[developments in soil science] synchrotron-based techniques in soils and sediments volume 34 ||...

Developments in Soil Science, Volume 34

C H A P T E R

13
Mercury Biogeochemistry in Soils
and Sediments
U. Skyllberg
Department of Forest Ecology and Management, Swedish University of Agricultural Sciences,Umea, Sweden

O U T L I N E

1. Biogeochemistry of Mercury and theImportance of Chemical SpeciationAnalysis 3801.1. Mercury Contamination, Net Methyl

Mercury Production and Accumulationin Biota 380

1.2. Linked Biogeochemistry of Fe, S, andHg and Need for ChemicalSpeciation Analysis 381

2. Chemical Speciation of Mercury inSoils and Sediments Using XAS 3822.1. The Usefulness of EXAFS

to Differentiate Between O/N andS Ligation 382

2.2. EXAFS Studies on Mercury ComplexFormation with Organic Ligands 383

2.3. EXAFS Studies on MercuryAdsorption to Mineral Surfaces UnderOxidized Conditions 391

2.4. EXAFS Studies on ChemicalSpeciation of Mercury Under Suboxicand Reducing Conditions 395

2.5. Chemical Speciation of Mercury inMine Waste and Downstream Effects 401

2.6. Comparison of Hg EXAFS with WetChemical Extraction and ThermoDesorption 403

2.7. Use of Hg EXAFS to MonitorRemediation of Mercury 405

2.8. Mercury Speciation in OrganismsUsing XANES and EXAFS 405

3. Conclusions 406

379
Copyright # 2010, Elsevier B.V.
All rights reserved.

380 13. MERCURY BIOGEOCHEMISTRY IN SOILS AND SEDIMENTS

1. BIOGEOCHEMISTRY OFMERCURY AND THE IMPORTANCE

OF CHEMICAL SPECIATIONANALYSIS

1.1. Mercury Contamination, NetMethyl Mercury Production andAccumulation in Biota

The unique properties of elemental mercury,Hg(0), being a liquid at room temperature witha high gas pressure, have been both a blessingand a curse for man. The element is easilyextracted from mineral deposits and apart fromits well-known use to concentrate (amalgamate)silver and gold, it has been extremely useful invarious industrial processes. An incautious utili-zation of mercury has, however, resultedin severe contamination of soil, sediments,and water resources in association to mines andindustries all over the world (Hylander andMeili, 2005). Especially the paper and pulpindustry have contributed to local contamina-tion by mercury due to the use of elemental Hgas an electrode in the chlor-alkali process andas a fungicide to preserve pulp-fiber. Many ofthese locally contaminated industrial sites stillhave to be remediated (e.g., Skyllberg et al.,2007).

Historical records in ice cores (Schuster et al.,2002) and peat soil (Biester et al., 2007; Martinez-Cortizas et al., 1999) reveal a global spreadingof elemental Hg due to natural and anthropo-genic activities for centuries. Gaseous, elementalmercury, formed during volcanic eruptions aswell as during combustion of petroleum pro-ducts, coal, and other fossil fuels, has a residencetime in the atmosphere on the order of 1 year(Slemr et al., 2003). After oxidation to inorganicHg(II), mercury is deposited on to oceans andland. The diffusive deposition of long-rangetransportedmercury caused by human activitieshas resulted in an accumulation in soils andsediments exceeding preindustrial concentra-tions by 3-5 times (Biester et al., 2007).

The biogeochemistry of mercury is compli-cated and involves all chemical states; solid,liquid, gaseous, and aqueous phases. In soilsand sediments elemental Hg(0) (l, g) may formunder anoxic conditions, both as a liquid andas a gas. Under oxidized conditions, inorganicHg with oxidation state þII predominate. Inor-ganic Hg(II) forms dissolved complexes withorganic and inorganic ligands, surface com-plexes with particles as well as solid phaseswith mainly sulfides (a- and b-HgS). Mercu-rous Hg, having oxidation state þI, is notstable in natural environments. Of major con-cern is the formation of alkylated forms of Hg(II), commonly designated organic mercury.Even if ethyl mercury (EtHg) and dimethylmercury (DMeHg) have been detected in envi-ronmental samples, monomethyl mercury(CH3Hg or MeHg) is by far the most abundantorganic form of mercury in the environment.Because of its lipofilic character MeHg readilypasses both the placental and the blood-brainbarrier and it is extremely toxic. As opposedto inorganic Hg, MeHg is not easily excretedand therefore bioaccumulates in organisms.As a consequence, almost all mercury in fishis in the form of MeHg (Downs et al., 1998).

Because of the toxicity of MeHg and itsbioaccumulation in higher organisms, anunderstanding of the environmental control ofthe transformation of inorganic Hg to MeHgis considered most important. Only if the meth-ylation process is understood in detail, actioncan be taken to minimize the effect on thesurroundings by locally contaminated sites.Perhaps it is possible to modify also large-scaleactivities in order to decrease the net produc-tion of MeHg at a landscape level. In Swedenalone 40,000 lakes have MeHg concentrationsin fish exceeding levels set by health authoritiesfor regular consumption (Hakanson, 1996), andthere is a current discussion about whetherforestry activities can be modified in order todecrease the net formation and export of MeHg(Bishop et al., 2009).

RSHgSR

kB

kD

kMHgc CH3HgX

SS

Hg

HgS�Hg(SH)2�

Hg2+ + xS2− + yH+ = HgSxHy2–2x+y

e.g HgS�, HgS22−, HgS2H–, Hg(SH)2�

Cell

FIGURE 13.1 A model of the assumed process of Hguptake and methylation within a bacterial cell. The rate ofmethylation is designated as kM; kB is the combined rate ofcompeting reactions that sequester Hg and make unavail-able for methylation and kD is the uptake rate. From Benoitet al. (2003). Printed with permission from AmericanChemical Society.

3811. BIOGEOCHEMISTRY OF MERCURY AND THE IMPORTANCE OF CHEMICAL SPECIATION ANALYSIS

The biogeochemistry and the transformationof inorganic Hg to MeHg has been covered inseveral review articles (e.g., Benoit et al., 2003;Morel et al., 1998; Ullrich et al., 2001). The forma-tion of MeHg is the net result of methylation(production of MeHg) and demethylation (deg-radation of MeHg) reactions. These reactionsare to a large degree mediated by microbes. Sul-fate reducing bacteria (SRB), and more recentlyiron reducing bacteria (FeRB), have been shownto be the two groups of microorganisms respon-sible for MeHg production in soils, sediments,and anoxic waters (Benoit et al., 2003; Fleminget al., 2006; Kerin et al., 2006; Ullrich et al.,2001). Demethylation processes may be abioticor biotic, and the latter seem to be mediated bya broad range of microorganisms. Some organ-isms utilize the energy in MeHg during oxida-tive demethylation, whereas other organismsinvest energy in a reductive demethylation reac-tion as a detoxification process. In general, themethylation process is most active under sub-and anoxic conditions, whereas demethylationreactions are generally promoted by oxic condi-tions. Therefore, sediments in lakes having oxy-gen deficient (hypolimnetic) bottom waters andsoils subjected to seasonal inundation, or havinga permanently high water table such as varioustypes of wetlands and discharge areas close tostreams and surface waters, are major sourcesof MeHg (Eckley et al., 2005; Hurley et al., 1995;St. Louis et al., 1994). In addition to oxygendeficiency, SRB require sulfate and FeRB requireFe(III) as terminal electron acceptors, and bothgroups of bacteria require energy rich organicsubstrate for their energy metabolism.

1.2. Linked Biogeochemistry of Fe, S,and Hg and Need for ChemicalSpeciation Analysis

Laboratory and field studies have indicatedthat a passive uptake of neutral inorganicHg–S molecules (e.g., Hg(HS)2

0) by SRB is onepossible mechanism behind MeHg production

in soil and sediments (Benoit, 1999a,b; Drottet al., 2007; Mehrotra and Sedlak, 2005). Thisprocess is depicted in Fig. 13.1. Laboratorystudies have shown that in addition to passiveuptake of neutral Hg species, an active uptakeof Hg complexed by small organic moleculesalso is possible (Golding et al., 2002). Smallorganic molecules such as thiols (RSH) maysignificantly facilitate the uptake of Hg byFeRB, as recently shown by Schaefer and Morel(2009). Thus, the chemical forms of Hg avail-able in pore waters and bottom waters may beextremely important for MeHg production.Similarly, it has been suggested that activeand passive uptake of MeHg complexes withHS� by various types of bacteria may be ofimportance for the demethylation process(Benoit et al., 2003; Drott et al., 2008).

