environmental chemistry and toxicology of mercury (liu/toxicology of mercury) || chemical speciation...

CHAPTER 7

CHEMICAL SPECIATION OF MERCURYIN SOIL AND SEDIMENT

ULF SKYLLBERG

7.1 INTRODUCTION

Reactions of mercury and its chemical speciation in soils and sediments havebeen covered in several reviews over the years (Andersson, 1979; Benes andHavlik, 1979; Schuster, 1991). The latest comprehensive review of the chemicalspeciation, linked to the bioaccumulation of mercury, in aquatic and terrestrialenvironments was published in 1998 (Morel et al., 1998). Since then, experimen-tal and field-oriented studies have contributed significant new findings. Especially,the role of natural organic matter (NOM) has been updated. Spectroscopic andbinding affinity studies conducted at environmentally relevant conditions haveprovided new information on the stability and structure of inorganic mercury(Hg) and methyl mercury (MeHg) complexes with thiol groups in NOM. Stud-ies have pointed out the stabilizing effect of NOM on nanoparticulate HgS(s).This has not only led to an improved understanding of the mobility of mercurybut has also questioned methods used to separate aqueous and solid phases, aswell as stability constants relying on those methods. Studies on chemical reac-tions between Hg and iron sulfides have provided quantitative data that could beincorporated in chemical equilibrium models. Despite these significant advancesthere is still a long way to go before the chemical speciation and solubility ofmercury is fully understood. It is noteworthy that there is still not a single studyin which the absolute concentration of Hg in the pore water of a soil or sedimenthas been satisfactorily explained by thermodynamic models, covering molecularspecies pertaining to aqueous and solid phases. A remaining important challengeis to determine the aqueous phase chemical speciation of Hg in equilibrium withsolid phases of HgS(s) at total concentrations of sulfides less than 10 μM. Alsothe issues of polysulfide formation and the interaction between Hg and MeHg

Environmental Chemistry and Toxicology of Mercury, First Edition.Edited by Guangliang Liu, Yong Cai, and Nelson O’Driscoll.© 2012 John Wiley & Sons, Inc. Published 2012 by John Wiley & Sons, Inc.

219

220 CHEMICAL SPECIATION OF MERCURY IN SOIL AND SEDIMENT

and various types of metal sulfides need to be resolved before thermodynamicmodels are good enough to be used for explanatory and predictive purposes insoils and sediments.

7.2 PHYSICOCHEMICAL PROPERTIES, OXIDATION STATES,CHEMICAL FORMS, STRUCTURES, AND CONCENTRATIONSOF MERCURY IN THE ENVIRONMENT

The Greek name “hydrarguros” means liquid silver, thereof the name “quick-silver.” The unique physicochemical properties, as reflected by its low melting(−38.87◦C) and boiling (356.57◦C) points, have facilitated the extraction and useof mercury by man since ancient times. Its electronic configuration, with filled fand d orbitals ([Xe]4f145d106s2), makes elemental Hg somewhat similar in sta-bility to noble gas elements. Because of weak interatomic attraction, mercury isthe only liquid metal at room temperature, and its vapor is monoatomic. Ele-mental mercury Hg0 occurs in liquid, gaseous, and aqueous phases (Table 7.1).It is quite soluble in water (60 μg/L at 25◦C) and is oxidized to the metastabledimer Hg2

2+, Hg(I), and to more stable Hg(II) forms. In natural environments,the mercurous ion Hg2

2+ may be an intermediate in the oxidation of Hg0 to themercuric ion Hg2+, but it is not long-lived enough to be detected by conventionalmethods. Release of Hg0(g) into the atmosphere by combustion of fossil fuelsand natural sources such as evasion from the sea (Mason and Sheu, 2002) andvolcanic eruptions, are the major sources of mercury in the environment. In theatmosphere, 95% of the total mercury is in the form of Hg0(g), and it is eventuallyoxidized and deposited as Hg(II) on land and in water as particles or dissolvedin rain water. Mercury(II) may also be present in organic forms (often referred toas “organic mercury” or “alkylated mercury”) such as monomethyl and dimethylmercury molecules. Monomethyl mercury (commonly referred to as methyl mer-cury, CH3Hg+, MMHg, or MeHg) is by far the most abundant organic form ofmercury in the environment. During the methylation process, inorganic Hg(II) istransformed to the MeHg molecule by the action of mainly iron (III) reducingbacteria (FeRB) and sulfate reducing bacteria (SRB) (see Chapter 5). Dimethylmercury (DMHg) is encountered in deep sea water, beneath the thermocline, but

TABLE 7.1 Oxidation States and Chemical Forms of Mercury Occurringin the Environment

Chemical Form Oxidation State Environment

Hg(g, l, aq) 0 Aquatic–terrestrialHg2+

2 +I MetastableHg2+ +II Aquatic–terrestrialCH3Hg+ +II Aquatic–terrestrial(CH3)2Hg(aq) +II Marine

PHYSICOCHEMICAL PROPERTIES, OXIDATION STATES, CHEMICAL FORMS 221

it has not been detected in the mixed layer where gas exchange with the atmo-sphere takes place (Fitzgerald et al., 2007). Phenylmercury acetate has been usedextensively as a pesticide in agriculture and pulp and paper industry but is notstable in the environment and degrades to inorganic Hg(II).

The biogeochemistry of mercury in soils and sediments is dominated by inor-ganic and organic mercury(II) complexes. Redox processes involving Hg(II) andHg0 and the partitioning of Hg0 between liquid, aqueous, and gaseous phases arealso of great importance. These processes are covered in Chapter 6. Because ofits large size and stable electronic configuration ([Xe]4f145d10), Hg(II) is easilypolarized by interacting atoms. The strong covalent character of the C–Hg bondis the reason for the high stability of the MeHg molecule. Similarly, the Hg2+ion forms very strong covalent bonds with soft Lewis bases (ligands) such ashalides and sulfur compounds, but relatively weak bonds with hard Lewis basessuch as fluorine (Fig. 7.1). These exceptional properties make Hg the “softest”of all metals (a so-called “Type B metal”). Typical for Hg(II) complexes is alow coordination number (2, 3, or 4). The linear two-coordination of Hg(II) iscaused by relativistic effects (Tossell and Vaughan, 1981) and is a very unusualstructure for a metal.

14

12

10

8

6

Al3+

Cd2+

Hg2+

Zn2+

4

2

0F Cl

Increasing softness

log

KM

L

Br I

Figure 7.1 Formation constants for metal–ligand complexes illustrate the “hard–soft–acid–base” (HSAB) principle. Mercury(II) is the softest of all metal ions (Lewisacids = electron acceptor), cadmium(II) is slightly less soft, zinc is on the borderlinebetween soft and hard acids, and aluminum is one of the hardest Lewis acids. Fluorineis a hard Lewis base (electron donor), and chlorine, bromine, and iodine are increasinglysoft Lewis bases. Source: Figure modified from Sposito (1994).


The total concentration of Hg(II) in the aqueous phase of soils and sedimentsis normally in the range 5–100 pM, and solid phase concentrations are often inthe range 50–500 ng/g (250–2500 pmol/g). In contaminated sediments, concen-trations may be 50 times higher in the aqueous phase and up to 1000 times higherin the solid phase. Concentrations of MeHg contribute in general to 0.1–10% ofthe total Hg (sum of Hg and MeHg) in solid phases and 5–80% of total Hg inaqueous phases of soil and sediment.

7.3 AQUEOUS PHASE: MAJOR LIGANDS AND THEIR AFFINITIESFOR MERCURY(II)

Below I will briefly cover the Hg and MeHg aqueous species recognized asquantitatively important and that can be considered relevant for thermodynamicmodeling in soil and sediment. In Appendix I, stability constants are listed forspecies for which the reliability of the value of the constant is considered fair.As will be explained, some stability constants are under debate, and for those Ihave taken a conservative standpoint and excluded them from Appendix I.

7.3.1 Oxygen-Containing Inorganic Ligands and Halides

The hydrated Hg(H2O)62+ (octahedral six-coordination) ion is an exception to the

rule of low coordinated Hg(II) complexes. This complex will undergo hydrolysiswith Hg(OH)2

0 as the dominant protolyte at neutral pH. As expected for soft“Type B metals”, the stability of Hg(II)–halide complexes increases with theatomic number of the halide (Fig. 7.1). Because of the abundance of chlorideions ([Cl−] ranging between 10−5 M in some freshwaters and 0.6 M in marineenvironments), Hg–chloride complexes are always in great excess over bromineand iodine complexes. For MeHg, the strong C–Hg bond strongly favors theformation of linear two-coordinated structures, and therefore, only the neutralforms MeHgOH0 and MeHgCl0 are quantitatively important.

7.3.2 Reduced Sulfur Ligands

7.3.2.1 Organic Thiols and Organic Sulfides. Recent studies show thatlow molecular mass (LMM) thiols (e.g., mercaptoacetic and mercaptopropionicacids, cysteine, and glutathione) are produced and build up under suboxic andanoxic conditions in hypolimnetic bottom waters. Total concentrations on theorder of 10–100 nM (and sometimes even in the micromolar range) can beexpected to be normal (Zhang et al., 2004; Hu et al., 2006). Thiol groups (RSH)associated with dissolved organic matter (DOM) are normally in the micromolarrange and are generally more resistant to oxidation than LMM thiols. Both Hgand MeHg form very stable complexes with organic thiols. Spectroscopic stud-ies and thermodynamic modeling conclusively show that the two-coordinated

AQUEOUS PHASE: MAJOR LIGANDS AND THEIR AFFINITIES FOR MERCURY(II) 223

Hg(SR)2 complex is highly dominant over the one-coordinated HgSR+ complex.At neutral and alkaline pH, with great excess of thiols in relation to Hg(II) andin absence of steric hindrance, three- and four-coordinated complexes can form(Koszegi-Szalai and Paal, 1999; Jalilehvand et al., 2006; Leung et al., 2007).The stability constant for these complexes are, however, not well established.The steric hindrance likely makes three- and four-coordinated Hg less abun-dant when complexed with thiol molecules associated with DOM or particulateNOM, as indicated by spectroscopic studies (Skyllberg et al., 2006). Studies onthe complexation of Hg and MeHg with organic thiols were recently summa-rized (Skyllberg, 2008). In Appendix I, stability constants for the formation ofthe Hg(SR)2 and MeHgSR molecules are listed. These constants should be seenas representative for complexes with both LMM and DOM-associated thiols.It should be noted that methodological difficulties have resulted in a number ofincorrectly determined stability constants for LMM Hg–thiol complexes reportedin the literature, as described in Casas and Jones (1980). Stability constants forthe complexation of Hg and MeHg with organic mono- and bisulfides such asmethionine (RSR) and cystine (RSSR) are much smaller (Latha et al., 2007), andthese complexes are not quantitatively important in soils and sediments.

In Fig. 7.2a–d, the outcome of thermodynamic modeling of Hg and MeHg spe-ciation under oxic conditions is illustrated. For comparative reasons, no organicligands are present in Fig. 7.2a, which is the traditional way to illustrate the aque-ous phase speciation of Hg. In the presence of LMM thiols, Hg(SR)2 complexeswill dominate even at quite modest concentrations of thiols (10 nM in Fig. 7.2b).In the presence of DOM, complexes with thiols [Hg(SR-DOM)2 and MeHgSR-DOM] will dominate the speciation of Hg and MeHg (Fig. 7.2c and 7.2d). Evenat concentrations of dissolved organic carbon (DOC) as low as 1 mg/L (corre-sponding to 46 nM of thiol groups if DOC is assumed to contain on average0.15% thiol groups on a mass basis, Skyllberg (2008) Hg(SR-DOM)2 complexesdominate in marine environments with chloride concentrations of 0.6 M.

