the competitive role of organic carbon and dissolved sulfide in controlling the distribution of...

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The competitive role of organic carbon and dissolved sulfidein controlling the distribution of mercury in freshwaterlake sediments

Nelson Belzilea,b,⁎, Chun-Yan Langc, Yu-Wei Chena, Mohui Wangc

aDepartment of Chemistry and Biochemistry, Laurentian University, Sudbury, Ontario, Canada P3E 2C6bCooperative Freshwater Ecology Unit, Laurentian University, Sudbury, Ontario, Canada P3E 2C6cDepartment of Applied Chemistry and Bioengineering, Chengdu University of Technology, Chengdu, 610059, China

A R T I C L E I N F O

⁎ Corresponding author. Tel.: +1 705 675 1151xE-mail address: [email protected] (N.

0048-9697/$ – see front matter © 2008 Elsevidoi:10.1016/j.scitotenv.2008.06.034

A B S T R A C T

Article history:Received 13 February 2008Received in revised form17 June 2008Accepted 20 June 2008Available online 26 July 2008

The detailed distribution of mercury was studied in sediments and porewaters of twofreshwater lakes, which were selected because of the contrasting conditions they present attheir respective sediment–water interface (SWI). One lake is characterized by a SWI thatremains oxic all year long whereas the other one shows a clear seasonal variation with theevolution of strongly anoxic conditions through the summer season. The results of the studyclearly identify the importance of redox conditions on the geochemical behaviour ofHgat theSWI of both lakes but a very limited influence of an oxidized layer enriched in Fe and Mnoxyhydroxides at the top of the sediment of the oxic lake. In both lakes, a competitive effecton the cycling andmobility of the elementwas observed between natural organicmatter andamorphousor organo-sulfide compounds. Theproportion ofHgassociated tonatural organicmatter in sediments showed a general increase with sediment depth. A fraction containingelemental Hg and Hg suspected to be bound to iron sulfides and organo-sulfides constitutedthe othermajor fraction of solidHg in the sediments of both lakes. This secondpool ofHgwasgenerally larger at the top of the sediment where the production of dissolved sulfides isusually more detectable and it decreases with depth, suggesting that the metal is partiallytransferred fromonepool being the sulfides including amorphous FeS andorgano-sulfides tothe organic matter pool. Methyl Hg represented less than 1% of the total Hg in sediments ofboth lakes. Our results obtained at different times of the summer season from two lakescontrasted by their SWI emphasize the competitive or alternating role played by dissolvedand solid natural organic matter and sulfides on the fate of Hg in freshwater systems.

© 2008 Elsevier B.V. All rights reserved.

Keywords:MercuryLake sedimentsPorewatersGeochemistry

1. Introduction

Mercury occurs naturally in its elemental form in the atmo-sphere and in sulfidic ores such as cinnabar (red HgS) inminerals. The terrestrial abundance of Hg averages 0.05 ppmor μg g−1 (Jonasson and Boyle, 1971) but the content varies withrock and sediment types (Turekian and Wedepohl, 1961).

2114; fax: +1 705 675 484Belzile).

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Originating from natural and anthropogenic sources, Hg isalso present in ambient air, in natural waters and in mostanimal and plant tissues (Fitzgerald and Lamborg, 2003). In theaquatic environment, mercury exists in various forms and it isbelieved that transformations and mobilization are normallycontrolled by a combination of several factors (Hudson et al.,1994). The partitioning of Hg between the dissolved, colloidal

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227S 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 5 ( 2 0 0 8 ) 2 2 6 – 2 3 8

and particulate phases varies widely spatially, seasonally andwith depth in the water column and sediments. In general, andparticularly in stratified systems, concentrations of Hg0 arehigher near the air–water interface whereas levels of total HgandMeHg are higher near the sediment (Morel et al., 1998). Themajor formof ionicHg in oxicwater isHg(II),which canpossiblybe complexed in variable amounts to hydroxide or chloridedepending on pH and ion concentration and to organic matter.Under anoxic conditions inwaters and sediments, themercuricion has very high affinity for sulfide, therefore the speciation ofdissolved Hg2+ in sulfidic waters is completely dominated bysulfide (HgS) and sulfide complexes such as Hg(HS)+, Hg(HS)2,HgSo, Hg(HS2)− and Hg(S2)2− (Morel et al., 1998; Benoit et al.,1999). However, in the presence of dissolved organic matter(DOM), reduced sulfur sites of DOM can bind strongly with Hgand even compete with inorganic ligands including free sul-fides (Ravichandran, 2004). This strong binding of Hg to DOMordissolved organic ligands affects its speciation, solubility,mobility and availability in aquatic systems. The two knownforms of solid mercuric sulfide HgS(s) cinnabar and meta-cinnabar both have a very low solubility product, and HgS(S) isthought to be the particulate mercury species that is buriedin sediments and the formation of HgS would control Hg2+

solubility in anoxic systems. In spite of the extremely lowsolubility of cinnabar, its actual solubility can be modified athigh sulfide concentration, due to the formation of the dis-solved sulfide and bisulfide mercuric complexes (Ravichan-dran, 2004; Paquette and Helz, 1997). This increasing solubilityof Hg with sulfide concentration undoubtedly plays a role inincreased dissolved mercury concentration observed in manyanoxic waters.