Concentrations of the dominant chemicalforms of Hg and MeHg in pore waters of soilsand sediments are on the order of approxi-mately 10�13-10�10 M (e.g., Drott et al., 2007;Skyllberg et al., 2003). These concentrations


are too low to be determined directly by analyti-cal methods, and the only way to determine thechemical speciation is by modeling. Chemicalmodeling in turn requires accurate input ofparameters and data. Besides measurements oforganic and inorganic ions and molecules insolution, a proper description of the solid phaseof soils and sediments is required. Because the“soft” chemical properties of Hg and MeHg(both are classified as type B metals), they willassociate very strongly with halides (Cl� < Br�

< I�) and reduced S compounds like sulfides(S2�), bisulfides (HS�), polysulfides (Sn), andorganic thiols (RSH). Because of generally lowconcentrations of Br� and I�, and the muchstronger association to reduced S than to Cl�,both Hg and MeHg mainly associate withreduced S compounds. This also holds trueunder marine conditions (e.g., Dyrssen andWedborg, 1991).

Under oxidized condition organic thiolspertaining to low and high molecular weightnatural organic matter (NOM) in solution aswell as at particle surfaces control the chemicalspeciation of Hg and MeHg. Under reducingconditions metacinnabar (b-HgS) is the mostlikely solid phase of mercury formed in soilsand sediments (Barnett et al., 1997). Cinnabar(a-HgS) may also be present, but mostly as arest product in mine tailings. Substitution ofFe(II) into HgS(s) or coprecipitation of Hg andFe to mixed HgFeS (s) phases is a possibilityin sediments (Jeong and Hayes, 2003;Wolfenden et al., 2005). It is also possible thatHg2þ (as well as MeHgþ) is adsorbed to surfacefunctional groups of FeS(s) particles (e.g.,Wolfenden et al., 2005). Furthermore, the reac-tion between Fe(II) and S(-II) indirectly regulatesHg chemistry by controlling the concentration ofdissolved HS�. A recent review and evaluationof stability constants for the association of Hgand MeHg to organic and inorganic sulfurligands is given by Skyllberg (2008).

To summarize, the fact that SRB and FeRBare recognized as the main microbes

responsible for the Hg methylation processes,their transformation of S(VI) ! (S-II) and Fe(III) ! Fe(II) combined with the strong chemi-cal interactions between chemical forms of Hg,S, and Fe means that the biogeochemistry ofthese three elements are tightly linked in a veryintricate way. To understand this complex sys-tem, proper determination of the chemical spe-ciation of Hg, S, and Fe is needed, and here X-ray absorption spectroscopy (XAS) techniquesplay an important role. In this chapter, the con-tribution of XAS to our current understandingof the speciation of Hg in soils and sedimentsis summarized. Geological deposits are brieflycovered by a chapter on the effects of minewastes.

2. CHEMICAL SPECIATION OFMERCURY IN SOILS ANDSEDIMENTS USING XAS

Brown and Surchio (2002) gave an overviewof XAS studies on metal (and metalloid) sorp-tion complexes at mineral/solution interfaces.In total four studies were listed for mercury.Six additional XAS studies on mercury minewastes were also covered in the review.Andrews (2006) summarized XAS studies onmercury speciation in environmental sampleslike coal, mine tailings, clays, NOM, plants,and fish. In recent years, the body of work usingXAS to study mercury in model systems, as wellas in natural soil and sediment samples, hasgrown considerably.

2.1. The Usefulness of EXAFSto Differentiate Between O/N andS Ligation

Several factors make Hg EXAFS an ideal toolfor unraveling the chemical speciation of mer-cury in complex systems like soils and sedi-ments. Because of its “soft” character and its

3832. CHEMICAL SPECIATION OF MERCURY IN SOILS AND SEDIMENTS USING XAS

electron configuration, Hg forms predominantly2-, 3-, or 4-coordinated covalent compoundswith sulfur. In excess of available sulfur ligands,Hg forms similarly low-coordinated complexeswith O and N ligands. In 2-coordinated com-pounds, the average bond length of for Hg–C(in organometallic compounds), Hg–O, andHg–N may increase with the atomic number(ion radius) of the ligand atom (Table 13.1), butthe differences are too small to be dis-tinguishable by a first-shell EXAFS analysis.However, backscattering of the photoelectronby S and C/O/N atoms is easily distinguished(owing to substantial differences in atomicmass), enabling a determination of the contri-bution form Hg–S and Hg–O/N associationsalso in very complex matrices like soils. Further-more, because of the sensitivity of the EXAFSfrequency to the bond length, 2-, 3-, and 4-coordinated compounds can be separated by afirst-shell analysis. The typical error for the bondlength is on the order of �0.02 A (the error ofcoordination numbers (CN) is on the orderof �20%). Thus, the error is smaller than the

TABLE 13.1 Typical Hg–S distances in 2-, 3-, and 4-coreported Hg–S, Hg–O, Hg–N, and Hg–C dis

2-Coordinate (A) 3-Coordin

Hg–S with RSHsa 2.345 � 0.025b 2.446 �a-HgS(s) 2.36-2.39

b-HgS(s)

Hg(H2O)6 2þ (aq)

HgO (s) 2.02-2.10

Hg–N 2.05-2.16e

Hg–C 2.01-2.08f

aManceau and Nagy (2008).bn ¼ 48.cn ¼ 11.dn ¼ 103.eHolloway and Melnik (1995).fQian et al. (2002), Hg EXAFS data determined for CH3Hg.gCharnock et al. (2003), Hg EXAFS data.hCollins et al. (1999), density functional calculations.

average increase in Hg–S bond lengths from2.34 to 2.44 A, when a third S ligand is added,and from 2.44 to 2.56 when a fourth S ligandis added (Manceau and Nagy, 2008). InTable 13.1 typical distances for Hg–C, Hg–O,Hg–N, and Hg–S bonds are tabulated for someselected model compounds.

2.2. EXAFS Studies on MercuryComplex Formation with Organic Ligands

2.2.1. Concentrations of Low MolecularWeightThiols in Soils, Sediments andWaters

In a series of recent studies, concentrationsof low molecular weight (LMW) organic mole-cules with reduced S functional groups havebeen determined in natural waters (Han et al.,2006; Tang et al., 2003; Zang et al., 2004). Inmost of these studies a reduction step wasused, which means that organic sulfides (RSR,RSSR) cannot be separated from thiols (RSH).The sum of thiols and organic sulfides has beenshown to be on the order of about 1-100 nM in

ordinated thiol compounds as well as typical ranges oftances in some selected model compounds

ate (A) 4-Coordinate (A) 6-Coordinate (A)

0.018c 2.566 � 0.047d

2.50-2.55g

2.34-2.40h


both fresh waters and marine ecosystems. It isonly in the study of Hu et al. (2006) in whichtrue thiols, having RSH groups available forcomplexation of Hg, have been determined innatural systems. Under highly reducingconditions in a stratified lake with hypolimnetic(oxygen free) bottom waters up to 340 nM of3-mercaptopropionic acid was determined.Under oxic conditions the concentrations wereon the order of 2-3 nM. Glutathione, which isthe most important thiol in living cells, wasmeasured at concentrations on the order of 1-10nM in the lakewater (Hu et al., 2006). Thus, evenif the number of studies so far is low, we canexpect that free thiols under sub- and anoxicconditions may reach concentrations on theorder of about 10 nM or more.

2.2.2. Complexation of Hg with LowMolecular Weight Thiols

Although Hg-thiol complexes have been thefocus in a number of EXAFS studies, two recentstudies on L-cysteine (Jalilehvand et al., 2006)and D-penicillamine (Leung et al., 2007) havegenerated the most detailed information on

4

A B

6 8 10k (Å–1) (R-a

12 14 16 0 1 2

S-Hg-SS-Hg-S

Hg-S

FT

Mag

nitu

de

c(k

).k3

Hg-S

FIGURE 13.2 Hg LIII-edge EXAFS data (solid lines) and fits(A) Full data (top), Hg–S single scattering path (middle), S–Hg–Stion (data not corrected for phase shift). The core of the proposerepresents C(CH3)2CH(NH3

þ)COO� [Leung et al. (2007)]. Repr

ligands relevant for natural conditions in soilsand sediments. In both these studies the com-plexation of Hg was studied using a combina-tion of Hg LIII-edge EXAFS and 199Hg-NMRspectroscopy. The EXAFS data were collectedat room temperature both for solid Hg-thiol pre-cipitates (at pH 3.7) and for aqueous solutionswith dissolved Hg-thiol complexes (at pH 11.0).The amplitude reduction factor S0

2 was fixed to0.9 and the bond distance R, the Debye-wallerfactor (s2), CNs, and the edge energy (DE0)were allowed to float during fitting. The s2 was0.004-0.007 A�2 in the first coordination shell.

Similar results were observed for both thiolsstudied and the results for penicillamine areillustrated in Figs. 13.2 and 13.3. Penicillamineand cysteine have three pKa values: pKa1� 2,pKa2� 8, and pKa3� 10. The first constant isthe acidity of the carboxyl group, whereas thesecond and third constants can be consideredaverage macroscopic constants for the aciditiesof the amino and thiol groups. The solid pre-cipitate (Fig. 13.2) was described by alinear structure with two thiol-S atoms asso-ciated to a central Hg atom. The Hg–S bond

) (Å)

2.35 Å

2.35 Å

3 4 5

R

R

S SHg

(dashed lines) for the Hg(HPen)2 solid compound at pH 3.7.multiple scattering path (bottom). (B) Radial structural func-d structure of Hg(Pen)2 is illustrated to the right. The letter Roduced by permission of the Royal Society of Chemistry.