7.3.2.2 Inorganic Sulfides, Bisulfides, and Mixture of Hydroxides andSulfides. The current theory for the formation of aqueous phase Hg–sulfidespecies follows the classical study of Schwarzenbach and Widmer (1963). Theymeasured the solubility of Hg2+ in equilibrium with “black HgS” (metacinnabar),taking advantage of the reaction Hg2+ + Hg0(l) = Hg2+

2 (log K = 2.0) occurringat the surface of an elemental mercury electrode. The total solubility of Hgwas determined as the radioactivity of 203Hg (after a three-day procedure dur-ing which the aqueous phase was slowly forced by N2 gas through an 8-μmfilter). Contamination by HgS particles passing the filter was recognized by theauthors as a major problem in the experiment, and precautions were taken to getreproducible results. The solubility of HgS(s) in the presence of dissolved HS−and H2S at pH 0–10.7 (in 1.0 M KCl) was explained by the formation of themononuclear complex Hg(SH)2

0 and its two protolytes HgS2H− and HgS22−.


−73 4 5 6

pH

(a) (b)

(c) (d)

MeHgCl

MeHgSR-DOM

Hg (OH)2

Hg(SR)2

HgCl2

HgCl3–

HgCl42– HgCl4

2–

log

[Cl– ]

(M

)

7 8 9 10

−6

−5

−4

−3

−2

−1

0

3 4 5 6

pH

log

[Cl– ]

(M

)

log

[RS

tot]

(M)

7 8 9 10−6

−6 –5 −4 −3log [Cl] (M)

MeHgOH MeHgCl

MeHgSR-DOM

−2 −1 0−14

−12

−10

−8

−6

−4

−2

−5

−4

−3

−2

−1

0

−73 4 5 6

pH

log

[Cl– ]

(M

)

7 8 9 10

−6

−5

−4

−3

−2

−1

0

HgClOH

Figure 7.2 Dominance diagram for aqueous phase species of Hg under oxic conditionsin the absence (a) and in the presence (b) of 10 nM of LMM thiols (RStot), and forMeHg under oxidized conditions in the presence of DOC 50 mg/L (having 2.3 μM ofthiol functional groups, DOM-RStot) (c) and a pH of 7.0 (d). Marine conditions areindicated by black areas, and freshwater conditions by gray areas. Stability constants usedin the calculations were taken from Appendix I, without correction for ionic strength andtemperature effects.

Barnes et al. (1967) argued that in the acidic region HgS(s) should react primar-ily with H2S(aq) (the dominant form of the H2S–HS− –S2− system at pH below∼7), suggesting that HgS(H2S)2

0(aq) rather than Hg(SH)20(aq) (the difference

being one H2S molecule) was the likely form of the major dissolved Hg-sulfidemolecule. After deprotonation, Hg(HS)3

− and HgS(HS)22− molecules were sug-

gested to be the dominating molecules at neutral pH and HgS22− at alkaline pH.

Recent extended x-ray absorption fine structure (EXAFS) spectroscopic studies(Lennie et al., 2003; Bell et al., 2007) have shown that Hg is two-coordinatedat pH 9–11.5, which is in agreement with both studies. So far, no spectroscopicstudies have been conducted at lower pH.


Stability constants for mixed chloride-bisulfide (HgClSH0(aq)) and hydroxide-bisulfide (HgOHSH0(aq)) listed in Appendix I are taken from Dyrssen andWedborg (1991). It should be noted that these constants are not experimentallydetermined but estimated by linear free energy relationshipsLinear Free EnergyRelationships (LFERs). The reported log K for the reaction HgS(s) + H2O =HgOHSH0(aq) is −22.3. A complication is that Dyrssen and Wedborg (1991), inan ambition to evaluate their theoretically determined constants by experimen-tal data, also reported a log K of −10.0 for the above reaction. The differenceof 12.3 log units between the two constants is due to a correction made fromliterature data from solubility experiments on CdS(s) and ZnS(s). Dyrssen andWedborg made this correction, despite noting that the solubility data of CdSand ZnS were biased by colloidal contamination of the aqueous phase, whichlater was verified (Daskalakis and Helz, 1993). The reported log K of −10.0for the reaction HgS(s) + H2O = HgOHSH0(aq) was later used to explain thepartitioning of Hg(II) in equilibrium with HgS(s) between octanol and water(Benoit et al., 1999b; Jay et al., 2000). The concentration of Hg in the octanolphase was calculated by difference after determination of Hg in the aqueousphase. Particulates of HgS(s) were separated from dissolved forms of Hg by0.2- (Benoit et al., 1999b) and 0.02-μm filters (Jay et al., 2000). Because col-loidal HgS has been shown to readily pass 0.1-μm filters (see below), it cannotbe ruled out that KOW distribution constants derived for HgOHSH0(aq) werebiased by colloidal contamination. This was shown to be the case in a similarexperiment in which 96% of HgS(s) in nanoparticulate form was partitioned intothe octanol phase when mixed with water at pH 7.0 (Deonarine and Hsu-Kim,2009). Because of the uncertainties associated with experimental data, a con-servative approach is taken in this chapter, and the theoretically derived logK of −22.3 for the reaction HgS(s) + H2O = HgOHSH0(aq) is used in thethermodynamic modeling (corresponding to a log K of 30.3 for the reactionHS− + OH− + Hg2+ = HgOHHS0(aq), Appendix I).

7.3.2.3 Polysulfides. Polysulfides are neutral or, when deprotonated, nega-tively charged ions that form chains consisting of 2–8 sulfur atoms. Often theformation of polysulfides is expressed as a reaction between elemental S (rhombicS8

0) and bisulfide: (n − 1)/8S80 + HS− = Sn

2− + H+, Kamyshny et al. (2004)but polysulfides exist also in the absence of S8

0 (albeit at a much lower concen-tration), and the formation can then be expressed as a reaction between bisulfideand sulfate: SO4

2− + 7HS− + H+ = 4S22− + 4H2O (Rickard and Luther, 2007).

Polysulfides are considered to be metastable in the environment, as they arecommonly formed during redox fluctuations when oxidized and reduced formsof sulfur get into contact. If the more stable forms of mono- and disulfides aredisregarded, polysulfides with n> 4 sulfur atoms dominate in the pH–Eh spacewhere rhombic S8

0 is the stable form (Fig. 7.3a and 7.3b). In fact, as illustratedin Fig. 7.3c, polysulfides with 3–8 S atoms show concentrations roughly similarto or higher than HS− in the presence of rhombic S8

0 at pH values above 8. Inthe absence of S8

0, polysulfide concentrations are much lower (Fig. 7.3d).


–0.52 4 6

pH

0

Total sulfur

2H2S

HS–

HS2

4

6 7 8

pH

–log

C

9 10

–25

–0.5

log

{S s

peci

es}

–0.4 –0.3 –0.2

Eh (V)

–0.1 0 +0.1

–24

–20

–16

–12

–8

–4

0PH2

> 1 b H2O H2S

H2S2

HS2

HS–SO–

pH = 6

{S}T = 10–3

6

8 10

–0.4

–0.3

–0.2

–0.1

Eh

(V)

0

0.1 Sulfur

H2S

HS–

SO2–

0.2

4

–0.52 4 6

pH(a) (b)

(c) (d)

8 10

–0.4

–0.3

–0.2

–0.1

Eh

(V)

0

0.1H2S8

H2S2HS6

HS5

HS2

HS8

0.2

2–

S82–

S22–

S62–

SO42–

–

–

HS3–

HS4–

HS5H2S2 HS6

– –

S22–

S82–

S32–

S72–

S62–

S52–

–

S2–

4

–

–

Figure 7.3 (a) Stable sulfur species in aqueous solution at 1 mM total concentration((25)

◦C). (b) Stable phases removed, emphasizing the regions of dominance if only themetastable polysulfide species are considered. (c) Mono-, di-, and polysulfide distribu-tion in the presence of excess S8

0. (d) Mono-, di-, and polysulfide distribution in theabsence of S8

0. Thermodynamic data for polysulfide formation are taken from Kamyshnyet al. (2004). Source: Reprinted from Rickard and Luther (2007) with permission fromAmerican Chemical Society.

Even if Schwarzenbach recognized the importance of polysulfide formation(Schwarzenbach and Fischer, 1960) before his classic work on HgS(s) solubility,it was not until the work of Paquette and Helz (1997) that the significantcontribution from Hg–polysulfides was recognized. Building on the study bySchwarzenbach and Widmer (1963), Paquette and Helz explained the highersolubility of HgS(s) in the presence of S8

0 by the formation of a polysulfidewith the general form HgSnSH−. This species can be considered an analogto the first protolyte of the disulfide species (HgS2H−) and consequentlydominates the speciation in the mid pH range (5.5–9.5). The number of S atoms


in the HgSnSH− species could not be specified, but according to Kamyshnyet al. (2004), n = 4–8 is a possible range. By 199Hg NMR spectroscopy,synthesis of HgS4S2−, HgS4S5

2−, and Hg(S5)22− has been verified in laboratory

experiments (Bailey et al., 1991). Theoretical calculations performed by Tossell(1999) suggest that the dissolution of cinnabar through the formation of apolysulfide with n = 4 (HgS4SH−) is an energetically favorable process. Sofar no spectroscopic evidence exists for aqueous phase Hg–polysulfides inequilibrium with HgS(s), despite studies at the high pH range of 9–11.5, wherepolysulfides should dominate (Lennie et al., 2003; Bell et al., 2007).

Jay et al. (2000) extended the experimental work to lower sulfide concen-trations, and at high pH two additional polysulfide species, Hg(Sn)

2−2 and

HgSnOH−, were suggested to explain their and Paquette and Helz’s solubilitydata (Fig. 7.4a and 7.4b). The first species is an analog to the second protolyteof the disulfide species (HgS2

2−), whereas the second species may be seen as apolysulfidic analog to HgSHOH0. Jay and coworkers showed that the stabilityconstant for the formation of Hg(Sn)2

2− improved the model substantially atpH values up to 8.5. The stability constant suggested for the HgSnOH− species(HgS(s) + HS− + (n − 1)S8

0 + H2O = HgSnOH− + H+; log K = −15.7(I =0.3M) was less well-constrained, and the data support was restricted to pHvalues above 9.0 and sulfide <100 μM. Similar to the uncertainty regardingthe stability assigned to HgSHOH0, it could be questioned whether a mixedpolysulfide/hydroxide complex would have the high stability suggested by Jayand coworkers. Even if a 0.02-μm filter was used to separate the aqueous andsolid phases in the study by Jay et al. (2000), the large increase in the formationof colloidal HgS observed above pH 8 (Ravichandran et al., 1999) may callinto question HgS(s) solubility data obtained above pH 9 owing to a possiblecolloidal contamination of filtrates by HgS(s) (see Section 7.4.2.1).