Thedistributionand speciationofHg in soils and sedimentsis of great significance for understanding its transportation,transformation, bioavailability and toxicity. Redox boundariesare the site of biological transformations where Hg can bemethylated under reducing conditions to produce solubleCH3Hg+ and volatile organo-mercury species such as (CH3)2Hg(Paquette and Helz, 1997; Ullrich et al., 2001). Usually themobility, bioaccumulation and toxicity in soil and sedimentincrease in the direction (Han et al., 2003): alkyl HgNsolubleinorganicHgNelemental Hg andHg-metal amalgamNmercuricsulfide, but there are exceptions to this rule. In general, Hgspeciation may be divided into two categories: (1) chemicalspeciationwhich classifies Hg compounds on the basis of theirchemical structure and affinity toward other compounds;(2) physical speciation based on the physical properties of Hgspecies with any potential carrier. Solid phase chemical spe-ciation based on sequential and selective extraction proce-dures is critical to understanding metal-contaminatedsystems and assessing metal distribution over various sedi-mentary phases. Even though inherent limitations exist due topossible re-adsorption or insufficient selectivity (Belzile et al.,1989; Nirel and Morel, 1990; Biester and Scholz, 1997) andbecause no other techniques are sensitive enough to provideinformation on the distribution of ametal such asHgwhen it ispresent at very low levels (low nmol/g) in complex matrices,the sequential extraction procedure remains a very usefultechnique to provide valuable information on the distributionof a trace element within complex matrices such as soil andsediment. A variety of sequential extraction schemes have

been designed for the speciation of Hg in solid substrates(Lechler et al., 1997; Wallschläger et al., 1998; Sahuquillo et al.,2003; Bloom et al., 2003), most of the time for contaminatedsites. Although there are some differences in the nature orstrength of the selected extractants, the main operationalprotocols are rather similar. In this study, the detailed depth-distribution of Hg species in porewaters and sediments wasinvestigated using porewater peepers and amodified selectiveextraction procedure of sediments initially proposed by Bloomet al. (2003). We intended to compare the geochemical be-haviour of Hg in two lakes that are not significantly con-taminated with this metal but very distinct in their respectivesediment–water interface (SWI). A particular attention wasgiven to seasonal variations of the redox conditions at the SWIof each lake and to the influence of increasing reducingconditionswith time in one of the two lakes on the distributionand evolution of Hg in sediments and corresponding pore-waters. This field study also includes the measurement of awhole series of parameters including iron and manganesecompounds to determine what role they could play in theoverall distribution of Hg in freshwater sediments.

2. Methodology

2.1. Sampling sites

Sediment and porewater samples were collected from twofreshwater lakes located in Sudbury, Ontario in June andSeptember of 2003. Clearwater Lake (46° 22′ N; 81° 03′ W) wasstrongly acidified in the past by atmospheric emissions of SO2

but its pH has gradually changed from to 4.2 in the 1970's toapproximately 6.4 nowadays. McFarlane Lake (46° 25′N, 80° 57′W) is a well buffered slightly alkaline lake at pH=7.5, locatedonly 5 km away from Clearwater L. and receives similar atmo-spheric loading of trace elements, mainly, Ni, Cu, Zn and Pbfrom the Sudbury smelters (Nriagu et al., 1982; Carignan andNriagu, 1985). The local contamination by Hg is minimal dueto its low concentration in the ore (b0.1 µg/g, unpublishedresults). In both lakes, sampleswere collected at littoral sites ofapproximately 8m in depth. These two lakes were selected forthis studybasedon criteria of acidification level andoxic statusat the sediment–water interface (SWI) defined in this study asthe first few centimetres above and within the sediment.Previous geochemical studies carried on selenium (Belzileet al., 2000) and antimony (Chen et al., 2003) suggest that theSWI of the sampling site of McFarlane L. is characterized bymuch more reducing conditions in late summer (lower dis-solved oxygen levels) as compared to that of Clearwater L.,which remain well oxygenated all year long.

2.2. Cleaning procedures

Due to its volatile nature, mercury is one element that is mostvulnerable to atmospheric contamination. To avoid it, allmaterials and vessels used for analytical purpose must bechosen carefully and cleaned appropriately before samplingand solution preparation. It was found that a preliminarycleaning of all sampling devices and vessels with diluteddetergent can remove an important fraction of adsorbed Hg.