4

A B

6 8 10k (Å–1) (R-a) (Å)

2.52 Å

2.34 Å 2.34 Å12 14 0 1 2

S-Hg-SS-Hg-S

Hg-S

Hg-NHg-N

Hg-CHg-C

FT

Mag

nitu

de

c(k

).k3

Hg-S

3 4 5

(–OOC)HCNH2

(CH3)2C

Hg SRS

FIGURE 13.3 Hg LIII-edge EXAFS data (solid lines) and fits (dashed lines) for a Hg(II)-penicillamine solution with [Hg(Pen)2]

2� as the dominating complex at pH 11. (A) Full data (top), Hg–S single scattering path, Hg–N single scattering path,Hg–C single scattering path, and S–Hg–S multiple scattering path (bottom). (B) Radial structural function (data not cor-rected for phase shift). The proposed structure of [Hg(Pen)2]

2� is illustrated to the right. For simplicity only one of thetwo rings are shown (R denotes C(CH3)2CH(NH2)COO�, making up the second ring) [Leung et al. (2007)]. Reproducedby permission of the Royal Society of Chemistry.


distance was 2.35 A, which is in agreementwith a 2-coordination. The multiple scattering(MS) peak occurring at �4 A in Fig. 13.2(uncorrected for phase shift) corresponds to atrue distance of 4.64 A (�2 � 2.35 A) and aCN of 2.1. Even if the CN typically is subjectedto an error of approximately �20%, the rela-tively strong MS peak (owing to an amplifica-tion of the backscattered X-rays along acollinear four-legged path Hg ! S1 ! Hg !S2 ! Hg and a weaker three-legged pathHg ! S1 ! S2 ! Hg) give further supportfor the linear form of the S–Hg–S complex.Thus, at pH 3.7 a Hg(HPen)2 complex was con-cluded to be the highly dominant complex,with the amino group remaining protonated.

In the aqueous solution experiments, themolar ratio of thiol to Hg was varied. At pH11 all three functional groups are dissociatedand the ligand is designated Pen2�. At a thiolto Hg molar ratio of 1.9 the Hg(Pen)2

2�waspredominating (Fig. 13.3). The dissociationof the amino group at pH 11 resulted in a

five-membered ring formation, with two thiol-Satoms at 2.34, two amino-N atoms at 2.52 Aand four C atoms at 3.24 A. With increasingligand to Hg ratio a mixture of three complexes;Hg(Pen)2

2�, Hg(Pen)34�, and Hg(Pen)4

6� isformed. This was shown as an increase in theaverage bond distance between Hg and the firstshell S atom, in agreement with the values tabu-lated for 2-, 3-, and 4-coordinated Hg–S com-pounds in Table 13.1. By use of a least-square,linear combination fitting of EXAFS spectra, therelative mixtures of Hg(Pen)2

2�, Hg(Pen)34�,

and Hg(Pen)46� could be calculated. The results

were largely in agreement with previous specia-tion calculations for penicillamine (Koszegi-Szalai and Paal, 1999), Fig. 13.4, where Ldesignates the Pen2� ligand.

As can be seen in Fig. 13.4, the speciation ofHg in presence of LMW thiols like penicilla-mine and cysteine (showing a similar pH-dependency) is highly dependent on pH andthe ligand to Hg molar ratio. According to theEXAFS studies conducted by Leung et al. at

00

0.2

0.4

0.6

0.8

1

1.2

1 2 3 4 5 6 7 8

pHA

Frac

tiona

l con

cent

ratio

n

HgL2H42+

HgL2H2

HgL2H3+

HgL3H3–

HgL2H–

00

0.2

0.4

0.6

0.8

1

1.2

1 2 3 4 5 6 7 8pHB

Frac

tiona

l con

cent

ratio

n

HgL2H42+

HgL2H2

HgL2H3+

HgL3H3–

HgL2H–

FIGURE 13.4 The pH dependent distribution of Hg(II)-penicillamine species at total uncomplexed penicillamine con-centrations of 5 � 10�3 M (A) and 0.4 M (B). L denotes the Pen2� ligand. Reprinted from Koszegi-Szalai and Paal (1999),with permission from Elsevier.


pH 11, the Hg(Pen)34� complex dominated

over Hg(Pen)22� at a thiol to Hg molar ratio

exceeding 5. The Hg(Pen)46� complex showed

only a small tendency to form and at maximumcontributed with 14% at a Hg to thiol molar ratioof 15.4. Cysteine showed a stronger tendency toform a 4-coordinated complex than the largerpenicillaminemolecule (likely due to sterical hin-drance), and the Hg(Cys)4

6� dominated at a thioltoHgmolar ratio exceeding 5.3 (Jalilehvand et al.,2006).

Given that the concentration of Hg in soilsand sediments commonly is on the order of0.01-0.1 nM and concentrations of free thiolsunder suboxic and reducing conditions mayreach 10 nM or even more in surface waters, amixture of Hg coordinated by 2 and 3 thiolsmay be possible in neutral or alkaline environ-ments. Under acidic conditions, however, the2-coordinated form should predominate. Innatural environment there will be a competi-tion between LMW and HMW thiols, and ifthe latter predominate (see the next section) ste-rical hindrance may favor 2-coordination.Other factors that need to be considered is com-petition for binding sites by Zn and Cu, whichshow weaker bonding to thiols as comparedto Hg, but these metals commonly occur atmuch higher concentrations.

2.2.3. Complexation of Hg in NaturalOrganic Matter

Mixed S and O/N Ligation in NOM. Xiaet al. (1999) reported the first Hg EXAFS studyon the complexation of Hg by NOM. Humicsubstances were gently extracted (by removalof flocculation polyvalent ions using a cation-exchange resin) from an organic soil situatedin the zone of soil water discharge between aforested upland and a lowland fen. The concen-tration of reduced sulfur groups (the sum ofthiols RSH, sulfide RSR, and disulfide RSSR)was determined using sulfur K-edge XANES.The addition of Hg-nitrate resulted in a Hg(II)to reduced S molar ratio of 3:1. The pH wasbelow 5 after addition of Hg. Thus, the suppo-sedly strongest sites; the thiol groups, weremore than saturated by Hg. As a result, oneoxygen (or nitrogen which cannot be separatedfrom O using conventional EXAFS) and onesulfur atom were observed in the first coordi-nation shell. Distances of 2.02 A for Hg–O and2.38 A for Hg–S bonds strongly suggested a 2-coordination of Hg. Thus, the first shell indi-cates that Hg is bound to thiolate groups(RS�) and to carbonyl (R ¼ O), carboxylate(R–O�), or amino (RNH2) groups. The secondcoordination shell data could be fitted bymodels having one S (at 2.93 A) and one C


(at 2.78 A) or twoC atoms at 2.78 A. The relativelyshort Hg–C distance indicates that these carbonatoms are directly bound to the first shell O(or N) atoms. A second shell C bound to athiol (RSH) group would be situated about3.3 A away from the Hg atom. The second shellHg–S interaction was suggested to be causedby an involvement of RSSR or RSSH functionalgroups.

Given results of EXAFS studies of aminoacids having O, N, and S functionalities(Jalilehvand et al., 2006; Leung et al., 2007), itis expected that thiol groups form the strongestbonds to Hg also in NOM. However, becauseof excessive addition of Hg in relation toreduced S groups, the possibility to identifydefinite structures involving only thiols andother reduced S functionalities was limited inthe study of Xia et al.(1999). Hesterberg et al.(2001) varied the molar ratio of Hg to reducedS between 2.4 and 0.25 (i.e., a maximum of fourreduced S groups were available for eachHg atom) in their EXAFS study of extractedhumic acids from a marine wetland soil. ThepH was 5.6. A curve-linear relationship wasestablished showing an increasing fraction ofHg–S bonding with decreasing Hg to reducedS ratio. At a molar ratio of 0.25, reduced Satoms contributed with 90% and O/N atomswith 10% of the first shell CN. Thus, the studyof Hesterberg et al. (2001) gave additional sup-port for a twofold coordination between Hgand S and O/N ligands. Since EXAFS data givethe average coordination chemistry, neither Xiaet al. (1999) nor Hesterberg et al. (2002)couldunequivocally distinguished a mixed ligationof S and O/N ligands from complexes in whichHg is exclusively bound to reduced S groups,or exclusively to O/N groups.Evidence for a Linear, Twofold Coordinationwith Thiols in NOM. Using the undulatorbeamline ID26 at the third-generation SRsource at ESRF, Grenoble, Skyllberg et al.(2006) could lower the Hg to reduced S molarratio to 0.01 (i.e., representing a theoretical Hg

saturation of only 1% of the reduced S groups).The coordination chemistry of Hg was deter-mined at room temperature in both intactorganic soils (pH 3.8-4.0) and in extracted humicacid from a wetland soil (pH 3-3.5). During fit-ting the parameters S0

2 (0.90) and s2 (0.001-0.01A2) were in fair agreement with EXAFS analysisof LMW thiols (see above). The concentration ofreduced S (RSHþ RSRþ RSSR) was determinedusing S K-edge XANES. Evidence for a lineartwofold coordination with two thiol groupscould be presented at Hg to reduced S molarratios of 0.01 and 0.05 (Fig. 13.5). The first shellHg–S bond distance was 2.33, well in agreementwith distances reported for thiols like penicilla-mine and cysteine. Furthermore, the strong MSpeak occurring at �3.9 A (corresponding to4.6 A after phase shift correction) is a directevidence of a linear S–Hg–S configuration. Thefour-legged path Hg ! S1 ! Hg! S2 ! Hgresults in a modeled distance two times theHg–S distance of 2.33 A. In the sample having aHg to reduced S molar ratio of 0.01, the MS pathwas modeled by a CN of 2.0, suggesting close to180� angle of the S–Hg–S structure.