The importance of polysulfide formation for HgS(s) solubility at pH valuesabove 4.0 is illustrated at two levels of aqueous phase sulfide concentrations inFig. 7.4c. In this figure, the species HgOHSH0 and HgSnOH− are not included.The great influence on the HgS(s) solubility by inclusion of a log K of 20.8 for thereaction HS− + (n − 1)S8

0 + H2O + Hg2+ = HgSnOH− + 2H+ and a log K of−10.0 for the reaction HgS(s) + H2O = HgOHSH0(aq) is illustrated in Fig. 7.4d.At 1 μM S(-II), the HOHgSH0 molecule dominates the solubility below pH 8 andHgSnOH− dominates at alkaline conditions. As can be noted in Fig. 7.4a and 7.4b,data collected at S(-II) <30 μM (∼10−4.5 M), which could give support for asignificant contribution from HgOHSH0(aq) and HgSnOH−, are currently lacking.Solubility experiments conducted at sulfide concentrations lower than 30 μMunder the prevention of colloidal contamination and independent spectroscopicstudies are needed before stability constants for these two species can be fullyestablished. I have therefore taken a conservative approach and excluded theHgSnOH− molecule and its proposed stability constant in Appendix I.

If all relevant components are considered under anoxic conditions, there willbe a competition among organic thiols, inorganic sulfides, and polysulfides for Hgand MeHg in the aqueous phase, as illustrated for a soil/sediment having 50-mg


–14

log

Hg-

spec

ies

(M)

–6.0 –5.0 –4.0

log S(–II)Tot (M)

–3.0 –2.0

–12

–10

–8 Total dissolved Hg

HgS2

HgS2H–

HOHgSH0

Total dissolved Hg

Hg(SH)2

Hg(SH)20

2–

–6

–4

–14

log

diss

olve

d H

g(II)

(M

)

4.0 5.0 6.0 7.0

pH

[S(–II) = 1 μM]

[S(–II) = 1 mM]

including

polysulfides

HgSnSH–

Hg(Sn)2–

I = 1 mM [S(–II) = 1 μM]I = 1 mM

D

C

B

A

8.0 9.0 10.0

–12

–10

–8

–6

–4

pH = 8.0I = 0.3 M

(a) (b)

(c) (d)

pH = 8.0I = 0.3 M

–14

log

Hg-

spec

ies

(M)

–6.0 –5.0 –4.0

log S(–II)Tot (M)

–3.0 –2.0

–12

–10

–8

–6

–4

0

0HgSH2

HgSnOH–

Hg(Sn)2

2–HgS2H

HOHgSH0

HgS5

–

2–

2

log

diss

olve

d H

g(II)

(M

)

4.0 5.0 6.0 7.0

pH

8.0 9.0 10.0–14

–12

–10

–8

–6

Figure 7.4 Solubility of HgS(s) in the absence (a) and presence (b) of elemental S (activ-ity set to unity) as a function of total dissolved sulfide concentrations [S(-II)] at pH 8.0.The open circles and solid diamonds represent data points from Paquette and Helz (1997)and Jay et al. (2000), respectively, and solid lines illustrate constants reported in ref Jayet al. (2000). The solubility as a function of pH is illustrated with (broken lines) and with-out (solid lines) consideration of the polysulfides HgSnSH− and Hg(Sn)2

2− in (c). In (d),the sum of concentrations of Hg(SH)2

0 + HgS2H− + HgS22− is represented by line A,

with successive addition the polysulfides Hg(Sn)SH− and Hg(Sn)22− (line B), the poly-

sulfide HgSnOH− (line C), and the species HgOHSH0 (line D). Stability constants usedin the calculations were taken from Jay et al. (2000), after correction for ionic strengtheffects.

DOC per liter in the pore water (Fig. 7.5). At 2.3 μM of RStot, concentrations ofHg(SR-DOM)2 and Hg–sulfides will be roughly equal at 0.5 μM of S(−II) =[H2S] + [HS−], and if polysulfides form, less than 0.1 μM of S(-II) is requiredto outcompete Hg(SR-DOM)2 at pH values above 6.5. At pH less than 5, theHg(SH)2

0 molecule will be the dominant inorganic complex even if polysulfidesare present. The relative competitiveness between inorganic and organic sulfur

LIQUID AND SOLID PHASES OF MERCURY IN SOILS AND SEDIMENTS 229

–83 4 5 6

pH

7 8 9

Hg(SR-DOM)2

Hg(SH)2

10

–7

log

[S-I

I] (M

)

–6

–5

–4

–3

0

2–

2–

HgS2H– HgS2

–83 4 5 6

pH

(a) (b)

7 8 9 10

–7

log

[S-I

I] (M

)

–6

–5

MeHgSH0

MeHgS–

MeHgSR-DOM

–4

–3

Hg(Sn)2HgSnSH–

Figure 7.5 Dominance diagrams for aqueous phase species of Hg (a) and MeHg (b)as a function of pH and concentrations of S(−II)(= [H2S] + [HS−]) in the presence ofDOC 50 mg/L (having 2.3 μM of thiol functional groups; DOM-RStot). In (a), dottedlines illustrate a situation without polysulfides, and solid lines with polysulfides (and theactivity of elemental S set to unity). Stability constants used in the calculations were takenfrom Appendix I, without correction for ionic strength and temperature effects.

ligands for MeHg (Fig. 7.5b) is quite similar to their competitiveness for Hg,with a maximum contribution from thiol complexation when pH equals the pKa

of MeHgSH0 (7.5).

7.4 LIQUID AND SOLID PHASES OF MERCURY IN SOILSAND SEDIMENTS

In geological deposits and in mine tailings after processing of ore, mercury isknown to occur in many different mineral forms, in association with oxygen, sul-fur, and chloride. The principal minerals are cinnabar (α-HgS) and metacinnabar(β-HgS), but minerals such as montroydite (HgO), mercuric chloride (HgCl2), ter-linguite (Hg2OCl), eglestonite (Hg3Cl3O2H), corderoite (Hg3S2Cl2), and schuet-teite (Hg3O2SO4) are also quite abundant in mine wastes (Kim et al., 2000).At sites without direct deposition of mercury from industrial or mining pro-cesses, and in presence of NOM, sulfides are the only stable solid phases ofmercury(II).

7.4.1 Elemental Mercury

Elemental mercury in gaseous and aqueous forms are well known to be ubiquitousin the environment, and the reduction of Hg(II) to Hg0 and subsequent evasioninto the atmosphere is an important source to the global mercury cycle (Masonand Sheu, 2002). The liquid form, Hg0(l), however, is practically unknown except


in soils and sediments heavily contaminated by local industry such as chlor-alkaliplants from which Hg0(l) has been released into the environment. By thermod-esorption analysis it has been shown that only traces of Hg0(l) persist in thesehighly contaminated soils and sediments (Biester et al., 2002). In marine andbrackish water environments, the major degradation of Hg0(l) is rapid oxidationby chlorine ions (Yamamoto, 1996). Interestingly enough, the oxidation of Hg0(l)stops immediately when the liquid phase [droplets of Hg0(l)] is removed, even ifthe concentration of Hg0(aq) is close to saturation with the solid phase (Amyotet al., 2005). Obviously the oxidation takes place at the interface between Hg0(l)droplets and water. The Hg0(aq) form is resistant to oxidation by chlorine but israpidly transformed to Hg(II) by photolytic reactions (Lalonde et al., 2004). Fora detailed discussion on photochemical reactions of mercury, I refer to Chapter6.

7.4.2 Mercury(II) Sulfides

Mercury(II) sulfides exist as two polymorphs; cubic metacinnabar (β-HgS), whichis considered metastable under low temperature conditions, and the more stablehexagonal cinnabar (α-HgS). Likely because of natural redox fluctuations, β-HgS(s) seems to be the predominant form in anoxic soils and sediments (Barnettet al., 1997). It has been shown by X-ray diffraction (XRD) and X-ray absorp-tion spectroscopy (XAS) that Hg2+ in the initial reaction with HS− forms anunstable low coordinated complex, with Hg–S bond distances similar to that incinnabar. This structure is explained by a gradual clustering of the predominanttwo-coordinated complexes Hg(SH)2

0, HgHS2−, and HgS2

2− in solution. Alreadyafter 5 s a cluster is formed with a local four-coordinated structure similar to β-HgS. After hours of further aging, it is graded into a more crystalline β-HgSstructure, which was not changed after 12 months of aging at ∼20◦C (Charnocket al., 2003). The structure of β-HgS is also consistent with an indicated fourfoldcoordination of Hg with polysulfides (Bailey et al., 1991) at suboxic conditions.

7.4.2.1 Colloidal Contamination of Filtrates in Studies on Mercury(II)Sulfide Solubility. Studies on HgS(s) solubility encounter the difficulty of sep-arating truly dissolved Hg species from colloidal forms of HgS clusters andHgS nanoparticles. Ravichandran et al. (1999) showed that the fraction of col-loidal HgS passing a 0.1-μm filter increased from less than 1% to 90% whenpH was increased from 8.0 to 10.0. In this experiment, HgS(s) was precipitatedby mixing solutions of Na2S and HgCl2. The same problem was recognizedwhen starting with a solid black metacinnabar phase and dissolving it in sulfidicsolution (Schwarzenbach and Widmer, 1963). This means that the solubility ofHgS(s) will be overestimated, especially at pH values above 8. Whether the fine0.02-μm filters used in recent studies on HgS(s) solubility (Paquette and Helz,1997; Jay et al., 2000) are fine enough to completely separate the finest colloidsproduced at alkaline conditions from the true aqueous phase we simply do notknow.

REACTIONS OF MERCURY(II) WITH SOIL AND SEDIMENT PARTICLE SURFACES 231

7.4.2.2 Solubility Product of Hg(II) Sulfide. The solubility product of mer-cury(II) sulfides has been determined in a few studies. Schwarzenbach andWidmer (1963) determined the log solubility product of a black HgS (the structurewas not determined, but it is generally assumed to be equivalent to metacinnabar)to be −36.8 (HgS(s) + H+ = HS− + Hg2+) in a 1.0 M KCl ionic medium. Forred HgS(s), cinnabar, a solubility product of −36.5 was determined at 0.7 Mionic strength (Paquette and Helz, 1997), Thus, despite the different structures ofα- and β-polymorphs, their solubilities do not seem to differ significantly. LFERssuggests that the solubility product of metacinnabar would be 0.4 log units lower(more soluble) than cinnabar (Potter and Barnes, 1978).

As can be seen in Fig. 7.4a and b, solubility data collected for HgS(s) havebeen restricted by the detection limit of methods used to determine total Hg in theaqueous phase and no data points exist at total S(-II) concentrations lower thanapproximately 30 μM. This means that the solubility at lower sulfide concentra-tions is extrapolated from the chemical speciation determined at higher sulfideconcentrations. Considering the uncertainty in analytical methods used for deter-mination of Hg in aqueous phase, the problems of separating truly dissolvedspecies from clusters and colloidal HgS, as well as the variation in crystallinityand structure of HgS(s) formed in natural soils and sediments in presence ofNOM, a log K value in the range −36 to −38 seems justified for the reactionHgS(s) + H+ = HS− + Hg2+ at this point.

7.5 REACTIONS OF MERCURY(II) WITH SOIL AND SEDIMENTPARTICLE SURFACES

As a consequence of the strong affinity for reduced sulfur groups, Hg2+ andMeHg+ ions will be adsorbed to functional groups associated with NOM andto metal sulfides. Because sulfur groups almost always are in great excess oftotal concentrations of Hg and MeHg, oxygen functional groups at Fe and Aloxyhydroxides and at the edges of phyllosilicates only play a role as indirectadsorbents of Hg and MeHg complexes. There are a number of experimentalstudies on the direct adsorption of Hg to oxidized surfaces, which are useful andrelevant for certain highly contaminated conditions (Kim et al., 2004; Skyllberg,2010), but they will not be discussed further in this chapter.