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Porewater collectors (peepers) were first washed in soapywater and then rinsed with double deionized water (DDW),then soaked in 6MHCl for 4 h and finally rinsedwithDDW. In asecond step, they were soaked in 5% (v/v) HNO3 for 3 days andabundantly rinsed with DDW. The 0.2 μm filtrationmembranewas successively soaked inDDW for 1 day and in 2% (v/v) HNO3

for 3 days and finally rinsedwithDDW.TheTeflon vials used tocollect porewaters for total Hg were first cleaned by addingabout 5 mL of the BrCl solution for more than 30 min andshaking them several times during this period. The vials wererinsed with tap water and DDW at least 8 times. Finally, theywere fully filled with fresh DDW, tightly capped and stored indouble plastic bags until being used shortly after. The 2-mLglass bottles for dissolved species such as organic carbon(DOC), sulfide, Fe and Mn were soaked in 10% (v/v) HCl for 6 h,and then abundantly rinsed with DDW. For DOC determina-tion, the glass bottles should further be rinsed with specialTOC-free deionized water at least three times and dried in aclean oven at about 120 °C for 5 h. After cooling, the bottleswere tightly capped until use.

2.3. Porewater and sediment sampling

Porewater samples were collected using in situ diffusion sam-plers (peepers) that had been previously filled with DDW anddeoxygenated by bubbling suprapure N2 gas for 72 h, immersedinacontainerofdeoxygenatedDDWwater beforebeing insertedin the sediment by a diver. The peepers were allowedequilibrating in sediment with interstitial waters for at least14 days (Carignan and Nriagu, 1985; Belzile et al., 2000). Twopeepers were attached back to back in order to collect largervolumes of interstitialwater at each corresponding depth (Chenet al., 2003). After equilibration, the peepers were retrieved fromthe sediment by the diver and water samples were collectedafter piercing the 0.2 µm filtration membrane with the tip of apre-cleaned micropipette. To minimize the contact with air,which could cause change of acidity and loss of volatile sulfide,the two first porewater subsamples were collected for pH anddissolved sulfide. pHmeasurementswere done immediately onthe boat in a small plastic tube containing 1.0mL ofwaterwith apointed small combined pH electrode. For dissolved sulfide, a1.0-mLporewater subsamplewas immediately transferred fromthepeepers to a 2-mLglass bottle already containing the amine-sulfuric acid fixing agent. Upon return to the laboratory, theprocedure was completed to measure the dissolved sulfidecontent by visible spectrophotometry at 670 nm (APHA, 1992;Carignan et al., 1985). To measure Fe and Mn, another 1-mLsubsamplewas pipetted into a 5-mL glass tubewhich contained100 μL of 20% (v/v) ultrapure HCl. For the determination ofdissolved organic carbon (DOC), a 1.5 mL aliquot was pipettedinto a 2-mL glass tube inwhich 10 μL of concentrated H3PO4 hadbeen added before sampling. For total Hg measurements inporewaters, a total volume of 10.0 mL was collected at eachdepth, which was then transferred into a pre-cleaned 30-mLTeflon vial. All Hg samples were kept on ice during the shorttransfer to the laboratory where 0.2 mL of a BrCl solution wasadded to each sample to convert all forms of Hg into Hg2+. Thetotal dissolved Hg was measured after a 2-d digestion at roomtemperature (Lang et al., 2005). Undisturbed cores of sedimentswere carefully collected by the diverwith a lightweight Plexiglas

corer at sites close to the porewater sampling location. Aftercapping and retrieval, cores were immediately transported tothe laboratory for extrusion under N2 atmosphere. Cores wereentirely sliced into 1.0-cm sections. In order to minimize redoxmodifications, the subsamples were placed in polyethylenebottles and frozenat−80 °Cuntil further treatmentandanalysis.

2.4. Sample treatment and analysis

For Hg measurements in porewaters, a Tekran model 2600cold vapour atomic fluorescence spectrophotometer with adual-stage gold amalgamation system was used. A protocoldeveloped in our laboratory was used to determine total Hg inporewaters and our results have shown that a digestionwith a2% (v/v) BrCl at room temperature was efficient in controllingpossible interferences due to dissolved organic ligands andsulfide (Lang et al., 2005). Dissolved concentrations of Fe andMn were obtained by flame and graphite atomic absorptionspectrometry and dissolved oxygen in overlying waters wasdetermined using the Wrinkler method. A Dohrmann DC-80total carbon analyzer was used for the determination of DOCin porewaters. A sequential method modified from Bloomet al. (2003) was used to study the distribution of Hg in the solidphase of the sediment. Extractions were carried out using400 mg of Clearwater L. or 200 mg of McFarlane L. freeze-driedand finely ground sediment samples in 30-mL clean Tefloncentrifugation tubes. A rinse with DDWwas included betweeneach step of the extraction procedure. For each extraction,16.0 mL of extracting solution was added to the sedimentsamples and the extraction was carried out for 18±2 h at roomtemperature with a wrist arm-shaker at 240 times per minute.The two first extractions bywater and acetic/hydrochloric acidof the Bloom protocol were replaced by a single extractionusing a 0.2 M oxalic acid solution buffered to pH 2 withammonium oxalate to remove Hg mainly bound to amor-phous and poorly crystalline Fe and Mn oxyhydroxides(Schwertmann, 1964; Borggaard, 1992). This fraction is definedas Hg–Ox in our study. Thismodification should have a limitedimpact on the overall process since the acidic Fraction F2 ofthe Bloom protocol should also extract some Fe and Mnoxides. The rest of the sequence was identical to the Bloomextraction procedure with a second step using a 1.0 M KOHsolution to remove Hg bound to organic matter and defined asthe Hg–KOH fraction. It was then followed by an extractionwith a cold 12 M HNO3 solution to extract Hg(0) and Hg boundto amorphous organo-sulfur or sulfides from the sedimentand defined as the Hg–HNO3 fraction. The final step was anovernight treatment of the residual with aqua regia at roomtemperature followed by 2.5 h on the hot plate at 125 °C toobtain the fraction identified as Hg–AR. This last stepdigestion was also used on a separate sediment subsampleto determine the total aqua regia concentration of Hg. Mercuryin each extracted fraction was determined by cold vapouratomic fluorescence spectrometry after BrCl digestion andelimination of the excessive BrCl by a 5% (w/v) NH2OH·HClsolution (Lang et al., 2005; Chen et al., 2002).