Despite the low Hg to reduced S molar ratio,the high brilliance of the beamline at ESRFprovided a good data quality for the secondcoordination shell. Two C atoms were detectedat a distance of in average 3.35 A. This is well inagreement with data for LMW thiols. A secondshell S contribution, as suggested by Xia et al.(1999), was also seen in the study of Skyllberget al. (2006), but only in the organic soil sam-ples reaching maximum 5% theoretical Hg sat-uration of reduced S functionalities. The bestmodel included 1.3 S atoms at a distance of3.08 A in the sample with an Hg to reduced Smolar ratio of 0.01, and 0.5 S atoms at 2.92 Adistance from the Hg atom in the sample witha molar ratio of 0.05. By inclusion of the longHg–S distance, model fits were improved by15-20%, as compared to a model having onlysecond shell Hg–C contributions. The thirdS atom could theoretically be a thiol but an

S

3rd Peak Parker-thread data

2nd Peak Parker-thread data

3 5 7 92nd Peak Parker-thread data

S

S/C

0 2 4 6 8 0 2 4 6 8R (Å) R (Å)

S/C

S-Hg-S

S-Hg-Sc

(k)×

k3

c(k

)×k

3

k(Å–1

) 3 5 7 9 11

k(Å–1

)

Hg/Org-SRED = 0.01 Hg/Org-SRED = 0.05

Inte

nsity

Inte

nsity

FIGURE 13.5 Radial structure functions for experimental data (solid line) and fits (dotted line) for an organic soil with aHg/Org-SRED molar ratio of 0.01 (left) and a Hg/Org-SRED molar ratio of 0.05 (right). The inserts show back transforms(solid line) and fits (dotted line) of the S/C peak (second peak) and the S–Hg–S multiple scattering peak (third peak), left,and the S/C peak, right. Data are not corrected for phase shift. From Skyllberg et al. (2006). Printed with permission fromAmerican Chemical Society.


organic sulfide or a disulfide is more likely,owing to their much weaker attraction to Hg.Second shell Hg–S interactions at 2.96-3.09 Ahave been shown for methionine (Church et al.,1986; Klemens et al., 1989), in combination withfirst shell Hg–S distances of 2.32-2.38 A.

In agreement with Xia et al. (1999) andHesterberg et al. (2001), O/N atoms increasinglycontributed to the first shell coordination as theHg to reduced S molar ratio was increasedbeyond the point of saturation. Second shell datasuggested that five-membered ring structures

R C

1.73

Å

2.33 Å

2.33 Å

1.73

Å

3.35 Å

3.35 Å

2.95

Å

111°

111°C R

Hg SS

S RA B

FIGURE 13.6 Proposed average structures for Hg comple0.01-0.05 in Fig. 13.5 (A) and an average structure for the orgaof 0.10-0.40 (B). Note that distances and angles reflect an averbe mixtures of different ligations. The N-atom represents eith(2006). Printed with permission from American Chemical Soci

with a combination of amino, carboxyls, andpossibly carbonyl groups may form (Fig. 13.6).Similar structures have been determined forwell-defined LMW weight organic molecules(see Skyllberg et al., 2006 for references). Itshould be noted that while Fig. 13.6A mayrepresent a true structure, the structure inFig. 13.6B is an average picture of a mixture ofdifferent Hg-complexes with pure or a mixedcomposition of S and O/N ligands.

Also binding affinity studies conducted at alow Hg(II) to reduced S ratio have reported

1.46

Å2.07 Å

C

N Hg

C

C O

O

S

R

R

1.53 Å

1.30 Å

3.24

Å

3.05 Å 2.33 Å

2.83

Å

3.35 Å

1.73

Å111°

xed by the organic soil with Hg/Org-SRED molar ratio ofnic soil and a humic acid with a Hg/Org-SRED molar ratioage composition and that true complexes may or may noter amino-N or a carbonyl-O group. From Skyllberg et al.ety.


conditional constants in favor of Hg-thiol asso-ciations in NOM from soils (Khwaja et al., 2006;Skyllberg et al., 2000) as well as from naturalfreshwaters (Gasper et al., 2007). The muchweaker affinity for oxygen groups, as well asfor a mixed S þ O ligation, can be illustratedby comparing the stability constant for the for-mation of HOHgOH (log K ¼ 22.2) withHOHgSH (log K ¼ 30.3) and HSHgSH (log K¼ 37.7, Dyrssen and Wedborg, 1991). Becauseof similar stoichiometry, these three constantsmay be directly compared, and the increaseby 7-8 log units when each O ligand isexchanged for an S ligand reflects a largeincrease in binding affinity.

The concentration of Hg was 100 mg kg�1

(expressed in relation to dry mass of NOM) atthe lowest addition of Hg in the EXAFS studyof Skyllberg et al. (2006). This value is �20-500times larger than concentrations of Hg in forestand wetland soils and sediments not affected bypoint sources of Hg [0.2-5 mg kg�1 (dry massNOM)]. In sediments highly contaminated by,for example, by chlor-alkali industry, concentra-tions on the order of 10-100 mg Hg kg�1 is notuncommon. Furthermore, in minerogenic soilsand sediments having on the order of 1% or lessorganic matter, the concentration of Hg may beexclusively bound to NOM and reach concentra-tions of 1-10 mg kg�1 (OM) also in moderatelycontaminated environments. Thus, it can be con-cluded that Hg EXAFS results, complemented bybinding affinity studies conducted at very lowHg concentrations, strongly suggest that Hg(II)mainly is 2-coordinated with thiol groups underoxic conditions in soils and sediments with OMexceeding approximately 1% and with low ormoderate mercury contamination.

2.2.4. Complexation of MeHg with NaturalOrganic Matter Thiols

MeHg is the mercury form that accumulatesmost effectively in higher organisms. In order tounderstand the processes of MeHg production,

degradation, and bioaccumulation, knowledgeabout the chemical speciation of MeHg in soils,sediments, and waters is of utmost importance.There are two studies in which the MeHgspeciation has been determined in organic mat-ter from organic soils and streams using HgLIII-edge EXAFS spectroscopic analysis (Qianet al., 2002; Yoon et al., 2005).

Qian et al.(2002) studied the complexation ofMeHg in an organic soil (the same soil wasused by Skyllberg et al., 2006 for studies ofHg complexation), in potentially soluble NOMfrom the soil as well as in samples from ahumic stream draining the soil. MeHg wasadded as dissolved hydroxide and after 7 daysof equilibration the samples were freeze-dried.pH was 3.8-4.0 in all samples. Cation-exchangeresins with carboxyl and thiol functionalities(added MeHg) were used as model com-pounds. The Fourier transform of the thiolresin model compound demonstrates twomajor peaks or shoulders, one maximum atabout �1.55 A (not corrected for phase shift)which corresponds to the Hg–C distance of2.04 A within the MeHg molecule, and onemaximum at �1.90 A which corresponds tothe Hg–S distance of 2.33 A (Fig. 13.7). Thecarboxyl resin shows only one broad peak com-posed of the sum of C and O backscattering.Thus, the bond lengths of Hg–C (2.04 A) andHg–O (2.09 A) are too close to be separated.The shoulder at �1.90 A representing S atomsin the first coordination shell gradually disap-pears as the molar ratio of MeHg to reduced Sgroups (Org-SRED) increases from 0.01-1.41 inthe PSOS sample. Model fits of the EXAFS datain k-space revealed that MeHg was complexedwith one thiol group, at and below a MeHg toOrg-SRED molar ratio of 0.24-0.37 in the NOMsamples investigated by Qian et al. (2002).Similar results were reported by Yoon et al.(2005) for terrestrial and aquatic humic sub-stances. Second coordination shell analysis inthe latter study further emphasized that

O

S

C

PSOS

CH3Hg(II) in carboxylic resin

CH3Hg(II)/Org-SRED = 1.41






CH3Hg(II) in thiol resin

Hgs (cinnabar)

0 2 4

R (Å)

6 8

PSOS

PSOS

PSOS

PSOS

PSOS

FPOS

CH3Hg(II) in carboxylic resin







CH3Hg(II) in thiol resin

K (Å–1)

2 4 6 8 10 12 14

Hgs (cinnabar)

PSOS

PSOS

PSOSInte

nsity

X(K

) x

K3

PSOS

FSOS

FIGURE 13.7 Radial structure functions (left, not corrected for phase shift) and k3-weighted Hg LIII-edge EXAFS data (right; solid line—experi-mental data, dotted line—fits) for potentially soluble organic substances (PSOS) extracted from an organic soil at pH 6 using a cation-exchange resin,as well as for the model compounds HgS and MeHg added to thiol and carboxyl resins. Reprinted from Qian et al. (2002), with permission fromElsevier.