7.5.1 Complex Formation with Natural Organic Matter

The role of NOM as an important complexing agent for Hg has been knownfor many decades, as reflected by numerous reports on the significant, positivecorrelation between total concentrations of Hg and DOC in stream water (Mierleand Ingram, 1991). The relationship is so strong that it was recently suggested thatDOC (or even better the UV absorption of DOC) can be used as a proxy for totalHg concentrations in streams (Dittman et al., 2009). The last decade Hg, LIII-edge EXAFS studies (Xia et al., 1999; Hesterberg et al., 2001; Qian et al., 2002;


Yoon et al., 2005; Skyllberg et al., 2006) and binding affinity studies (Lovgrenand Sjoberg, 1989; Hintelmann et al., 1997; Skyllberg et al., 2000; Karlsson andSkyllberg, 2003; Black et al., 2007; Gasper et al., 2007) have revealed that thebonding of Hg and MeHg to NOM pertaining to solid and aqueous phases is dueto the complexation with thiol groups.

7.5.1.1 Inorganic Mercury. The current view is that Hg forms a two-coordinated linear complex with two thiol groups, Hg(SR)2, having a Hg–Sdistance of approximately 2.33 A (Skyllberg et al., 2006). A third sulfur group,which may be a mono- or disulfide group (–S–S–), is involved in a weakerbond (at ∼3.0 A), Fig. 7.6a. This structure has been established by increasingthe Hg concentration to 100 μg/g, which is 300–1000 times higher than thatin most soils and sediments. Still, at this Hg concentration only about 1% ofthe thiol groups, as determined by a combination of Hg EXAFS and S K-edgeXANES, were saturated by Hg.

The spectroscopic results have been complemented by binding affinity studiesconducted at much lower concentrations of Hg and MeHg in isolates of NOMfrom soil and natural water samples. Because of the very low aqueous concen-trations of Hg in equilibrium with the Hg(SR)2 complexes at natural conditions,the binding affinity can be determined only after addition of a competing lig-and. This method is known as competing ligand exchange (CLE) or competitivecomplexation . To separate Hg complexes formed with NOM and with the com-peting ligand, liquid–liquid, solid phase, or dialysis separation methods are used.Because of dissimilarities in methods, samples, conditions, and conventions usedto calculate conditional constants it is quite difficult to compare and evaluate therelevance of reported results. For such an evaluation I refer to compilations madeby Black et al. (2007), Gasper et al. (2007), and Skyllberg (2008). Overall, mostCLE studies conducted at reasonably low Hg to NOM molar ratios have resultedin a log K on the order of ∼20–32 (Hg2+ + L′ = HgL′, where L′ represents adeprotonated ligand with a concentration calculated by a fitting procedure or set

S SCl

Hg

S

S

Fe

S

(a) (b)

C

C

2.33 Å

2.33 ÅHg

2.95

Å

Figure 7.6 Proposed structure for complex formation between Hg and thiol in NOM(Skyllberg et al., 2006) (a) and Hg with pyrite surface groups in presence of 0.2 M ofchloride ions (b). Source: Figure (b) reprinted from Bower et al. (2008) with permissionfrom Elsevier.


equal to NOM). Small values are generally obtained at high additions of Hg inrelation to RSH groups, and values of log K on the order of ∼10 are obtained ifHg is more than saturating the thiol groups and complexation by carboxyl groupsis dominant (Lovgren and Sjoberg, 1989).

As shown by Skyllberg (2008), a log K ∼ 27, for the reactionHg2+ + NOM = Hg–NOM, or a log K ∼ 30, for the reaction Hg2+ + L′ = HgL′at a total ligand concentration of 1 μM, is indicative of Hg–thiol complexation.The LMM thiols L-cysteine, DL-penicillamine, and mercaptoacetic acid havea log β2 for the reaction Hg2+ + 2RS− = Hg(SR)2 in the range 37.8–44.5(Skyllberg, 2008). If the pKa values assigned to RSH and RNH2 groups forthe former two molecules (7.9, 8.5 and 10.4, 10.5, respectively) are consideredmacroscopic in nature, an average of 42 ± 2 can be considered the current bestavailable estimate of log β2 for thiol groups associated with short carbon chainsand having carboxyl and amino groups in the near vicinity.

From combined measurements using S K-edge XANES and Hg LIII-edgeEXAFS, the concentration of thiol groups in NOM from a soil and stream witha high affinity for Hg and MeHg has been calculated to be on the order of0.15% of organic C on a mass basis, or 45 nmol/mg (C) (Skyllberg et al., 2005).Using this estimate in combination with data from Skyllberg et al. (2000), logβ2 values for the reaction Hg2+ + 2RS− = Hg(SR)2 were calculated to be inthe range 43.3–47.7 for organic soils along an upland peat gradient (Skyllberg,2008), as well as for extracted and purified humic substances from soils takenalong the same gradient (Khwaja et al., 2006). The molar RStot to Hg ratio wasbetween ∼10−4 and 10−5 in the former and between ∼10−3 and 10−1 in the latterstudy. Thus, both the size of the stability constants and spectroscopic observa-tions strongly point at the formation of Hg(SR)2 complexes associated with NOMin the aqueous and solid phases of soils. There is a current lack of spectroscopicstudies on the interaction between Hg and NOM in aquatic ecosystems; however,binding affinity studies suggest an interaction with thiols, presumably in a similarlinear RS–Hg–SR configuration as established for terrestrial environments.

7.5.1.2 Methyl Mercury. Similar to Hg–NOM complexes, MeHg forms alinear two-coordinated complex with one Hg–C bond of 2.05 A (the bondingbetween the methyl group and mercury within the methyl mercury molecule)and one Hg–S bond of 2.35 A with the thiol group (Qian et al., 2002; Yoonet al., 2005). The number of CLE studies on MeHg–NOM interactions is smallerthan for Hg (Hintelmann et al., 1997; Amirbahman et al., 2002; Karlsson andSkyllberg, 2003; Khwaja et al., 2010), but since the affinity of MeHg is lower,the methodological problems are less expressed and results are quite consistent.Even if many studies report results pointing at a complex formation with thiols inDOM and particulate NOM from soils, few have actually measured these groups.Karlsson and Skyllberg (2003) and Khwaja et al. (2010) made independent deter-minations of RSH groups using S XANES and calculated stability constants atMeHg to RSH molar ratios less than 2 × 10−6 and 8 × 10−3, respectively. Usinga model with pKa = 9.96 (similar to mercaptoacetic acid), a log K in the range


15.6 (pH 5.1)–17.1 (pH 2.0) at 0.25 M ionic strength (Karlsson and Skyllberg,2003), and 15.5–16.0 at pH 3 and I = 0 (Khwaja et al., 2010), was reported forthe reaction MeHg+ + RS− = MeHgSR in an organic soil and Suwanee Riverfulvic acid, respectively. Considering experimental errors, these constants arewell in agreement with stability constants in the range 15.7–17.5 as reported forthe formation of MeHgSR complexes with LMM thiols, having pKa values inthe range 8.5–9.96. However, in order to fully account for the pH dependencyof the data collected in the above two studies, a thiol group with a pKa valueof 4.0 is needed (Khwaja et al., 2010). Thiols with such a high acidity havebeen determined for clustered cysteine groups (Freisinger, 2008) and thiophenols(Summa et al., 2007). Thus, there are indications that Hg and MeHg may bebound to different types of thiol groups in NOM, something that needs to befurther investigated.

7.5.2 Reactions with Iron(II) Sulfides

Pyrite (FeS2) is generally considered the most abundant iron sulfide in soils,sediments, and waters (Rickard and Luther, 2007). It is a crystalline mineral,which is quite resistant to short-term oxidation events in sediments and soils.Mackinawite (FeSm) is a short-range ordered, nanocrystalline iron sulfide, whichreadily forms in sediments and soils as a reaction between Fe2+ and HS− ions.In terrestrial and freshwater environments it is considered just as abundant aspyrite (Rickard and Morse, 2005). Experimental studies show that the tetragonalcrystalline structure of FeSm forms within milliseconds. Black iron(II) monosul-fides (mackinawite) is commonly referred to as amorphous, even though trulyamorphous FeS(s) has not been identified (Rickard and Luther, 2007). Because offluctuating redox conditions and kinetic constraints, pyrite and mackinawite existside by side in sediments. Although several different reaction paths have beensuggested for the formation of FeS2(s), a two-step oxidation of mackinawite toFeS2 via greigite (Fe3S4), through reactions with sulfur species with intermediateoxidation states (e.g., elemental S, polysulfides, or thiosulfate) or with Fe(III), hasbeen frequently argued as being of major importance in sediments (Wilkin andBarnes, 1996). In laboratory experiments, FeSm and FeS2(s) have been shownto actively remove Hg from solution (Bower et al., 2008; Liu et al., 2008), andin order to model the chemical speciation of Hg, the mechanisms and thermo-dynamics of these reactions need to be resolved. At this point, no studies havecovered interactions between MeHg and iron sulfides.

7.5.2.1 Reactions with Pyrite. The reaction between Hg and FeS2(s) hasbeen studied by use of adsorption–desorption and spectroscopic studies. DespiteHgS(s) being thermodynamically more stable than FeS2(s), significant dissolu-tion of FeS2(s) under the formation of HgS(s) has not been observed duringthe timeframe of any experimental study (Hyland et al., 1990; Ehrhardt et al.,2000; Behra et al., 2001; Bower et al., 2008). This may be explained by the highkinetic stability of pyrite. Adsorption–desorption experiments have shown that


surface bonds formed between Hg and ≡Fe–S–S surface groups are relativelyweak (e.g., in comparison to Hg–S bonds in NOM) and 40% of surface-boundHg was desorbed by the addition of 0.1 M of chloride (Behra et al., 2001).X-ray photoelectron spectroscopy (XPS) Suggested the formation of a linear≡Fe–S–S–Hg–OH surface complex in the absence (Behra et al., 2001) and alinear ≡Fe–S–S–Hg–Cl in the presence of 0.1 M NaCl (Behra et al., 2001;Bower et al., 2008) (Fig. 7.6b). In contrast to several other metals (Huerta-Diazand Morse, 1992), there are no indications that Hg is pyritized, that is, incorpo-rated into the FeS2 structure.

7.5.2.2 Reactions with Mackinawite. In contrast to reactions with pyrite,Hg has been shown to react with and dissolve FeSm during the formation of β-HgS(s) (Jeong et al., 2007, 2010; Skyllberg and Drott, 2010). In another study,a four-coordinated HgS compound (in agreement with β-HgS(s) formation) wasformed under partly oxidized conditions, whereas at strictly reducing conditionsa distorted surface complex was suggested to form Wolfenden et al. (2005). Asshown with pyrite, the ability of iron sulfide to withstand dissolution by Hglikely is related to its kinetic stability and crystallinity. Therefore, the time ofFeSm ageing could be important for whether reactions with Hg will lead tosurface complex formation, dissolution–precipitation reactions, or a mixture ofthese processes. In the study by Skyllberg and Drott (2010), FeSm was allowedto age for 8 h, and no significant surface complexation was observed by Hg LIII-edge EXAFS at Hg to FeSm molar ratios of 0.002–0.012. In contrast, surfacecomplexation of Hg onto FeSm aged for three days was observed at a Hg toFeSm molar ratio less than 0.05 in an adsorption experiment (Jeong et al., 2007).At higher molar ratios, FeSm was dissolved and β-HgS formed (as determinedby XRD). Substitution of Hg2+ for Fe2+ ions in the FeSm structure has not beenobserved in any spectroscopic study. In order to resolve at which conditionssurface adsorption, dissolution, and β-HgS(s) formation occurs, more strictlydesigned time-dependent experiments are needed in which adsorption is followedby spectroscopic methods. Furthermore, the influence of NOM on the reactionbetween Hg and FeS(s) need to be considered in order for results to be relevantunder natural conditions in soils and sediments.