Methyl Hg was first extracted into CH2Cl2 in an alkalinesolution, back extracted into an aqueous solution, and thenethylated to form methyl–ethyl mercury, which was pre-concentrated on a Tenax trap. The compound was then


released from the trap by heating and delivered to a gas chro-matography column to be separated, decomposed by pyrolysisas Hg0 and determined by CVAFS (Bloom, 1989). Fe and Mnextracted along with Hg in all sediment fractions weremeasured by flame AAS and defined as Fe–Ox and Mn–Oxwhen extracted by the oxalate buffer, as Fe–KOH and Mn–KOHwhen extracted by KOH; the fractions extracted by cold HNO3

were defined as Fe–HNO3 and Mn–HNO3 and the residualfractionswere defined as Fe–ARandMn–ARafter the aqua regiatreatment. For the determination of total concentrations ofHg, Fe and Mn, all sediment samples were digested in Teflonbombs according to a procedure previously described (Belzileet al., 2000; Chen et al., 2003). To estimate the fraction of totalorganic carbon (TOC) in sediments, a known mass (∼0.5 g) ofoven-dried (40 °C) sediment was subjected to a temperature of750 °C in amuffle furnace for 4 h. The estimatedTOCcontent orloss on ignition (LOI) was obtained from the difference ofsample weight before and after ashing. This is considered as avalid estimation of TOC in low carbonate sediments.

3. Results and discussion

A typical calibration curve for the determination of Hg inporewater shows that the blank value was around 165 in peakarea compared to a value of 383 for the lowest standard of1.0 ng/L or 5 pM. The R2 value obtained with 4 standards, thehighest being 500 pM, is typically 0.9999. A relative standarddeviation 4.8% was obtained on repeated measurements oflake and pore waters showing a concentration close to 5.0 pM.Standard additions on aliquots of a porewater sample showedrecoveries between 95 and 106%. The quality of the digestionand analysis was controlled through repeated determinationsafter every 10 samples of two certified standard materials(CRM) of sediment (PACS-2 and MESS-3) from the NationalResearch Council of Canada. Recoveries in the CRM werebetween 96 and 99% for total Hg. For the sequential extractionprocedure, the addition of all extracted fractions and residual

Fig. 1 –Profiles of pH in porewaters of C

led to values varying between 92 and 98% of the total Hg mea-sured separately, for each sample and the relative standarddeviation on repeated extractions was less than 8%.

3.1. Redox status of the sediment–water interface (SWI)

The pH and redox status of the two chosen lakes and theireffects on the geochemical behaviour of Se (Belzile et al., 2000)and Sb (Chen et al., 2003) have previously been reported.Several parameters measured in this study confirmed againthe differences in pH and redox conditions existing at the SWIof the two lakes. The SWI of Clearwater L. was clearly oxic withdissolved oxygen (DO) concentrations around 9.0mg/L, both inJune and September; pH at the SWIwas 5.7 and varied between5.6 and 6.2 in the sediment porewaters. The SWI of McFarlaneappeared still well oxygenated in June at 8.2 mg/L butsignificantly depleted in DO at less than 4.0 mg/L in earlySeptember. Porewaters of McFarlane L. showed more neutralpH conditions than Clearwater L. with values around 6.6 at theinterface and in porewaters for both sampling times (Fig. 1). Itshould be mentioned however, that the real DO concentrationof the SWI cannot be obtainedby theWrinklermethodbecausethe real thin anoxic layer of overlying water can be easilydisturbed and destroyed by the diver's movements whenfilling the bottle (Chen et al., 2003). The redox status of the SWIin the two study lakes was also confirmed by profiles of dis-solved Fe and Mn (Fig. 2). In Clearwater L., both dissolvedspecieswere onlymeasurable below the SWI (the thin oxidizedlayer of the sediment), as they are released under reducingconditions from the dissolution of respective oxyhydroxides.Reducingconditionsat the SWIofMcFarlaneL.were confirmedby measurable levels of dissolved Mn at ∼25 µM and Fe at∼12 µM in overlying waters (Fig. 2). Considering the neutral pHconditions of McFarlane L., concentrations of dissolved Fe anddissolved Mn were relatively high over the sediment likelybecause the low oxygen levels do not favour the kinetics ofprecipitation of Fe oxyhydroxides (Stumm and Morgan, 1996).Redox conditions suspected in overlying and pore waters are

learwater L. (a) and McFarlane L. (b).