390

13.MERCURYBIO

GEOCHEMISTRYIN

SOILSAND

SEDIM

ENTS


organic sulfides (RSR, RSSR) were not involvedin the complexation.

Given the much stronger binding affinitybetween MeHg and thiols than between MeHgand amine or oxygen functional groups (Karls-son and Skyllberg, 2003), and the small MeHgto thiol molar ratio in natural environments,the EXAFS results strongly suggest that MeHgis complex by one thiol group in a linear C–Hg–S structure under oxidized conditions insoils and sediments.

2.2.5. Using MeHg as a Probe for Thiols byCombining Hg EXAFS and S XANES

Sulfur K-edge XANES alone is not sensitiveenough to give a reliable separation and quan-tification of RSH, RSR, and RSnR functional-ities. Therefore, these functional groups oftenare reported as a sum of reduced organic S:Org-SRED. The strong affinity for thiol groupsin a monodentate complex makes MeHg a veryuseful probe for these sites. A successive titra-tion of thiols and the weaker O/N functional-ities by MeHg monitored by Hg EXAFS couldbe used to calculate the concentration of thiols[RSH] in mol kg�1 by Equation (13.1). The coor-dination numbers of S and O/N ligands aredesignated CNS and CNO/N, respectively, and[ ] is concentration in, e.g. mol kg�1.

½RSH� ¼ CNS=ðCNS þ CNO=NÞ � ½MeHg�ð13:1Þ

The calculated thiol concentration can in turn becompared with Org-SRED; reduced S having anelectronic oxidation state of 0.2 (correspondingto an absorption maximum at 0.3-0.4 eV aboveelemental S) as determined by S XANES spectro-scopic analysis. In Table 13.2, data for such ananalysis is reported for data from Qian et al.(2002). In average, thiols contributed with 37%of Org-SRED in the organic soil, 32% in the PSOSsoluble organic substances from the organic soiland with 24% of Org-SRED in the humic stream.

A similar type of calculation can be donewith Hg2þ used as a probe, and if we assumethat Hg saturate all thiol groups before O/Nfunctionalities come into play (this is somewhatan oversimplification since thiols may to someextent combine with O/N functionalities inthe complexation of Hg2þ). Because of the diva-lent charge of Hg2þ Equation (13.1) has to bemodified to Equation (13.2):

½RSH� ¼ CNS=ðCNS þ CNO=NÞ � 2½Hg� ð13:2ÞUsing this equation on data reported by Sky-llberg et al. (2006), in average 20% of Org-SREDwas represented by thiols in the organic soiland 29% in a humic acid extracted from a wet-land soil. Thus, as an overall estimate, using datafrom both MeHg and Hg titrations, the concen-tration of thiol groups in relation to all reducedorganic S functionalities (RSH, RSR, and RSSRand RSnR) may be on the order of 20-30% inNOM from organic soils and streams. Thefinding of Skyllberg et al. (2006) that Hg(II) iscomplexed by two thiols at a Hg(II)/Org-SREDmolar ratio of 0.05, furthermore suggests that atleast 10% of Org-SRED (0.05 � 2 � 100) are thiolsthat are mobile or situated close enough to com-bine in a linear twofold coordination with Hg(II).

2.3. EXAFS Studies on MercuryAdsorption to Mineral Surfaces UnderOxidized Conditions

2.3.1. Iron and Aluminum Oxyhydroxides

In highly contaminated soils with low con-centrations of available thiols, oxygen groupsat mineral surfaces may contribute to the com-plexation of Hg(II) under oxidized conditions.It should, however, be noted that even at a con-centration of 1% NOM in mineral soils or sedi-ments the concentration of thiol groups in mostcases is more than enough to complex all Hg.Thus, the situation with Hg(II) complexeddirectly to mineral surface groups under

TABLE 13.2 Calculation of concentrations of thiol groups (RSH) in an organic soil (OS), potentially solubleorganic substances (PSOS) from the organic soil and in stream organic substances (SOS) from the nearby draining

stream

Sample

[CH3Hg]

(mmol

kg�1)

Org-STOT

(mmol

kg�1)

Org-SRED(mmol

kg�1) CNS CNO/N

([RSH]a)

Equation (13.1)

(mmol kg�1)

[RSH]/

Org-STOT

[RSH]/

Org-SRED

OS (0.53)b 31 128 59 0.83 0.54 19 0.15 0.33

OS (1.08) 64 128 59 0.66 0.88 27 0.21 0.46

OS (1.62) 96 128 59 0.27 1.18 18 0.14 0.31

OSaverage 0.17 � 0.04 0.37 � 0.08

PSOS (0.47) 32 162 69 0.65 0.56 17 0.11 0.26

PSOS (0.95) 65 162 69 0.53 0.72 28 0.17 0.40

PSOS (1.41) 97 162 69 0.30 1.10 21 0.13 0.30

PSOSaverage 0.14 � 0.03 0.32 � 0.07

SOS (0.47) 23 131 50 0.85 0.60 13 0.10 0.27

SOS (1.19) 59 131 50 0.24 1.09 11 0.08 0.22

SOSaverage 0.09 � 0.01 0.24 � 0.04

aThe concentration of RSH was calculated by Equation (13.1), using coordination numbers for sulfur (CNS) and oxygen/nitrogen (CNO/N) as determined by Hg LIII-edge EXAFS with MeHg as a probe for RSHs.bNumbers given in parenthesis are the MeHg/Org-SRED molar ratio for each sample. This table was adapted from Skyllberget al. (2005) and reproduced by permission of the Royal Swedish Academy of Sciences.


oxidized conditions may be very uncommon.Indirect adsorption of Hg-NOM complexes tomineral surfaces, on the other hand, is moreof a rule. Mixed organo-mineral colloides areoften responsible for transport and resuspen-sion of Hg(II) in streams and water bodies.

Kim et al. (2004b,c) studied the complexationof Hg(II) to surfaces of goethite (a-FeOOH), g-alumina (Al2O3), and bayerite [b-Al(OH)3] inthe pH range 4-8. Mercury was added as Hg(II) nitrate and equilibrated with the mineralsurface. Excess Hg in the supernatant wasremoved prior to EXAFS experiments. Hg LIII-EXAFS fluorescence data were collected atroom temperature. The final surface coverageof Hg at the goethite surface was 0.39-0.42 mmolm�2, corresponding to 7300 mg Hg kg�1 (drymass). At bayerite, the surface coverage wassimilar (0.39-0.44 mmol m�2) but the much

smaller surface area resulted in a final Hg con-centration of 750 mg Hg kg�1 (dry mass). Theadsorption to g-alumina was much less than togoethite and bayerite and the experiment wascomplicated by a conversion of the surface to abayerite-phase during the experiment. Further-more, the weak bonding of Hg to g-aluminaresulted in an artificial reduction of Hg(II) toHg(I) by the high-intensity SR. This so-called“radiation damage” has not been reported inmany Hg-studies, but using a high-intensitybeamline, Qian et al. (2002) observed a “radia-tion damage” for Hg(II) complexed by a carbox-ylic resin. Similar to the experience of Kim et al.,“radiation damage” was facilitated by relativelyweak Hg–O bonding.

At the goethite surface, Hg(II) was adsorbedas an inner-sphere, bidentate complex in acorner-sharing arrangement to the Fe(O,OH)6

12

2

Fe

OG= 0.39μ mol/m2

G= 0.40

G= 0.42

G= 0.42

1.5

1

0.5

10A

pH 7.4

pH 6.7

pH 5.9

Tran

sfor

m M

agni

tude

pH 4.3

B

C

D

8

6

4

2

0

0 2 4 6 8 10 12 0 1 2 3 4 5 6

K (Å–1)R +D(Å)

c(k

).k

3

FIGURE 13.8 Fits (gray) of k3-weighted Hg LIII-edge EXAFS data (black) and corresponding RSF for Hg(II) sorbed ongoethite over the pH range 4.3-7.4. Uptake values G are indicated to the right. A vertical guideline in the RSF shows afeature consistent with second neighbour Fe atoms. Reprinted from Kim et al. (2004b), with permission from Elsevier.