7.5.2.3 Reactions with Mackinawite in the Presence of Natural OrganicMatter. In soils and sediments NOM and FeSm may coexist, acting as veryimportant components in the control of pH and pe. The binding of Hg to thiolgroups in NOM is well studied, but what will happen in the presence of FeSm? Inthe only existing study, it was shown that FeSm aged for five days was dissolvedby Hg in the presence of NOM under the formation of Hg(SR)2 complexes andβ-HgS(s) (Skyllberg and Drott, 2010) at pH 5.7–6.1. Hg(SR)2 complexes andβ-HgS(s) (as determined by Hg EXAFS) coexisted after 30 h of equilibration.An increase in FeSm from 2 to 20 mass% (and decrease in NOM from 98 to80 mass%) resulted in a change of Hg speciation from 43 to 100% β-HgS(s)(consequently, Hg(SR)2 decreased from 47% to 0%). A simple thermodynamic


calculation based on the composition of the two components was shown tobe in reasonable agreement with a log solubility product of β-HgS(s) of −37and a log β2 for the formation of Hg(SR)2 of 42 or greater. Results from thisexperiment suggest that in the presence of NOM, Hg complex formation at thesurface of FeSm is of minor importance as compared to secondary precipitation ofβ-HgS(s).

7.5.3 Reactions with Other Metal Sulfides

By using XPS and Auger electron spectroscopy it was observed that Hg readilydissolved galena (PbS) under the formation of HgS(s) (Hyland et al., 1990). Thisis explained by the higher solubility of the lead(II) sulfide. Some Hg also formedweak complexes that passivated the galena surface from continued formation ofHgS(s). Studies on the adsorption of mercury to metal sulfides such as CuS,Cu2S, and ZnS have so far not been conducted.

7.5.4 Indirect Adsorption to Oxygen Groups at Mineral Surfaces

Abundant and reactive oxygen functional groups associated with the surfacesof clay minerals, Mn, Al, and Fe oxyhydroxides, are known to be of majorimportance for the complexation of transition metals, such as Mn, Co, Ni, Cu,and Pb, in soils. These metals form inner- and outersphere bonds with hydrated≡XOH surface groups (where X = Mn, Fe, Al, Si). Because of ubiquitous NOMand the strong affinity for thiol groups, Hg and MeHg will not bind directlyto oxygen functional groups, except at exceptionally contaminated conditions invery low organic matter environments. However, reactive ≡XOH surface groupsmay form ternary complexes with low- and high-molecular-mass Hg(SR-DOM)2

and MeHgSR-DOM molecules. The most likely interaction is a “Type B” ternarysurface complex formation where the ≡XOH surface group interacts with anoxygen (carboxyl) and/or nitrogen functional group of the organic molecule.Using Hg-cysteine (−OOCHN3

+CH2S)2Hg as a model, depending on pH andthe point of zero charge of the mineral surface, the negatively charged carboxylgroup, the positively charged amino group, or both groups may interact with the≡XOH surface group. There are numerous studies on the interaction betweenHg and oxidized mineral surfaces such as illite (Hamilton et al., 1995), kaolinite(Sarkar et al., 2000), and goethite (Barrow and Cox, 1992), but only a few on theinteraction with MeHg (Gunneriusson et al., 1995; Desauziers et al., 1997). In thefew studies in which NOM- (Yin et al., 1996) or thiol- (Senevirathna et al., 2011)containing molecules have been included, the conclusion has been that Hg(II)does not adsorb directly to ≡XOH surface groups. There is so far only one studyin which Hg, thiols, and mineral surfaces have been considered in a mixed system,highlighting the importance of ternary surface complexes for the regulation ofadsorption and desorption of Hg-thiols onto surfaces of kaolinite (Senevirathnaet al., 2011). Desorption of Hg(SR-DOM)2 and MeHgSR–DOM complexes islikely responsible for the observed concurrent release of Hg and MeHg together

SOLUBILITY AND CHEMICAL SPECIATION OF MERCURY(II) IN SOILS AND SEDIMENTS 237

with Fe and Mn from lake sediments into the bottom water (Regnell et al.,2001). Reduction of Fe(III) and Mn(III, IV) oxyhydroxides during summer orwinter stratification results in a degradation of the surfaces of these mineralsand a subsequent release of Hg and MeHg complexed by organic substances.Under anoxic conditions, iron(II) sulfide surfaces may be important adsorbents forHg(SR-DOM)2 as well as Hg complexes with inorganic sulfides and polysulfides.These processes still remain to be studied in controlled experiments.

7.6 STABILIZATION OF NANOPARTICULATE MERCURY(II) SULFIDESBY NATURAL ORGANIC MATTER

From a thermodynamic point of view, NOM associated thiols will compete withinorganic sulfides for Hg under anoxic conditions. Thus, in the presence of DOMthe dissolution of HgS(s) will be enhanced as a consequence of Hg(SR-DOM)2

complex formation. In addition to this process, nanoparticulate HgS(s) physic-ochemically stabilized by DOM has been shown to contribute to an apparentdissolution of HgS(s) (Ravichandran et al., 1999; Deonarine and Hsu-Kim, 2009).Thus, in experiments where HgS(s) has been allowed to precipitate out from solu-tion, DOM has been demonstrated to have a dual role of decreasing the growthrate and aggregation of HgS polymers and at the same time stabilizing HgScolloids (Deonarine and Hsu-Kim, 2009; Slowey, 2010). Results from solubilityexperiments with cinnabar and DOM may suggest that hydrophobic fractions ofDOM enhance the apparent solubility more than hydrophilic fractions of DOM(Ravichandran et al., 1998; Waples et al., 2005), independent of the concentrationof thiol groups in these fractions. In apparent contrast, LMM amino acids withthiol groups have been shown to be much more efficient than molecules withamino and carboxyl groups alone to inhibit aggregation and to stabilize HgSnanoparticulates in experiments where metacinnabar has been allowed to precip-itate after reaction between HS− and Hg2+ (Deonarine and Hsu-Kim, 2009). Inthe latter type of experiment, it has been shown that reactions between DOM,Hg2+, and HS− and the formation of metacinnabar are dynamic over time anddo not seem to reach equilibrium even after weeks (Slowey, 2010).

7.7 SOLUBILITY AND CHEMICAL SPECIATION OF MERCURY(II)IN SOILS AND SEDIMENTS

7.7.1 Linkage to the Biogeochemistry of Natural Organic Matter,Sulfur, and Iron

Natural dynamics in the formation and degradation of inorganic and organic sul-fur species and their reactions with NOM and iron, to a large extent, determinethe speciation of Hg and MeHg in soils, sediments, and waters. During oxicconditions, Hg and MeHg exist exclusively as organic complexes with low- andhigh-molecular-mass thiols, but when redox conditions are lowered to the regimes


of Fe(III) and sulfate reduction, the situation gets more complex. A typical sit-uation in sediments of estuaries and lakes is the formation of an oxic–anoxicinterface in surface sediments where terrestrial sources of Fe(III), transported bystreams and rivers, react with H2S–HS−, produced by SRB in anoxic sediments,under the formation of sulfur species with intermediate oxidation states (elemen-tal S, polysulfides, thiosulfate) (Fig. 7.7). Under such circumstances, formationof organic thiols will take place because of diagenetic incorporation of H2S intoorganic substances (Anderson and Pratt, 1995). In the presence of zero-valentsulfur, Hg-polysulfides will form and increase the solubility of Hg (Fig. 7.4).Indications of an increased solubility caused by the formation of Hg-polysulfideshave been observed in controlled sulfurization experiments in the presence ofiron oxyhydroxides (Slowey and Brown, 2007). Yet, these reactions remain to bedemonstrated in natural soils and sediments.

During periods with high flow conditions, as well as during spring and autumncirculation in dimictic lakes, the oxic–anoxic interface in lake sediments movesdownward, and during low flow conditions, such as summer and winter waterstratification, the oxic–anoxic interface moves upward. These fluctuations cause

O2oxic

anoxic

Fe2+

H2S

FeS2

H2S, Sx

FeSaq, FeS

H2S S2O3Sx

O2OH–

Fe(OH)3(s)

S8

Fe3+L

H2O2

H2O

–

2–

2–

2–

Figure 7.7 The dynamic interaction between iron and sulfur biogeochemistry atoxic–anoxic sediment or soil interfaces. Source: Reprinted from Rickard and Luther(2007) with permission from American Chemical Society.


dynamic changes in the speciation of Hg and MeHg, as well as in the trans-portation processes between the sediment pore water and bottom water. Alsomethylation and demethylation reactions, which are tightly associated with theactivity of FeRB and SRB, will greatly affect the quantity of MeHg. The dynam-ics of Hg and MeHg solubility and mass transfer within sediments and in theinterface between pore and bottom waters have been frequently studied (Hurleyet al., 1994; Regnell et al., 2001; Goulet et al., 2007; Merritt and Amirbahman,2007). However, because of an incomplete understanding of reactions with notonly NOM and polysulfides but also with sulfides such as FeSm and pyrite, therehave been few attempts to determine the chemical speciation of Hg and MeHg inthese dynamic environments. Dissolved and particulate concentrations and par-titioning coefficients (log Kd) have been reported for total Hg and MeHg innumerous studies, but there are so far very few in which the partitioning of Hgor MeHg between aqueous and solid phases of soils or sediments has been usedto evaluate chemical speciation models.

7.7.2 Dominance Diagram for Solid and Surface Phases

On the basis of the stability constants reported in Appendix I, the dominating Hgspecies in the solid phase of soils and sediments can be calculated as a functionof pe and pH (Fig. 7.8a). This diagram differs from traditionally publisheddiagrams by including Hg complexed by thiol groups associated with NOM.The Hg(SR-NOM)2 complex dominates over the whole pH scale, and the borderto HgS(s) formation coincides with the H2S/HS− –SO4

2− redox transition. It isknown since at least 35 years (Andren and Harriss, 1975) that the complexationof Hg to NOM, under all except extreme conditions of contamination,outcompetes solid phases such as HgO and HgCl2, as well as their aqueousphase analogs Hg(OH)2

0(aq) and HgCl20(aq). Despite this insight, a lack ofproper methods has until recently restricted the possibility to determine stabilityconstants for Hg-NOM complexes. Furthermore, the convention to expressHg–NOM complexes as conditional HgL′ or Hg–NOM complexes, rather thanhaving a defined molecular structure, has precluded the use of stability constantsfor Hg–NOM complexes together with well-defined inorganic molecules inthermodynamic modeling. Therefore, recent spectroscopic information onHg–NOM complexes (using Hg LIII-edge EXAFS), and a quantification of thiolgroups in NOM by use of S K-edge XANES, has not only provided chemicalstructures but also facilitated the formulation of stoichiometric reactions andthermodynamic modeling of Hg and MeHg in soils, sediments, and waters.