Fig. 2 –Concentrations of dissolved Fe (a) and (c) and Mn (b) and (d) across the sediment–water interface of Clearwater L.(top panels) and McFarlane L. (bottom panels) in June and September.


also confirmed by the profiles of total and extractable Fe andMn (presented in Figs. 4–6 as Tot and Oxal, and discussedbelow). The presence of anoxic conditions at the SWI ofMcFarlane in June and September was confirmed by thenoticeable smell of H2S brought to the surface of the lake bythe diver's bubbles, which was particularly obvious during theSeptember sampling.

3.2. Effect of dissolved sulfide and DOC on porewater Hg

With the exception of one overlying water sample in McFar-lane L. collected in June, all Hg concentrations in porewaterswere ≤200 pM or ∼40 ng/L and depth profiles showed littlevariation from the surface of the sediment to 25 cm belowsurface. These porewater concentrations are comparable toother non contaminated environments such as other Ontariolakes (He et al., 2007). For comparative purpose, the distribu-

tion coefficients (Kd) were calculated with the total Hg concen-trations in the solid phase assuming a thermodynamicequilibrium between porewater and sediment. The log valuesranged between 3.76 and 4.48 with a general tendency to de-crease with depth, which is likely due to lower Hg concentra-tions in the solid phase in pre-industrial sediments (below15 cm in depth). This range of log Kd values is comparable toothers reported in the literature (Bloom et al., 1999; Ham-merschmidt and Fitzgerald, 2004; He et al., 2007;). The profilesof dissolved sulfide (Fig. 3b, e and h) show low or undetectableconcentrations of dissolved sulfide in Clearwater and McFar-lane lakes in June but much higher concentrations (more than20 μM) across the SWI of McFarlane L. in September (differentscale compared to the other two profiles). This is a conse-quence of the depletion of dissolved oxygen and reduction ofsulphate at this SWI. Calculations of ion activity product (IAP)were performed assuming that all dissolved Fe was present as

Fig. 3 –Concentrations of dissolved Hg (a, d, g), dissolved sulfide (b, e, h) and dissolved organic carbon (c, f, i) across thesediment–water interface of Clearwater L. (top panels: June) and McFarlane L. (middle: June and bottom: panel: September).



Fe2+ at depthswhere the concentrationof dissolved sulfidewasdetectable and assuming a solubility control by amorphousiron sulfide. The solubility product pKsp ([Fe2+][HS−] γFe2+ γHS−/(H+) ) for amorphous ferrous sulphide of 2.95 (Davison, 1991)was used to estimate the saturation level with respect toamorphous FeS in the September profile. Calculations weredone using concentrations of dissolved Fe2+, dissolved sulfideand converted values of pH obtained in collected porewaters(Belzile et al., 1996). The dissociation constants (I=0 andT=25 °C) used in the calculations were 10−7.02 and 10−13.90 forH2S (Smith and Martell, 1976), respectively. More recent esti-mates put the pKa2 value at less than −17. It was found that −log IAP values were very close (between 2.7 to 3.2) to that ofpKsp of amorphous ferrous sulfide of 2.95, indicating a closesaturation with respect to that solid in that zone. It is likelythat dissolved sulfide are oxidized above the concentrationpeak in the water column and precipitated as FeS or pyrite

Fig. 4 –Distributions of Fe (a), Mn (b) and Hg (c) extracted fractionin Clearwater L. in June.

(FeS2) below the peak in the sediment. These two iron sulfidecompounds can play a role in controlling the solubility of Hgby sorption or co-precipitation (Behra et al., 2001; Jeong et al.,2007). This will be further discussed later.

The concentrations of DOC in porewaters varied betweenthe two lakes and the two seasons (Fig. 3c, f and i). DOC con-centrations in Clearwater L. in June decreased from a valuearound 3 mg/L in overlying waters to a background value of 1–1.5 mg/L in sediments. When the DOC profile is compared tothat of dissolved Hg in the same lake, it suggests that higherDOC concentrations could help in maintaining higher con-centrations of dissolved Hg, when dissolved oxygen is presentand dissolved sulfides are absent. In McFarlane L., the DOCprofile in June also showed slightly higher values in overlyingwaters and a decrease in depth. However, the Septemberprofile in the sediment of this lake (Fig. 3i) suggests the remo-bilization of organic matter at depth and the upward diffusion

s in sediments, and (d) expressed as a percentage of total Hg,


of this DOC. This DOC profile and other profiles discussedbelow confirm the highly dynamic nature of the McFarlane L.basin, which is likely activated by intensifying reducing con-ditions at its SWI as season advances. The much lower con-centrations of dissolved Hg in overlying and porewaters ofMcFarlane L. in September also suggest that dissolved sulfidesat the SWI and solid Fe sulfides suspected highly to be formed(see above discussion) can be in competition with organicmatter to control the solubility of Hg under such strongly re-ducing conditions. Further work is needed to clarify the natureof the released DOC at depth although it seems to have limitedinfluence on the dissolved Hg profile.

