreno RA, et al. Trace elements in king eiders and commoneiders in the Canadian Arctic. Arch. Environ Contam Toxicol2001;41:491–500.

Wayland M, Gilchrist HG, Neugebauer E. Concentrations ofcadmium, mercury and selenium in common eider ducks inthe eastern Canadian arctic: influence of reproductive stage.Sci Total Environ 2005;351/352:323–32.

Wedephol KH, Correns CW, Shaw DM, Turekian KK, Zemann J.Handbook of geochemistry. New York: Springer Verlag; 1978.Vol. II/s, 80-C-l to 80-O-4.

Weiss HV, Chew K, Guttman M, Host A. In: Reed JC, Slater JE,Gunn WW, editors. Proceedings of a Symposium on BeaufortSea Coast and Shelf Research. Arctic Institute of NorthAmerica; 1974. p. 737–46.

WHO. Evaluation of certain food additives and contaminants.Thirty-third report of the Joint FAO/WHO Expert Committee onFood Additives. Technical Report SeriesGeneva: World HealthOrganization; 1989.

WiseMH, Linn IJ, Kennedy CR. A comparison of the feeding biologyof mink Mustela vison and otter Lutra lutra,. J Zool Lond1981;195:181–213.


Wobeser G, Swift M. Mercury poisoning in a wild mink. J Wildl Dis1976;12:335–40.

Wong MP. Chemical residues in fish and wildlife harvested inNorthern Canada. Environmental StudiesOttawa: Indian andNorthern Affairs Canada; 1985.

Wolfe AP. A 6.500-year diatom record from southwestern FosheimPeninsula, Ellesmere Island, Nunavut. Geol Surv Can Bull2000;529:249–56.

Wood PB, White JH, Steffer A, Wood JM, Percival HF. Mercuryconcentrations in tissues of Florida bald eagles. J Wildl Manage1996;60:178–85.

World Health Organization (WHO). Air quality guidelines forEurope. Second edition. European seriesWHO RegionalPublications 92-890-1358-3; 2000.

Wren CD. Distributions of metals in tissues of beaver, raccoon andotter from Ontario, Canada. Sci Total Environ 1984:177–84.

Yannai S, Berdicevsky I, Duek L. Transformations of inorganicmercury by Candida albicans and Saccharomyces cerevisiae. ApplEnviron Microbiol 1991;57(1):245–7.

Young DR, Johnson JN, Soutar A, Isaacs JD. Mercury concentrationsin dated varved marine sediments collected off SouthernCalifornia. Nature 1973;274:273–4.

Documents

Critical review of mercury fates and contamination in the arctic tundra ecosystem