It can be noted in Fig. 7.8a that Hg0, which in traditional pH-pe diagramshas been illustrated as the dominant form both at pe values above and below thestability field of HgS(s) (Andersson (1979), is outcompeted by Hg(SR–NOM)2

complexes. Thus, the stability of Hg0(l) is restricted to highly anoxic conditionswhere the boundary between HgS(s) and Hg0(l) depends on the solubility productchosen for the former. In Fig. 7.8a, a log Ksp of −37.6(I = 0) for metacinnabaris used. If instead a solubility product of −39.1 (Martell et al., 2003) is used,


2–10

–5

0

pe 5 S8Hg(SR-NOM)2

Hg0(I)H2S

SO4

HS

2–β -HgS(s)

0

10

15

20

3 4 5 6

pH

(a) (b)

7 8 9 10 3–9

–8

–8

log

[S-I

I] (M

)

–6

β -HgS(s)

50% org-C

MarineFeSm stabilityfields

15% org-C

2% org-CHg(SR-NOM)2

–5

–4

–3

–2

4 5 6

pH

7 8 9

fresh waterFreshwater

Figure 7.8 Dominance diagrams for soils and sediments with 100 ng Hg/g dry mass and90 mass% pore water content. (a) The solid phase contains 15% organic C (NOM having aconcentration of thiol functional groups corresponding to 0.8 mM on a pore water basis):the concentration of inorganic sulfur is 0.1 mM. Solid lines show the stability of mercuryspecies, black broken lines show the stability of water and grey broken lines show thestability of sulfur species. (b) Limiting concentrations of S(−II)(= [H2S] + [HS−]) forthe precipitation of metacinnabar in sediments with different concentrations of organic C.Regulation of S(-II) by the solubility of FeSm is indicated at freshwater (area bordered bydotted line) and marine conditions (black field). The stability of FeSm is in equilibriumwith Fe2+ concentrations of 10−3 to 5 × 10−5 M in freshwater and 10−6 to 5 × 10−8

in marine environments. Stability constants used in the calculations were taken fromAppendix I, without correction for ionic strength and temperature effects. The stabilityconstant (log K) for FeSm + H+ = Fe2+ + HS− is −3.5 (Rickard, 2006).

the boundary between HgS(s) and Hg0(l) will coincide with the lower boundaryfor the stability of water. The suggested stability of Hg0(l) in sediments is inagreement with thermodesorption analyses showing that traces of elemental Hgreleased by chlor-alkali industry may still persist under highly anoxic conditions(Biester et al., 2002).

The stability constants for β-HgS(s) and Hg(SR-NOM)2 can be used tocalculate environmental conditions at which β-HgS(s) starts to precipitate. Asillustrated in Fig. 7.8b, the limiting concentration of S(−II) = [H2S] + [HS−]needed to precipitate β-HgS(s) is 0.004, 0.25, and 2.5 μM at pH 6 and less insoils/sediments with 2%, 15%, and 50% of organic C, respectively. At pH valuesbelow 6.2, Hg(SH)2

0(aq) is the dominant dissolved Hg–sulfide complex, andthe solubility of β-HgS(s) is practically independent of pH (Fig. 7.4c). When thepH is raised above 6.2, the solubility of Hg increases because of an increasingdominance of HgS2H− and HgS2−

2 in solution. Because of this, the limitingS(-II) concentration for precipitation of β-HgS(s) increases above pH 6.2. AtpH 7 in a pure organic peat soil (having 50% organic C), 4 μM of S(-II) will beneeded to precipitate β-HgS(s); in a marine sediment (pH 8.3) with 2% organic


C, only 0.1 μM of S(-II) is required, whereas in a salt marsh (pH 8.3) with 15%organic C, 5.0 μM of S(-II) would be required. It should be noted that thesevalues are highly dependent on the stability constants chosen. The Ksp reportedfor β-HgS(s) varies by roughly three orders of magnitude, and the uncertaintyin the log K for the formation of Hg(SR-NOM)2 may be even greater.

The formation of β-HgS(s) in sediments and soils is due to an interplayamong organic matter, sulfides, and iron. If HS− is regulated by FeSm, β-HgS(s)likely will be found as a precipitate in most acid wetland soils as noted fromFig. 7.8b. However, in neutral soils and freshwater sediments containing FeSm

(having intermediate or high concentrations of Fe2+), whether β-HgS willform or not depends on the organic matter content. Under marine conditions(having low concentrations of Fe2+), the presence of FeSm under almost allcircumstances would indicate the formation of β-HgS(s). In salt marsh soils,high organic matter contents and relatively high concentrations of Fe2+ willcounteract the formation of β-HgS.

7.7.3 Modeling of Mercury(II) Speciation in the Aqueous Phaseof Soils and Sediments

Two different types of thermodynamic modeling procedures may be recognized:

1. Calculation of dissolved Hg and MeHg species from total concentrationsin the aqueous phase without taking reactions with solid phases into con-sideration.

2. Calculation of the Hg and MeHg solubility reactions, validated by actualdata from soils and sediments.

The first type of model is frequently used to calculate concentrations of Hgand MeHg aqueous species involved in diffusion and biouptake by methylat-ing bacteria and higher organisms in soils and sediments. Pure bacterial cultureexperiments have shown that Hg complexed by DOM decreases the uptake ofHg (Kelly et al., 2003), obviously by lowering the concentration of small Hgmolecules available for uptake. The current view is that Hg uptake may be bothpassive, by a diffusion of neutral forms of Hg through cell membranes of methy-lating bacteria (Barkay et al., 1997; Morel et al., 1998), as well as active, by theuptake of Hg complexed by LMM organic thiols (Schaefer and Morel, 2009). Forpassive uptake, first-order uptake rates will be directly related to the concentra-tion of neutral forms of Hg (Benoit et al., 2003). This means that only chemicalforms with reasonably high concentrations will be of quantitative importance fora passive uptake.

A thermodynamic calculation, using the constants in Appendix I, at anoxicconditions (under which HS− is regulated by the solubility of FeSm) in fresh-water, brackish water, and marine conditions shows that the highly dominantneutral form is Hg(SH)2

0. Possibly, HgClSH0 and HgOHSH0 can be includedas candidates for passive uptake under certain conditions (Table 7.2). Under


oxic conditions, HgCl20, Hg(OH)20 and HgOHCl0 are the only known neutral

forms, and neither of them reaches concentrations at a reasonable level for passiveuptake. This result is partly in contradiction to previous perceptions (Morel et al.,1998), when much smaller stability constants were used for Hg bound to organicthiols in NOM. Under oxic (and suboxic) conditions, the lack of neutral formsshowing a reasonable concentration (>10−20 M) leaves active uptake of LMMorganic complexes with thiols as the only realistic uptake mechanism by methy-lating bacteria. Mercury in these complexes may be two-coordinated, and mixedcomplexes with chloride and hydroxyl ions reach concentrations that may be ofquantitative importance for biouptake (Table 7.2). It is noteworthy that the LMMHgClSR(aq) complex may reach higher concentrations than Hg(SR)2(aq) undermarine conditions. This may suggest that mixed thiol–chloride Hg complexescould contribute to bacterial uptake under high-salinity conditions.

7.7.4 Chemical Speciation Models Validated by Solubility Datain Soils and Sediments

Very few Hg speciation models have been validated by actual data from soils,sediments, or waters. An attempt was made with the mercury cycling model(MCM); (Hudson et al., 1994), a thermodynamic model for Hg and MeHg par-titioning between seston (particulate NOM and plankton) and DOM in the watercolumn of lakes linked to a kinetic model for biouptake. Owing to the lackof stability constants for Hg and MeHg interactions with particulate and dis-solved organic matter (OM), log K for the reactions Hg2+ + L− = HgL+ andMeHg+ + L− = MeHgL and the concentration of organic ligands (HL) wereoptimized. The fitted log K values of 17.5 and 12.5 for Hg and MeHg bindingto DOM were compared to log K values of 22.1 and 15.7, respectively, reportedfor HgSR+ and MeHgSR complexes (Dyrssen and Wedborg, 1991). Because theoptimization included facilitated Hg2+ and MeHg+ uptake by plankton, it wasnot a true validation of the thermodynamic part of the model. It should be notedthat a log β2 of 42 for the formation of Hg(SR-NOM)2 corresponds to a log K of∼33 if written as Hg2+ + L− = HgL+ (Skyllberg, 2008). Thus, even if the out-come of the modeling was interpreted as an involvement of thiols in the bindingof Hg and MeHg, data were not consistent with the formation of two-coordinatedHg-thiol complexes.

In order to explain octanol–water partitioning of Hg in equilibrium withHgS(s) (Benoit et al., 1999b) and observations of an inverse relationshipbetween MeHg concentrations in sediments and dissolved inorganic sulfides(Benoit et al., 2003) along a nutrient gradient in the Florida Everglades(Gilmour et al., 1998) and the Patuxent River (Benoit et al., 1998), Benoitet al. (1999a) developed a semiempirical thermodynamic model for Hg. Thetheory behind the model builds on the hypothesis that the net formation ofMeHg is related to an uptake of neutral forms of Hg-sulfides by methylating


TABLE 7.2 Aqueous Phase Concentrations (Moles per Liter) of Major ChemicalSpecies of Hg (Excluding Polysulfide Complexes) and Possible Candidates forDiffusion, Evasion, and Biouptake in the Sediment–Bottom Water Interface UnderOxic and Anoxic Conditions

Freshwater/ Brackish MarineWetland Soil Water Conditions

pH 7.0 7.5 8.3Cl− 10−4 0.1 0.6DOC 4.2 × 10−3 (50 mg/L) 8.3 × 10−4 (10 mg/L) 4.2 × 10−4 (5 mg/L)

Oxic Anoxic Oxic Anoxic Oxic AnoxicDOM thiols 3.8 × 10−6 3.8 × 10−6 1.2 × 10−6 1.2 × 10−6 6.3 × 10−7 6.3 × 10−7

LMM thiols 10−9 10−7 10−9 10−7 10−9 10−7

HS− 0 3.2 × 10−8 0 2.0 × 10−6 0 3.2 × 10−5

S(-II)tot 0 3.3 × 10−7 0 2.6 × 10−6 0 3.4 × 10−5

HgCl20 3.8 × 10−29 1.4 × 10−29 1.9 × 10−24 1.3 × 10−27 1.1 × 10−23 2.2 × 10−29

Hg(OH)20 2.4 × 10−29 8.7 × 10−30 1.2 × 10−27 8.2 × 10−31 7.6 × 10−27 1.5 × 10−32

HgOHCl0 4.8 × 10−29 1.7 × 10−29 7.6 × 10−26 5.9 × 10−29 4.6 × 10−25 9.2 × 10−31

HgOHSH0 0 8.2 × 10−21 0 1.6 × 10−21 0 7.8 × 10−23

HgClSH0 0 2.4 × 10−21 0 1.5 × 10−20 0 6.9 × 10−22

Hg(SH)20 0 3.9 × 10−12 0 4.1 × 10−13 0 4.0 × 10−14

Hg(SR-LMM)2 1.5 × 10−18 5.5 × 10−15 3.8 × 10−17 2.6 × 10−16 1.5 × 10−16 3.1 × 10−18

HgClSR-LMM 1.2 × 10−19 4.3 × 10−18 6.0 × 10−18 4.1 × 10−19 5.7 × 10−16 1.2 × 10−21