2.04 Å2.03 Å

3.29

Å

3.27 Å

Hg

FIGURE 13.9 Proposed Hg(II) bonding configurationon goethite, with Hg(II) sorbing as a bidentate inner-spherecomplex linked in a corner-sharing (binuclear) arrangementto two A-Type oxygens of adjacent Fe(O,OH)6 octahedra.Reprinted from Kim et al. (2004b), with permission fromElsevier.


octrahedra at mainly the 110 face. The first peakin the RSF (Fig. 13.8) represents Hg bound totwo oxygen atoms at a distance of 2.02-2.05 A,well in agreement with distances to O atomsin twofold coordinated model compounds(e.g., HgO). The small peak occurring at �2.9 A(not corrected for phase shift) was modeled by0.4-0.6 Fe atoms, providing direct evidence foran inner-sphere complexation. The true secondshell Hg–Fe distancewas 3.19-3.28 A, suggestinga structure as depicted in Fig. 13.9. The resultconfirms the EXAFS results reported by Collinset al. (1999) at pH 4.6 and extends the finding toa pH-interval of 4.3-7.4.

Similar to goethite, bayerite formed inner-sphere complexes with Hg in the pH-range5.1-7.9. The picture was, however, more compli-cated with Al atoms occurring at two distinctdistances in the second coordination shell

4.272.

07

2.50

3.16

Cl

Hg

FIGURE 13.10 Proposed Hg(II) bonding configurationon goethite in presence of chloride, with Hg(II) sorbing asa Type-A ternary surface complex bonded monodentatewith chloride. Reprinted from Kim et al. (2004c), withpermission from Elsevier.


(at 3.06-3.09 and 3.35-3.40 A). This suggeststhat Hg may form a mixture of bidentate, cor-ner-sharing (binuclear) complexes with twoAl-octahedra similar to the structure at goethite(Fig. 13.9) and bidentate, edges-sharing (mono-nuclear) complexes with only one Al-octahedra.Monodentate mononuclear complexes with asingly coordinated oxygen is also a possibility.

Kim et al. (2004c) followed up their work onthe pH-dependency with a study on Hgadsorption to goethite (a-FeOOH), g-alumina(Al2O3), and bayerite [b-Al(OH)3] in the pres-ence of chloride (Cl�) and sulfate ions (SO4

2�)at pH 6.0. Chloride ions form relatively strongcomplexes with Hg(II) and at 1 mM of Cl�,HgCl2 (aq) will be the predominate inorganiccomplex at pH-values below approximately7.0. This means that HgCl2 (aq), rather thanHg(OH)2 (aq) or Hg2þ (H2O)6 (aq), is the formof Hg(II) that may control the surface complex-ation to mineral surfaces in soils and sedimentswith a salinity of 1 mM or more (and in theabsence of dissolved organic matter). Sulfate,on the other hand forms relatively weak com-plexes with Hg(II), and is outcompeted byhydroxyl ions even at very low pH and molarconcentrations of sulfate.

Because of the neutral charge of the HgCl20

(aq) complex it will not be attracted to positivelyor negatively charged surface groups at mineralsurfaces. Thus, the adsorption to the Fe- and Al-oxyhydroxide surfaces decreasedwith increasingchloride concentration. Once adsorbed, however,Hg will retain one Cl atom and form a type Aternary complex with goethite surfaces at pH6.0 (Fig. 13.10). Less surface bonding of Hg inpresence ofClwas further emphasized by the factthat radiation damage [photoreduction of Hg(II)to Hg(I)] was observed at high chloride con-centration. Photoreduction was occurring in allexperiments with g-alumina and at concentra-tions exceeding 1 mM of Cl in equilibrium withbayerite.

In the presence of sulfate ions the uptake ofHg(II) increased slightly at the goethite surface.

This was not interpreted as caused by achanged mode of complexation at the surface,as compared to systems without sulfate, butrather because a build up of sulfate near thesurfaces reduced the electrostatic repulsion thatpositively charged Hg-ions encounter whensorbing to a surface with a positive net charge(i.e., at pH below its point of zero charge).The mode of complexation at the surface wasnot observed to change in the presence ofenvironmentally relevant concentrations ofsulfate on neither gotheite nor bayerite.

2.3.2. Phyllosilicates

Studies on Hg adsorption to montmorilloniteand vermiculite have been reported by Brigatti


et al. (2005) and by Malferrari et al. (2008). Inboth studies Hg(II) was added to give a concen-tration equivalent to the cation-exchange capac-ity, corresponding to 28,000-80,000 mg Hg kg�1

(dry mass). Results showed that Hg(II) wascomplexed by six O atoms. Three oxygens wereassigned a distance of 1.95-2.00 A and three adistance of 2.3-2.4 A. Together with secondshell Hg–Hg scattering, the former distancesindicates a formation of precipitates like HgO(s)and Hg(OH)2(s). The longer distance wasinterpreted as indicative of hydrated Hg2þ ionsadsorbed in the interlayer. Further studies atmore relevant (lower) Hg concentrations areneeded to reveal surface complexes with phyl-losilicates which are relevant for natural sit-uations in soils and sediments. Given theprevalence of Hg for reduced S functionalities,and the low density of aluminol and silanolgroups on phyllosilicate edges, it is not likelythat surface complexes between Hg and phyllo-silicates surface groups are of any importancein the environment.

2.4. EXAFS Studies on ChemicalSpeciation of Mercury Under Suboxic andReducing Conditions

2.4.1. Model Systems: HgS(s)

The biogeochemistry of mercury under low-redox conditions is in current research focus(e.g., Munthe et al., 2007). The reason for thisis that inorganic Hg is transformed to the highlytoxic MeHg under sub- and anoxic conditions.Ofmajor concern is the formation of bioavailableforms of Hg in solution, like Hg(SH)2

0 (aq) andLMWHg-thiols. To understand themercury bio-geochemistry under sub- and anoxic conditions,the association of Hg and MeHg to organic S,inorganic S, and various types of iron sulfidesneed to be fully understood. So far such studiesare quite few. For MeHg, not a single study onthe adsorption to HgS(s) or FeS(s) or FeS2(s)exist. For Hg, X-ray photoelectron spectroscopy

(XPS) has been used in addition to Hg EXAFSto study such systems.

Cinnabar (a-HgS) is the stable low-temperaturephase of HgS(s). It is trigonal and built as infinite–S–Hg–S–Hg– chains. As a consequence two Satoms are situated at distances of 2.38, 3.10, and3.30 A from the central Hg atom (e.g., Charnocket al., 2003). Due to distortions in the structure,the two distant groups of S atoms commonly donot significantly improve fits to of EXAFS data.The same holds for the backscattering of threeHg atoms at 3.79 A and 12 Hg atoms at 4.14 A.In sediments and soils, metacinnabar (b-HgS)seems to be the most stable form of HgS (Barnettet al., 1997). In crystalline form metacinnabar iscubic and coordinated with four S atoms at a dis-tance of approximately 2.50-2.55 A. Higher shellcontributions from Hg and S commonly do notimprove fits. In Fig. 13.11 first derivatives of HgLIII-edge XANES spectra and Fourier transformsof EXAFS data for cinnabar and metacinnabarare compared. Note differences in EXAFS fre-quency and corresponding shift in the RSF peakfrom�1.9 A (not corrected for phase shift) for cin-nabar to �2.1 A for metacinnabar.

Charnock et al. (2003) combined XAS andXRD to study the kinetics of HgS(s) formationfrom sulfidic solution. They found that the pre-cipitation is a multistep process starting with a2-coordinated structure (one Hg–S bond lengthat 2.32 A and one at 2.96 A) that rapidly (after5 s) is transformed to 4-coordinated clusterswith a local structure similar to metacinnabar.After hours of aging a cubic structure is formedand eventually turned into crystalline metacin-nabar, b-HgS(s).

2.4.2. Model Systems: Reactions betweenHg and Iron Sulfides

Ehrhardt et al. (2000) and Behra et al. (2001)used XPS to study the adsorption of Hg(II) ontosurfaces of pyrite (FeS2). In the latter study, XPSwas combined with XAS. In the study byEhrhardt et al. it was shown that Hg(II) was

k3 w

eigh

ted

inte

nsity

k (Å−1)

2 4 6 8 10

Nor

mal

ized

1st

deriv

ativ

e

A

B C0 1 2 3 4 5

E (keV)R (Å)

12.3112.2912.27

FIGURE 13.11 Hg LIII-edge EXAFS datafor metacinnabar (b-HgS, solid line) and cinna-bar (a-HgS, hatched line), illustrating differ-ences in phase-shift (A), Fourier-TransformedHg–S distances (B) and first derivatives (C).


adsorbed both to S–S surface groups of pyriteand to O functional groups at oxidized patcheswith Fe-oxyhydroxides. No signs of HgS(s), sul-fate, or thiosulfate formation were observed atpH 4.0. In the study of Behra et al. the pH inter-val was extended from 2 to 12. Mercury wasadded as Hg-nitrate at the maximum adsorptioncapacity of FeS2 for Hg of 25 mmol m�2,corresponding to 2000 mg Hg kg�1 (dry mass).At this concentration both oxidized and reduced(pyritic) surfaces were likely saturated by Hg(II).