3.3. Influence of natural organic matter (NOM) and sulfideson solid phase distribution

The surface enrichment in Fe and Mn oxyhydroxides(expressed as the Oxal fraction) in Clearwater L. (Fig. 4a and b)

Fig. 5 –Distributions of Fe (a), Mn (b) and Hg (c) and extracted fracHg, in McFarlane L. in June.

is characteristic of an oxygenated SWI. For this lake, only theJune results are shown since the September profiles were verysimilar. The situation was different at the SWI of McFarlane L.where no such surface enrichment in Fe oxyhydroxides wasobserved in June (Fig. 5a) or in September (Fig. 6a) due to theestablishment of increasing reducing conditions from the firstsampling to the second one. The Mn profiles suggest thepresence of higher concentrations of reducibleMn (Mn–Oxal) insurficial sediments in June but this fraction had significantlydecreased in September under more reducing conditions(Figs. 5b and 6b). Total Hg concentrations in both lakes weregenerally ≤1.0 nmol/g on a dry wt basis and decreased withdepth. The results of sequential extraction provide interestinginformationon thedistributionof totalHg in sediments and thesums of all Hg fractions were very close to the total concentra-tion. The fraction of Hg extracted by oxalate was very smalleven in surficial oxidized sediments of Clearwater L., whichsuggest that Fe and Mn oxyhydroxides play a minor role in

tions in sediments, and (d) expressed as a percentage of total

Fig. 6 –Distributions of Fe (a), Mn (b) and Hg (c) extracted fractions in sediments, and (d) expressed as a percentage of total Hg,in McFarlane L. in September.


controlling the solubility of Hg. The results of the extractionrather suggest a competitive role of natural organic matter and(iron) sulfides in controlling the solubility of Hg.

Many studies provide experimental evidence for the com-plexation of Hg with natural organic matter (NOM) in naturalwaters and in soils (e.g. Ravichandran, 2004) and a study basedon selective extractions in peat soils and sediments report thata large fractionwas associatedwith organicmatter, particularlythe humic/fulvic and organic-sulfide bound fractions (Di Giulioand Ryan, 1987). It is therefore not surprising to measure animportantproportionofHg in sedimentsof both lakes extractedbyKOH (Figs. 4c, 5c and6c). The estimated concentrationofTOCor LOI (Fig. 7) can reach 30% in mass in the sediments ofMcFarlane L. and around 15% in the other lake. The proportionof Hg attached to the so-called humic organicmatter as definedin Bloom et al. (2003) varied from 28 to 72% of total Hg insediments of Clearwater L. collected in June, with higher values

in the surface layer. It varied from8 to 68% inMcFarlane L., withvalues generally increasingwith depth both in June (Fig. 5c) andSeptember (Fig. 6c), and theKOH fraction became thedominantfraction of the sediment in this lake below 15 cm in depth. Thedominance of the Hg-KOH fraction in McFarlane L. sedimentscould be due to the depletion of dissolved free sulfide as theother species controlling Hg distribution in sediments at depthalthough the possible sorption of themetal on amorphous ironsulfide or pyrite cannot be eliminated. The Hg content in theKOH fraction was reasonably (R2: 0.52; pb0.01; N:25) correlatedwith the organic content of the sediment in McFarlane L. butverypoorly correlatedwith thesameparameter inClearwater L.(R2: 0.1; p:0.12; N:25). If the fraction of Hg extracted by KOH isindeed related to humicmatter, the difference between the twolakes might be due to the fact that only a small proportion ofTOC in Clearwater L. sediments is present as humic or fulvicacids (Belzile et al., 1997).

Fig. 7 –Concentration of estimated total organic carbon insediments of both lakes.


Since this extraction protocol was specifically designed forHg species associated to organicmatter in soils and sediments,the corresponding fractions of Fe–KOH (Figs. 4a, 5a and 6a) andMn–KOH (Figs. 4b, 5b and 6b) were very low. However, thesurface enrichment with Fe–Oxal and Mn–Oxal particularlyvisible in June (Figs. 4a, b, 5a and b), not only revealed the redoxstatus of the SWI at that time but confirmed that the Fe andMnoxyhydroxides extracted in the Oxal fraction have littleinfluence on the Hg cycle in freshwater sediments. This issomewhat in contrast with other elements such as As (Belzileand Tessier, 1990), Se (Belzile et al., 2000), Sb (Chen et al., 2003)and other trace metals that can be adsorbed onto Fe and Mnoxyhydroxides and cycled with them in the sediment underreducing conditions. According to Bloom et al. (2003), the KOHfraction should also normally contain methyl Hg, whichshould constitute only aminor fraction of the total. Extractionsof methyl Hg (Bloom, 1989) were performed on the sedimentsof both lakes and the proportionof this compoundwere indeedlower than 1% (b4 pmol/g) in most cases and with no cleartrends as a function of depth (profiles not shown). However, itdoes not necessarily indicate that the production of methyl Hgis low because of its highmobility. The fraction ofmethyl Hg isusually low in sediments but it has been shown by Jin et al.(1999) that the presence of low concentrations of selenium(∼3 nmol/g) in lake sediments can reduce themethylation rateof Hg. We can speculate that relatively high concentrations ofSe in Clearwater L. (more than 10 nmol/g) and McFarlane L.(more than 100 nmol/g) (Belzile et al., 2000) will not favour themethylation of Hg. A preliminary investigation done in ourlaboratory indicates that it could be the case.