HgOHSR-LMM 6.0 × 10−20 2.2 × 10−18 9.5 × 10−18 6.5 × 10−23 9.6 × 10−20 1.9 × 10−23

Hg(SR-DOM)2 1.0 × 10−11 3.6 × 10−12 1.0 × 10−11 5.5 × 10−15 1.0 × 10−11 2.0 × 10−17

HgS2SH− 0 2.5 × 10−12 0 8.3 × 10−12 0 5.0 × 10−12

HgS2−2 0 1.2 × 10−14 0 1.3 × 10−12 0 5.0 × 10−12

aThe total concentration of Hg is 10−11 M. The concentration of organic thiol groups associatedwith DOM is calculated as 0.15 mass% of DOC (Skyllberg, 2008). The concentration of LMMthiols is set to 1 nM under oxic conditions and 100 nM under anoxic conditions. The pKa of thiolsis set to 10.0. Under anoxic conditions, the concentration of HS− is regulated by the solubility ofFeSm + H+ = Fe2+ + HS−; log K2,SP = −3.5(I = 0; Rickard, 2006) with [Fe2+] = 1 × 10−3, 5 ×10−6, and 5 × 10−8 M in freshwater, brackish water, and marine conditions, respectively. Stabilityconstants in Appendix I, corrected for the ionic strength effect by Davies equation, are used in thecalculation. Temperature effects are not considered.

bacteria. It was also observed that while the solubility of Hg, as would beexpected (c.f. Fig. 7.4a and b), increased with increasing concentrations ofdissolved sulfides along the Patuxent River gradient, this was not the casein Florida Everglades. By including a reaction for diagenetic incorporationof HS− into an organic matter solid phase (ROH + HS− = RSH + OH−),and a complexation reaction between Hg2+ and organic thiols [HgSR+ andHg(SR)2], a model was created in which the Hg solubility was roughlyindependent of the dissolved sulfide concentration. Furthermore, by using alog K of 40.5 for the reaction HS− + OH− + Hg2+ = HgOHSH0(aq), which


corresponds to a log K of −10.0 for the reaction HgS(s) + H2O =HgOHSH0(aq), the HgOHSH0 molecule became quantitatively dominant andwas significantly positively correlated to the sediment concentration of MeHg atboth sites.

The model proposed by Benoit and coworkers can be said to be one of the fewthat has been validated by Hg solubility data (even if a strict validation with calcu-lation of merit of fit was not conducted) from natural sediments. It is also the firstthat considered a solid phase of organic thiols associated with NOM as importantcomponents regulating the solubility of Hg, even if the concentration of thiolswere not determined. There are, however, several uncertainties in the model. Thediagenetic incorporation of HS− in NOM is a complicated reaction that has notbeen verified by experimental data, neither in terms of kinetics nor thermodynam-ics. The log K for this reaction can therefore be regarded as a fitting parameterin the model. It may be plausible that (if other factors are being constant) the dia-genetic formation of organic thiols increases with the concentration of dissolvedsulfide, but this remains to be shown. Furthermore, as discussed in Section 7.3.2.2,the log K of 40.5 set for the formation of HgOHSH0, is likely substantiallyoverestimated.

Drott and coworkers, using a model including DOM-associated thiols,reported a positive relationship between the sum of the calculated concentrationsof Hg(HS)2

0(aq) and HgOHSH0(aq) and the methylation rate constant (km),as well as total concentrations of MeHg, in contaminated sediments (Drottet al., 2007). Similar to that of Benoit et al. (1999a), this model was biased byan unsubstantiated, large log K for the formation of HgOHSH0(aq), even ifthe uncertainty in this constant was recognized and its size varied in order toaccount for this. Reactions with HgS(s) were considered in the regulation ofthe concentration of HgOHSH0(aq), but reactions with solid phases were notallowed to limit the overall solubility of Hg. Thus, the model was not validatedby solubility data.

7.7.5 Thermodynamic Modeling of Partitioning Coefficients ((Kd)for Hg and MeHg in Soil and Sediment

The partitioning of Hg and MeHg between solid (particulate) and aqueous phasesis commonly reported for soils, sediments, and waters. The distribution coeffi-cient Kd (L/kg) is defined as the concentration in the solid (or particulate) phase(mol/kg or g/kg) divided by the concentration in the aqueous phase (mol/L org/L). In an oxic environment, the partitioning of Hg and MeHg will be deter-mined by the binding to thiol groups and in turn their partitioning betweensolid and aqueous phases of NOM. If we assume that the binding affinity ofHg to thiols is independent of whether RS belongs to the solid (designated RS(ads)) or aqueous (designated DOM-RS (aq)) phase, the stability constant logK for Equation 7.1 is 1.0 (all concentrations are expressed as moles per liter


solution).

Hg(SR − DOM)2(aq) + 2RS−(ads) =2DOM-RS−(aq) + Hg(SR)2(ads);K = 1 (7.1)

Kd = [Hg(SR)2(ads)]/[Hg(SR − DOM)2(aq)]

= [RS−(ads)]2/[DOM − RS−(aq)]2 (7.2)

If the concentration of thiols in relation to organic C is assumed to be equal insolid and aqueous phases, concentrations of solid and dissolved organic carbon(SOC and DOC) can be used as proxies for thiols in solid [RS− (ads)] andaqueous phases [DOM-RS− (aq)], respectively. The range of SOC concentrationsin most soils and sediments is roughly 10–500 g/kg, and the range of DOCconcentrations in pore waters is roughly 10–300 mg/L, with concentrations ofSOC and DOC more or less positively correlated. This means that a log Kd forHg calculated by Equation 7.2 should roughly cover a range of 6.0–6.4 L/kg[(10 × 10−3)2/(10 × 10−6)2 = 106 with SOC 10 g/kg and DOC 10 mg/L and(500 × 10−3)2/(300 × 10−6)2 = 106.4 with SOC 500 g/kg and DOC 300 mg/L].This means that Kd for Hg under oxic conditions can be approximated as Kd

for organic carbon raised to 2. For MeHg, which forms monodentate complexeswith thiols, the partitioning of MeHg simply follows the partitioning of organicC. Thus, Kd for MeHg is equal to Kd for organic C, which results in a range oflog Kd between 3.0 and 3.2 if the above data on SOC and DOC are applied toEquations 7.3 and 7.4.

MeHgSR − DOM(aq) + RS−(ads) =DOM − RS−(aq) + MeHgSR(ads); K = 1 (7.3)

Kd = [MeHgSR(ads)]/[MeHgSR − DOM(aq)]

= [RS−(ads)]/[DOM − RS−(aq)] (7.4)

In agreement with the above calculation, log Kd for Hg is generally observed tobe in the range 5.0–6.5, whereas the log Kd for MeHg often is reported to be inthe range 2.5–3.5 in soils and sediments. The roughly three orders of magnitudegreater Kd for Hg is thus attributed to the exponent of 2 in Equation 7.2, causedby the requirement of two independent thiol groups for the formation of theHg(SR)2 molecule. If the complexation of Hg with RS− instead would havebeen expressed as a bidentate complex (i.e., the thiol groups belonging to thesame molecule; HgS2R), which has been proposed (Khwaja et al., 2006), the logKd for Hg would have been 3.0–3.2, similar to MeHg. Even if the activity ofsurface complexes is not a trivial issue, the fair correspondence between measuredand theoretical log Kd may therefore be taken as support for an involvement oftwo independent thiol groups in the complexation of Hg.


0.01

log

Kd

(L/k

g)

2

3

4

5

6

7 pH 4.0 pH 7.0

2.0 4.0

MeHg, no surface complexation with MeSMeHg, no surface complexation with MeS(s)

No HgS, no polysulfides

No HgS, no polysulfides

HgS, no polysulfides HgS, no polysulfides

HgS and polysulfidesincluding HgSnOH–

HgS and polysulfidesexcluding HgSnOH–

HgS and polysulfidesincluding HgSnOH–

S(–II)Tot (μM)

6.0

(a) (b)

8.0 10.0 0.01

log

Kd

(L/k

g)

2

3

4

5

6

7

2.0 4.0

S(–II)Tot (μM)

6.0 8.0 10.0

Figure 7.9 Theoretical modeling of the partitioning of Hg and MeHg between solid andaqueous phases in a hypothetical wetland soil or sediment having 15% organic C, DOC50 mg/L, and an ionic strength of 0.5 mM. The concentration of thiols (RS) was setto 0.15 mass% of organic C in both aqueous and solid phases. The water content was90% of wet soil mass. Total concentrations of Hg and MeHg were 100 and 0.1 ng/g,respectively. The figures were modified from Skyllberg (2008). Stability constants usedin the calculations were taken from Appendix I, without correction for ionic strength andtemperature effects. The stability constant for the formation of HgSnOH− is taken fromJay et al. (2000). In presence of polysulfides, the activity of elemental S is set to 1.0.

As illustrated in Fig. 7.9, the maximum log Kd is obtained under oxic condi-tions, in the absence of inorganic sulfides. When sulfides are formed and build upin solution, log Kd will decrease because of the formation of soluble Hg-sulfidesand polysulfides. Thus, even at the point when HgS(s) is formed, which occursat ∼0.6 μM S(-II) in the hypothetical soil/sediment shown in Fig. 7.9a and at 1.2μM S(-II) in Fig. 7.9b, log Kd will continue to decrease with increasing concen-trations of S(-II) in solution. It can be noted by comparing Fig. 7.9a and 7.9b thata greater solubility of Hg increases at higher pH, which is in opposition to thegeneral notion that the solubility of metals increase with decreasing pH. Evenif the free Hg2+ ion increases at lower pH, the shift in speciation of sulfidesfrom HS− to H2S results in decreased concentrations of dissolved Hg-sulfides.Another effect is the formation of polysulfides, which contributes to an increasedsolubility at higher pH. In Fig. 7.9b, the dramatic effect on the solubility causedby the stability constant suggested for the HgSnOH− molecule (Jay et al., 2000)is illustrated, increasing the solubility of Hg similar to levels of MeHg. In sum-mary, under strong acidic conditions, the log Kd for Hg is predicted to be inthe range of 4.5–6.0, whereas in neutral and alkaline soils and sediments, logKd would vary roughly between 3 and 6, depending on whether polysulfides areformed or not.


In a review of studies from marine and estuarine sediments, log Kd for Hgwas reported to vary between 5 and 6 (Fitzgerald et al., 2007); a similar rangewas reported in freshwater sediments (Turner et al., 2001). Most of these studiescover true anoxic conditions (sulfide concentrations exceeding 10–20 μM), andpolysulfides may not be anticipated to form. There are, however, several studiesin which a log Kd on the order of 3.5–4.5 has been reported, under conditions inwhich polysulfide formation can be suspected. Goulet et al. (2007) reported anaverage log Kd of 4.2 in the sediment of a riverine wetland in which elementalS was measured (concentrations up to 8 μM in pore water). Similarly, an abruptdecrease in log Kd from 4.5–5.7 in the oxidized water column (where Hg wouldbe complexed by particulate and dissolved organic thiols alone) to a log Kd

of 3.4–4.1 in the top 30 cm of the underlying sediment of Little Rock lake(Hurley et al., 1994) may be explained by polysulfide formation in the sediment(elemental S was not measured). The observation that log Kd dropped markedlyfrom 4–5 to 3–4 when the concentration of acid volatile sulfur (AVS) decreasedfrom above 20 to 0–10 μmol/kg (Hammerschmidt and Fitzgerald, 2004) in a near-shore marine sediment may be taken as an indication that polysulfide formationunder suboxic conditions contributed to the solubility.