The XPS data from Behra et al. (2001) areillustrated in Fig. 13.12. The Fe2p3/2 spectrashow a predominance of pyritic Fe(II)-S sur-faces, but at higher pH also Fe(III)-O sites con-tribute significantly to the spectrum. The S2pspectra show the pyritic S–S surface groups, S(-I), and no signs of S(-II). The shoulder at�164 eV indicates formation of polysulfidegroups (–Sn–). The doublet seen in the Hg4f7/4spectra was by Ehrhardt et al. (2000) attributed

to Hg(II) adsorbed to Fe-oxyhydroxide patches(104.5-104.8 eV) and to pyritic sites (100.6-100.8 eV). Lack of a peak at 99.2-99.8 eV wastaken as an indication that Hg0 was not formedin these experiments. Owing to a small surfacearea of the sample, the collected Hg LIII-edgeEXAFS data were quite noisy, and only first-shell data were modeled at a pH of 3.5. It wassuggested that Hg(II) formed 2-coordinated,ternary complexes with one S–S group at thesurface and with one OH� ion in solution(S–S–Hg–OH). Consensus was not obtainedconcerning the Hg–O distance of 2.24-2.25 A,which is considerably longer than distances of�2.05-2.10 A in 2-coordinated HgO (s) and Hg(OH)2 (s), but substantially shorter than theHg–O distance of 2.41 A in the Hg(H2O)6

2þ

ion. In 0.1 M NaCl, the OH� ion was exchangedfor a Cl� ion resulting in the ternary complexS–S–Hg–Cl. The S–Hg distance was 2.37-2.41 A and the Hg–Cl distance was 2.33 A. Both

N(E)(cts/s)

N(E)(cts/s)

Fe(III)-O711 eV

S(-I)162.7 eV

104.6 eV

104.5 eV

104.8 eV

100.6 eV

100.6 eV

100.8 eV

2.10 eV

1.75 eV

1.95 eV

S(-II)

Fe(III)-S708.5 eV Fe(II)-S

707.5 eV

pH 11.5

pH 11.5

pH 11.5

pH 3

pH 3

pH 3

pHi 9.5, pHf 6

pHi 9.5, pHf 6

pHi 9.5, pHf 6

Fe2p

S2p

Hg4t

104

103

0

N(E)(cts/s)

103

0

0

A B

C

715 710 EBF (eV)

EBF (eV)

EBF (eV)170

110 100

160

FIGURE 13.12 XPS spectra of (A) Fe 2p, (B) S 2p, and (C) Hg 4f in Mg Ka after sorption of Hg(II) onto FeS2 for differentpHs. Experimental conditions: I ¼ 0.01 M NaNO3; 10 g L�1 FeS2; [Hg(II)]total ¼ 100 mM; 12 h (no Hg) þ 24 h (with Hg) con-tact time. Initial pH, and if different from initial, final pH are indicated for each spectrum. From Behra et al. (2001). Printedwith permission from American Chemical Society.



distances indicate a 2-coordination. No signsof redox transformations involving Hg(II) andS(-I), forming Hg(0) and S(�II, þVI) wereobserved in these experiments. Neither couldprecipitation of a HgS(s), nor a mixed Fe/HgS(s) phase, be observed at the pyrite surface.

Even if pyrite is the most common of allFe-sulfides, metastable or microcrystalline FeS(mackinawite), may dominate in reduced,organic rich sediments. A Hg LIII-edge EXAFSstudy of the interaction between Hg(II) andmackinawite by Wolfenden et al. (2005) is there-fore of great interest. In a first experiment,Hg(II) was added together with dissolved Fe(II)and S(-II) in aqueous solution in order to allowfor a coprecipitation of Hg(II) with mackina-wite. The total Hg concentration was 500 mgkg�1 (dry mass). The resulting EXAFS resultsshowed an Hg–S distance of 2.38 characteristicof a 2-coordinate Hg center similar to cinnabar.Because a true coprecipitate of Hg with FeSshould be 4-coordinated, there were no signsof Hg(II) substitution for Fe(II) in the mackina-wite lattice. This observation is in line withtheoretical considerations (Jeong and Hayes,2003; Jeong et al., 2007) suggesting that thesubstitution of Hg(II) into mackinawite is unfa-vorable due to electronic configurations andby geometrical restriction. Iron (II) is 4-coordi-nated with S(-II) in mackinawite, similar to thestructure in metacinnabar in which Hg(II) is 4-coordinated by S(-II), but the Fe–S distance of�2.24 A is much shorter than the Hg–S distanceof �2.5 A. Thus, unless the mackinawite struc-ture is substantially rearranged, coprecipitationof traces of Fe(II) in HgS is more favorable thanvice versa. In a second experiment the adsorptionof Hg(II) to the surface of mackinawite wasinvestigated. The Hg–S distance was the sameas in the coprecipitation experiment (2.37 A),thus indicating amaintained cinnabar-like struc-ture, but the fitted number of S atoms changedfrom 2 to 4. Given the uncertainty in CNs in theEXAFS analyses of �20%, the interpretation isnot clear. Interestingly enough,when the sample

with Hg adsorbed to FeS was partly oxidized[S XANES determination showed approxi-mately 50% S(-II) and 50% S(þVI)], both Hg–Sbond distances (2.53 A) and CNs pointed at afourfold coordination of Hg. These results sug-gest either a precipitation of metacinnabar(b-HgS at the FeS surface, or a formation ofHg(II)–S–Fe surface complexes. Unfortunatelythe data quality was not good enough to allowan analysis of the second coordination shell.Jeong et al. (2007) showed in a wet chemicalstudy that surface adsorption of Hg in presenceof Cl� ions (0.05-0.2 M) occur at nanocrystallinemackinawite surfaces if the molar ratio of initialconcentrations of HgCl2

0 to FeS is lower than0.05 (which is the case in many sediments).When the ratio increased the adsorption capac-ity of FeS(s) was exceeded and metacinnabarwas precipitated.

2.4.3. In Search for Hg-PolysulfideFormation

Polysulfide formation in the aqueous phaseis of large importance for the biogeochemistryof Hg because they may increase the solubilityand bioavailability of Hg(II). Polysulfides areformed under sub- and anoxic conditions andthey have been shown to enhance the solubilityof Hg(II) in presence of HgS(s) and elementalS8

0. Paquette and Helz (1997) explained theenhanced solubility of HgS(s) in presence ofelemental S in the pH range 1-12 at a total sulfideconcentrations of 10�3-10�1 M by the formationof the polysulfide HgSnHS� (n ¼ 2-6). Jay et al.(2000) argued that the species HgSnOH� andHg(Sn)2

2� could better explain theHg(II) solubil-ity in the pH interval 7-10, especially at lowertotal sulfide concentrations (10�6-10�3 M), thanthe species suggested by Paquette and Helz(1997). They also proposed that a lipofilic HgS5(aq) species may contribute to the bioavaiablepool of Hg for methylating bacteria.

No definite evidence for Hg-polysulfide for-mation was observed in two EXAFS studies onaqueous phase Hg-sulfide systems (Bell et al.,


2007; Lennie et al., 2003). In the study of Lennieet al. pH was 11. First shell analysis and MScontributions clearly showed the dominanceof Hg complexed by two S atoms at 2.30 A ina linear S–Hg–S arrangement. This is in agree-ment with solubility data from Schwarzenbachand Widmer (1963) showing a predominanceof the HgS2

2� complex. In the study of Bellet al. one sample showed possible signs ofpolynuclear Hg–S formation, but it could notbe ruled out that colloids of HgS(s) passed thefilter. No signs of polysulfide formation wereobserved in the sample with a pH of 9.5 with6 mM of total sulfides.

2.4.4. Experiments on Hg–Fe–STransformations

Slowey and Brown (2007) studied the trans-formation of Hg(II) to HgS(s) as a consequenceof addition of a bisulfide (HS�) solution at dif-ferent rates (40-310 mM h�1) to Hg(II) initiallyadsorbed to newly synthesized goethite. This“sulfidation” experiment was conducted underN2 atmosphere and pH was maintained at 7.5.Concentrations of dissolved Hg, Fe, and sulfidewere monitored in solution and Hg EXAFS andS XANES data on the solid phase were col-lected at different sulfide loadings. Initiallythe concentration of Hg(II) decreased in solu-tion. This was explained by the formation ofmetacinnabar [b-HgS(s)] at the expense of Hg(II) adsorbed to goethite (Fig. 13.13). It shouldbe noted that the EXAFS analysis did not enablea definite dismissal of adsorption of Hg(II)to possible traces of FeS(s), or formation ofHg/FeS(s) coprecipitates.