Another important fraction of Hg was extracted from thesediments with the cold 12 M HNO3 solution. In Clearwater L.,this fraction varied between 26 and 52% of the total Hg in June(Fig. 4c). Higher values in surficial sediments also suggest thatthe fraction extracted by HNO3 is partially replaced by thatextractable by KOH. This inversion is noticeable in all solidprofilesmore obviously inMcFarlane L. (Figs. 5, 6c and d) and is

much larger than the variation (less than 8%) on extractions.The same fraction varied between 25 and 67% in June (Fig. 5c)and between 17 and 87% in September (Fig. 6c) of total Hg in thesediments ofMcFarlane L., with again higher values in surficialsediments where it also represented the dominant fraction ofHg in the solid. According to theBloomprotocol, this extractioncould include all Hg(0), Hg bound to amorphous organo-sulfur,Hg–Ag amalgams (likely negligible in our lakes) or crystallineFe/Mn oxide phases. However, we do not expect to find a largeamount of Hg bound to this last phase because the Bloomprotocol was slightly modified in our study to replace the twofirst fractions by an extraction with an oxalate buffer thatwould normally dissolve most of the amorphous and poorlycrystalline Fe/Mn oxyhydroxides (Schwertmann, 1964).

It is reported by Bloom et al. (2003) that cinnabar andmeta-cinnabar would only be dissolved by aqua regia. However, wesuspect that some amorphous HgS or Hg ions or complexesadsorbed or co-precipitated on iron sulfidic phases such asamorphous FeS or pyrite (FeS2) could likely be extracted by thecold 12 M HNO3 solution. The strong statistical correlationsexisting between the McFarlane L. fractions Hg–HNO3 and Fe–HNO3 (likely dissolving FeS and FeS2) and between Hg–HNO3

and the fraction of total reducible sulfur (TRS), previouslymeasured at the same location in the sediments of McFarlaneL. (Chen et al., 2003), would support this hypothesis. The im-portant role played by amorphous iron sulfide and pyrite inbinding trace elements (through sorption and/or co-precipita-tion) has been clearly presented in several studies dealingwithtrace elements (Morse and Arakaki, 1993; Huerta-Diaz et al.,1998; Belzile et al., 2000; Chen et al., 2003) and there has beenlittle evidence to date to support the existence of pure tracemetal sulfides in sediments. In a study on the formation anddetermination of elemental sulfur, Chen et al. (1997) haveidentified the first centimetres of the sediments in the sameMcFarlane basin as the site of the most intense productionof acid volatile sulfur that includes amorphous FeS. More re-cently, a sorption study Jeong et al. (2007) demonstrated thatadsorption on synthetic mackinawite (FeS) was mainlyresponsible for the removal of Hg(II) at low molar ratios of[Hg(II)]/[FeS]0.

Values of ion activity products were calculated at depthswhere dissolved Hg2+ and dissolved sulfide could be detected,i.e. between 8 cm above and 8 cm below the SWI (Fig. 3h).Considering the extremely low solubility of mercuric sulfide(HgS(s)=Hg2++S2− log K≈−53), it is not surprising to obtaincalculations indicating high supersaturation with respect tocinnabar or meta-cinnabar in those sediments (results notshown). In order words, the co-existence of measurable con-centrations of dissolved Hg and dissolved sulfide is notexpected if cinnabar is present. However, the formation ofmercury-sulfide complexes such as Hg(HS)+, Hg(SH)2, Hg(HS2)−,or HgS22−, all characterized by relatively high thermodynamicconstants of formation (Table 1), or the formation of Hg poly-sulfides complexes (Jay et al., 2000), or that of dissolved or-ganic matter-Hg-sulfide as recently proposed by Miller et al.(2007), can possibly explain the presence of measurable con-centrations of dissolved Hg in the sediments of McFarlane L.in June and September. If mercuric sulfide compounds such ascinnabar do not directly control the solubility of Hg underreducing conditions, it could be likely done by sorption on Fe

Table 1 – Thermodynamic constants of mercury andsulfur compounds (T=25 °C; I=0)

Reaction log K Reference

HgS(cinnabar)+H+=Hg2++HS− −39.1 NIST (2003)Hg2++HS−=Hg(HS)+ 22.29a Benoit et al. (1999)Hg2++2HS−=Hg(HS)2 40.39a Benoit et al. (1999)Hg2++HS−=HgS0+H+ 29.8a Benoit et al. (1999)Hg2++2HS−=HgS22−+2H+ 25.51a Benoit et al. (1999)Hg2++2HS−=HgS(HS)−+H+ 34.6a Benoit et al. (1999)

a Recalculated by Zhang et al. (2004) from an initial value of I=0.3 Musing the Davis equation.


sulfide compounds, especially in McFarlane L. sedimentswhere TRS are more abundant than in Clearwater L. (Belzileet al., 2000). Amorphous Fe sulfide and pyrite are well knownfor the important role they play in controlling the solubilityof other trace metals (Huerta-Diaz et al., 1998; Billon et al.,2001) andmetalloids such as As (Moore et al., 1988; Belzile andLebel, 1988), Sb (Chen et al., 2003) and Se (Belzile et al., 2000).