The log Kd for MeHg varies roughly between 1.5 and 5.4 in marine sed-iments (Hammerschmidt and Fitzgerald, 2004; Fitzgerald et al., 2007). Valuesin the low end may be consistent with the model used in Fig. 7.9b, whereasvalues above 4.0 are not. An obvious weakness of the proposed model is thelack of thermodynamic data on surface complexation of MeHg to iron sulfidesand other metal sulfides. The formation of such complexes could counteract thedecrease in log Kd with an increase in dissolved sulfide concentrations, but itremains to be shown that they are strong enough to explain log Kd values onthe order of five or above, as reported in some studies. Another explanation forsome unexpectedly high log Kd for MeHg may be related to the hydrophobicityof MeHg, recently proposed to increase the affinity for certain types of thiols(Khwaja, et al., 2010). These thiols may show different concentrations in rela-tion to organic C of solid and aqueous phases. Furthermore, the fact that thehighest log Kd for MeHg (up to 6.0) has been reported as a partitioning betweenthe aqueous phase and suspended particles (including living organisms) in fresh-waters (Watras et al., 1998; Babiarz et al., 2001) suggests that the partitioningof MeHg may include Hg methylation and buildup of MeHg in bacteria andphytoplankton.

Even if there are numerous studies in which log Kd has been reported, fewincludes measurements of all parameters needed for modeling Hg and MeHg solu-bility. As illustrated in Fig. 7.9, in addition to concentrations of sulfide, elementalS is the most important parameter behind Hg solubility, which often is missing.It would also be useful to determine the concentration of LMM and macromolec-ular organic thiols (using, e.g. S K-edge XANES). Also, total concentrations ofHg and MeHg in the sediment will influence the log Kd. If HgS(s) is present, log


Kd will increase with one order of magnitude if the total concentration increase10 times. As discussed above, salinity (chloride concentrations) may not be afactor affecting log Kd of either Hg or MeHg through chloride complexation.But salting out of Hg-DOM and MeHg-DOM complexes is an important indirecteffect that may contribute to a higher log Kd in saline environments (Babiarzet al., 2001; Turner et al., 2001).

7.8 METHODS FOR STUDYING THE CHEMISTRY OF MERCURY(II)IN SOILS AND SEDIMENTS

A complete chemical characterization of Hg and MeHg in soils and sedimentsrequires methods for the determination of total concentrations, as well as con-centrations of chemical species pertaining to aqueous, solid, surface, and gaseousphases. Total concentrations of mercury (Hg + MeHg + traces of other Hg forms)can be determined after complete digestion of the soil or sediment sample, orin aqueous phase, after reduction of mercury to Hg0 followed by cold vaporatomic fluorescence spectroscopy (CVAFS). In contrast, a separate determinationof MeHg and Hg requires combinations of extraction, liquid–liquid separation,clean-up, derivatization, separation, and detection steps, preferably using stableMeHg and Hg isotopes as internal standard for quality control (Bjorn et al.,2007).

Individual chemical species of Hg and MeHg are much more difficult toquantify. Selective wet chemical extraction has been used to get approximateinformation about Hg species pertaining to solid phases and surfaces (Bloomet al., 2003). With this method, organically bound Hg (extracted by strong acidor base) may be semiquantitatively separated from HgS(s)-dominated phases(dissolved by aqua regia) in soils and sediments. In highly contaminated minewastes also HgCl2, HgO(s), and Hg0(l) phases may be identified and semiquan-titatively determined (by dissolution in deionized water, acetate, and strong acid,respectively). A comparison with Hg LIII-edge EXAFS spectroscopy revealed thatHgS(s) was most accurately determined by extraction, whereas the quantificationof the more soluble Hg-phases such as HgCl2 and HgO(s) were dependent on theoverall matrix, particle size, and crystallinity of the mine waste samples (Kimet al., 2003). In less contaminated samples, such as most soils and sediments, con-centrations are too low to be quantified by Hg LIII-edge EXAFS spectroscopy.However, if the concentration of reduced S groups is in great excess, additionsof Hg (or MeHg) may give an idea of the chemical speciation also at ambientconcentrations. A parallel determination of the binding ligands by, for example,S K-edge XANES is in many cases very useful. Mercury LIII-edge EXAFS isthe only available method for the direct determination of the local binding envi-ronment (molecular structure) of Hg and MeHg in environmental samples. Arecent summary of XAS study conducted on mercury in soils and sediments isgiven in Skyllberg (2010). The more or less distinct release of gaseous Hg as a

FUTURE RESEARCH NEEDS 249

consequence of increased temperature has been used to semiquantitatively sepa-rate Hg0(l), organically bound Hg, and HgS(s) by the thermodesorption method(Biester and Scholz, 1997). This is currently the best available method for thedetermination of Hg0(l) in contaminated sediments.

Because of the very low concentrations and methodological difficulties, thereare currently no methods for a direct determination of dissolved Hg and MeHgspecies in soils and sediment pore water. Thus, the concentration of these speciesneeds to be indirectly calculated from stability constants. Ligand-exchange meth-ods have been successfully used to determine conditional constants for the inter-action between Hg, MeHg, and organic or inorganic sulfur-containing ligands inaqueous and solid phases of soils and sediments (Gasper et al., 2007; Skyllberg,2008).

7.9 FUTURE RESEARCH NEEDS

In order to understand the critical processes in mercury biogeochemistry, suchas methylation, demethylation, gaseous evasion, diffusion, solubilization, mobi-lization, bioaccumulation, and biomagnification, a correct determination of thechemical speciation of Hg and MeHg in aqueous and solid phases of soils,sediments, and waters is required. Major knowledge gaps include molecular inter-actions of Hg and MeHg with not only the most abundant metal sulfides such asFeS2(s), FeS(s) but also the less abundant ZnS(s), CuS(s), and Cu2S(s). The inter-action between MeHg and HgS(s) and FeS(s) is currently unknown. It is likelythat the observed log Kd for MeHg on the order of four to six in some envi-ronments is partly due to these types of interactions. The aging of FeSm, whichis affected by the dynamics of reduction and oxidation and presence of NOM,seems to play a major role. When it comes to aqueous species, the solubilityof β-HgS(s) needs to be studied at S(-II) concentrations below 30 μM. In thesestudies, the problem of separating truly dissolved species from nanoparticulateHgS(s) needs to be addressed. A combination of solubility modeling and inde-pendent spectroscopic studies is required to resolve the stability constants for theformation of various Hg-polysulfides and the HgOHSH0 complex. To improvethe understanding of critical processes such as methylation and demethylation,we need multidisciplinary approaches in which the chemical speciation of Hgand MeHg is linked to the biogeochemistry of organic C, S, and Fe and thebiological activity of FeRB and SRB.

APPENDIX

Acidity and Thermodynamic Constants Used in the Thermodynamic Modelingof Soils and Sediments Reported in Table 7.2 and Figs 7.2, 7.5, 7.8, and 7.9.


Chemical Reaction Log K References

Proton dissociation of ligands

H2O(l) =OH− + H+

13.7

H2S(aq) − HS− +H+

7.0

RSH(aq) =RS− + H+

9.96 Skyllberg (2008)

Formation of inorganic Hg(II) complexes

OH− + Hg2+− =HgOH+

10.3 (I = 0 M) Powell et al. (2004)

2OH− + Hg2+ =Hg(OH)2

0(aq)21.4 (I = 0 M) Powell et al. (2004)

3OH− + Hg2+ =Hg(OH)3

−20.3 (I = 0 M) Powell et al. (2004)

Cl− + Hg2+ =HgCl+

7.3 (I = 0 M) Powell et al. (2004)

2Cl− + Hg2+ =Hg(Cl)2

0(aq)14.0 (I = 0 M) Powell et al. (2004)


−14.9 (I = 0 M) Powell et al. (2004)


2−15.5 (I = 0 M) Powell et al. (2004)

Cl− + OH− +Hg2+ =HgClOH0(aq)

18.0 (I = 0 M) Powell et al. (2004)

HS− + Hg2+ =HgSH+

20.6 (LFER) Dyrssen and Wedborg(1991)

2HS− + Hg2+ =Hg(SH)2

0(aq)

37.7 (I = 1.0 M) Schwarzenbach andWidmer (1963)

2HS- + Hg2+ =HgS2H- + H+

31.5 (I = 1.0 M) Schwarzenbach andWidmer (1963)

2HS- + Hg2+ =HgS2

2- + 2H+23.2 (I = 1.0 M) Schwarzenbach and

Widmer (1963)HS- + OH- + Hg2+

= HgOHSH0(aq)30.3 (LFER) Dyrssen and Wedborg

(1991)HS- + Cl- + Hg2+

= HgClSH0(aq)25.8 (LFER) Dyrssen and Wedborg

(1991)

FUTURE RESEARCH NEEDS 251


2HS− + (n-1)S0 +Hg2+ = HgSnSH-

+ H+

32.6 (I = 0.3 M) Jay et al. (2000)

2HS− + 2(n −1)S0 + Hg2+ =Hg(Sn)2

2− + 2H+

24.8 (I = 0.3 M) Jay et al. (2000)

S0

RS− + Hg2+ =HgSR+


2RS− + Hg2+ =Hg(SR)2

42.0 Skyllberg (2008)

HgOHSR + Cl− =HgClSR + OH−

−3.7 Hilton et al. (1975)

RS− + OH− +Hg2+ = HgOHSR


Formation of MeHg(II) complexes

OH− + CH3Hg+ =CH3HgOH0

9.5 Loux (2007)

Cl− + CH3Hg+ =CH3HgCl0

5.4 Loux (2007)

SH− + CH3Hg+ =CH3HgSH0

14.5 (LFER) Dyrssen and Wedborg(1991), Loux (2007)

SH− + CH3Hg+ =CH3HgS− + H+

7.0 (I = 0.1 M) Schwarzenbach andSchellenberg (1965)

RS− + CH3Hg+ =CH3HgSR

16.5 (I = 0 M) Loux (2007), Karlssonand Skyllberg (2003)

Solubility of solid phasesBlack

HgS(s) + H+ =HS− + Hg2+

−36.8 (I = 1.0 M) Schwarzenbach andWidmer (1963)

Red HgS(s) + H+ =HS− + Hg2+

−36.5 (I = 0.7 M) Paquette and Helz (1997)

(continued)



Formation, dissolution, volatilization, and dismutation of Hg0

Hg2+ + 2e− =Hg0(l)

28.7 Sillen and Martell (1971)

Hg0(l) = Hg0(aq) −6.48 (I = 0 M) Clever et al. (1985)Hg0(g) = Hg0(aq) −2.59 (I = 0 M) Sanemasa (1975)Hg2+ + Hg0(aq) =

Hg22+

8.46 (I = 0 M) Hietanen and Sillen(1956)

The ionic strength is reported for experimentally determined constants. LFER =constant determined by linear free energy relationships. Some constants wererecalculated using the acidity constants for H2S and H2O and for solubilities ofHgS(s) as reported in the original work or in this table.

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