Profiles of Hg extracted by aqua regia, Hg(AR) of both lakesshow values that represent less than 5% of the total. This isanother confirmation of the negligible presence of cinnabarin our samples, which should be extracted by this last fractionof the Bloom protocol. This is somewhat in contrast what hasbeen reported for soil contaminated with Hg (Revis et al.,1989). Besides the strong influence of organic matter such ashumic and fulvic acids and that of sulfides on the distributionof Hg in sediments, other mineral phases such as Fe and Mnoxyhydroxides do not seem to have much influence on thedistribution of Hg in the studied sediments, even in Clear-water L. where a distinct oxidized layer of Fe and Mnoxyhydroxides exist in permanence (Fig. 4a and b; Oxalfractions). In McFarlane L, the seasonal onset of reducingconditions at the SWI does not favour the formation andenrichment of the surficial sediment with those oxyhydr-oxides. Such mineral phases are known to play a significantrole on the cycling of several other metals and metalloids(Tessier et al., 1996). This limited influence of amorphous Feand Mn oxyhydroxides on the mobility of Hg in sedimentsand soils has been previously reported (Gambrell et al., 1980;Wallschläger et al., 1998). Although the results of thesequential extraction clearly indicate the strong connectionof Hg with the KOH and HNO3 fractions, it is important toremember that the attribution of Hg to specific biogeochem-ical categories works best for highly contaminated samples,as stated by Bloom et al. (2003). However, we are convinced ofthe validity of our results and we believe that those twofractions play an important competitive role on the cycling ofHg in freshwater sediment as confirmed by several otherstudies already cited. It is important to reiterate that theaddition of all extracted fractions and residual led to valuesvarying between 92 and 98% of the total for each sample.

3.4. Comparing the two lakes

Considering that the two study lakes have approximately thesame sedimentation rate of around 1 mm per year (Nriaguet al., 1982; Carignan and Nriagu, 1985) and assuming an

equal contribution from atmospheric deposition due to theproximity of the two lakes, the pool of Hg should be similar inboth lakes, for the studied depth of sediment. However, itlooks like the first 25 cm of surface sediment of McFarlane L.containmoreHg than those of Clearwater L. Itmight be due toa more intense cycling of the element in McFarlane L. whereoxic and anoxic conditions alternate at the sediment–waterinterface or to the physical positioning of this lake being fedby two upper lakes whereas Clearwater L. is a head lakesurrounded by a limited watershed. It is also possible that theslightly more alkaline conditions prevailing in McFarlane L.water and sediments favour a better retention of sulfidespecies and the complexation of Hg to those species inporewaters and in sediments. Only profiles of dissolvedconstituents and solid fractions measured in June werepresented for Clearwater L. because those obtained in Augustunder similar oxic conditions at the SWI were very similar tothe June profiles. The situation in McFarlane L. was differentbecause of the gradual depletion of dissolved oxygen in thehypolimnion of the studied basin and the onset of stronglyreducing conditions at its SWI. It is clear that such rapidchanges in the redox status of the SWI make the McFarlaneenvironment as very dynamic in terms of dissolved sulfideand DOC concentrations (Fig. 3). The consequences on thedissolvedHg profile are not obvious but further investigationson the exact nature of the dissolved Hg species couldelucidate those changes.

4. Summary

Profiles of dissolved species and distributions in the solidphase show that the early diagenesis of Hg in freshwater sedi-ments is affected by several factors. The influence of thermo-dynamics (pH and redox status) has been evidenced by thecontrasting redox conditions existing at the SWI of the twolakes close to the end of summer. The mobility of Hg in sedi-ments strongly depends on the complexation with naturalorganic matter (Hg–KOH), the formation of elemental Hg,Hg organo-sulfides or Hg adsorbed on amorphous sulfides(Hg–HNO3). They represent the two major fractions of Hg ex-tracted from the solid sediment and competing to control thesolubility and mobility of the metal in the two study lakes.Overall, it might mean that the organic matter plays an evenlarger role in themobility of Hg if we accept the possibility thatorgano-sulfide compounds are significantly represented inthe Hg–HNO3 fraction. The presence of iron and manganeseoxyhydroxides in the oxic layer of the sediment or cinnabar inthe anoxic one does not seem to play a major role on thecycling of Hg. We are now investigating on the role that couldbe played by Se on the methylation of Hg in sediments.

Acknowledgements

This work received financial support from the NaturalSciences and Engineering Research Council of Canada throughthe COMERN research network. Technical assistance from JianTong and diving by John Varney and Rick Carrey are sincerelyacknowledged. The manuscript highly benefited from thejudicious comments of two anonymous reviewers.


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the competitive role of organic carbon and dissolved sulfide in controlling the distribution of...

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