After the initial decrease in Hg(II) concen-tration, at the moment when concentrations ofdissolved S(-II) and Fe(II) became detectable,the concentration of Hg(II) increased from0.2 to 30 nM. This was interpreted as an effectof Hg-polysulfide formation in solution. Theobservation was corroborated by S K-edgeXANES spectroscopy, showing a formation ofelemental S8

0, along with S(IV) and S(VI). It

was concluded that sulfide S(-II) reacted withFe(III) of goethite under the formation of ele-mental S(0), polysulfides and a solid phase ofFe(II/III) hydroxysulfate green rust. The over-all consequence was an enhanced dissolutionof b-HgS(s) by Hg-polysulfides. This processis likely to occur in sub- and anoxic sedimentswith input of Fe(III) and sulfate.

2.4.5. Chemical Speciation of Mercury inIntact Soils and Sediments

There are very few XAS studies on the Hgspeciation in intact soils and sediments. Wanget al. (1995) used Hg EXAFS spectroscopyto characterize a soil sample from a mercurycontaminated flood plain (280 mg Hg kg�1

dry mass) at Lower East Fork Poplar Creek,oak ridge, Tennessee, USA. A first shell Hg–Sdistance of 2.55 A and Hg–O distance of2.04 A was interpreted as a mixture of Hg4-coordinated with S and 2-coordinated withO ligands. A least-square linear combinationfit resulted in 80% metacinnabar and 20% mon-troydite (HgO). Barnett et al. (1997) combinedscanning electron microscope (SEM) withX-ray dispersive spectroscopy, transmissionelectronmicroscope, and select area electron dif-fraction to characterize samples from the sameflood plain [15-2700 mg Hg kg�1 (dry mass)].They concluded that submicrometer, crystallinemetacinnabar dominated the Hg speciation inthese soils. They also constructed a thermody-namic model to calculate the redox potential[as a consequence of reactions between theelectron-donor DOC (CH2O) and the electron-acceptors O2 (aq), MnO2(s), Fe2O3(s), andSO4

2�], on Hg speciation. Based on data on pHand initial concentrations of the above compo-nents and their reaction with 15 mg DOC L�1,pe could be calculated. The result is depicted ascircles in Fig. 13.14, indicating a stability ofHgS(s), in fair agreement with the micro- andspectroscopic determinations.

Bernaus et al. (2006a) used Hg XANES spec-troscopy to characterize a soil contaminated by

20 20

10more sulfide0.2

Hg-S

Hg-OHg-Fe

0.1

0.07

0.04

no sulfide

8

6

4

2

00 1 2 3 4

16

12

8

k3 c

(k)

k (Å–1) R + D(Å)

F[k

3 c(k

)]

4

0

–42

A

C

B4 6 8 10 12

100

80

60

40

20

00 0.05

β-HgS(s)

μm

ol H

g L–1

adsorbed

0.1[S(II)] : [ = FeOH]

0.15 0.2

FIGURE 13.13 k3-weighted Hg LIII-edge EXAFS data (A), radial structure functions (B), and quantitative Hg speciation(C) for a sulfidation experiment. Dotted lines are experimental data and solid lines are theoretical fits. In (B) D is phase shiftand the numbered annotations indicate sulfide loadings. Reprinted from Slowey and Brown (2007), with permission fromElsevier.


a chlor-alkali plant (4-1150 mg Hg kg�1). Princi-ple component analysis (PCA) was combinedwith linear combination fitting to model thesoil data. A mixture of cinnabar and corderoite(Hg3S2Cl2) gave the best model fits.

Wolfenden et al. (2005) used Hg EXAFS toinvestigate the speciation of Hg in two lake sedi-ments with 10.5% and 12.4% organic carbon.Because of low concentrations, Hg(II) nitrate

was added to reach 440-470 mg Hg kg�1 (drysediment mass). The best model fit wasobtained by a twofold coordination by either Satoms at 2.37 A or Cl atoms at 2.29 A. The atomicmass of S and Cl are too close to be distin-guished by conventional EXAFS. However, X-ray imaging at 1 mm resolution showed thatHg always was associated with S, and no corre-lation was observed with Cl. Thus, coordination

0 2 4 6 8 10 12 14pH

0

10

–10

20

pe

HgCl2 (s)

Hg2Cl2 (s)

HgO (s)

Hg (l)

HgS (s)

a

b

c

FIGURE 13.14 Theoretical stability of mercuryforms at a function of pe and pH. Three principalredox regions correspond to the reduction of oxygen(a), pyrolusite (b), and hematite/Fe(II) couple (c) byoxidation of dissolved organic matter. Circles showdata calculated by the hematite/Fe(II) couple.Diagram constructed with 2 � 10�4 M chloride and6 � 10�4 M total sulfur at 25 �C. From Barnettet al. (1997). Printed with permission from AmericanChemical Society.


with two S atoms was much more plausible. Inthe absence of a second shell analysis, it wasnot possible to establish whether the S atomswere organic thiols or an inorganic a-HgS(s)phase.

2.5. Chemical Speciation of Mercury inMine Waste and Downstream Effects

Mine waste is regionally a severe environ-mental problem. In areas with geological depos-its andmining activities, like the coastal range ofCalifornia, mine waste is amajor source of Hg(0)to the atmosphere (Coolbaugh et al., 2002), aswell as a source of dissolved and colloidal Hg(II)to soils, sediments, lakes and estuaries (Conwayet al., 2008; Rytuba, 2000).

The chemical speciation of Hg in mine wastesand tailings from a portion of the in total morethan 70 abandoned mining sites in the Californiacoast range has been described in a number ofpapers (Kim et al., 2000, 2004a; Lowry et al.,2004). Also the effect of Hg(0), its transformationsand release from gold mine tailings has been

investigated (Slowey et al., 2005). The researchgroup of Brown has successfully used HgEXAFS, in combination with wet chemicalextraction and other spectroscopic methods, todetermine and quantify different Hg phases bothin mine tailings as well as in colloids spread withstreams downstream the mining areas. Twotypes of geological Hg ore deposits have beenexamined; hot-spring Hg deposits and silica-car-bonate Hg deposits. Tailings consist of roasted(at 600 �C), so-called calcined Hg ore andremains of untreated Hg ore. The Hg ore is pri-marily cinnabar. During the roasting process cin-nabar (hexagonal HgS) is transformed primarilyto metacinnabar (cubic HgS), but also to moresoluble phases like corderoite (Hg3S2Cl2), mon-troydite (HgO), mercuric chloride (HgCl2), terlin-guite (Hg2OCl), and schuetteite (Hg3O2SO4).Thus, both the geological origin and the roastingprocess affect the composition of Hg in minewastes (Kim et al., 2004a).

A least-square fitting procedure was used tolinearly combine EXAFS data from a set of rea-sonable model compounds. In Fig. 13.15 an

Turkey run calcineLinear comb. fit1

0

–1

–2

–3

–4

21 3 4 5 6 7 8 9

Cinnabar

Metacinnabar

k3 .

c(k

)

k (Å–1)

FIGURE 13.15 Experimental HgLIII-edge EXAFS data (black line) andlinear combination fits (gray line) andthe two components that contribute tothe linear fit (dashed lines). Whenscaled to a total of 100%, cinnabarand metacinnabar contributed with58% and 42%, respectively. Reprintedfrom Kim et al. (2000), with permissionfrom Elsevier.


example is given from Kim et al. (2000). Themodeling of this particular calcine sample sug-gests that 42% of the originally pure cinnabarore was transformed to metacinnabar duringthe roasting processes and during possiblesubsequent processes in the mine waste deposit.The linear combination fitting procedure wasestimated to give an accuracy of approximately25% for heterogeneous mine wastes, havingHg concentrations of approximately 200-1000mg Hg kg�1 (Kim et al., 2000). Bernaus et al.(2006b) used m-EXAFS analysis to determinethe speciation of Hg in calcines from the Alma-den mine in Spain. Three different model com-pounds were used to fit the EXAFS data usinga principal component analysis procedure.Cinnabar varied between 5% and 89%,montroy-dite between 5% and 55%, and schuetteitebetween 6% and 49% in the samples analyzed.Absence of metacinnabar in the calcine samplesat Almaden was explained by the authors by alower roasting temperature than at the miningsites in California.

An interesting question is in what form Hgmay be transported downstream from minewaste tailings. Because of the low aqueoussolubility of most Hg phases, colloidal trans-port during episodic events may account for

as much as 95% of the Hg release from minetailings exposed to whether and wind (Lowryet al., 2004). Furthermore, the concentrationof Hg increased with decreasing size of parti-cles. Transmission electron microscope andHg EXAFS analyses showed that 80% of Hgin colloids generated from calcines and wasterock was in the form of cinnabar or metacin-nabar. In Fig. 13.16 linear combined fits areshown for colloids generated in a columnexperiment.

Because of the use of Hg(0) to collect Au(0)and Ag(0) particles (through amalgamation),mercury contamination is highly associatedwith gold and silver mining activities. In astudy of placer gold mining tailings, Sloweyet al. (2005) used sequential chemical extrac-tions (SCEs) to show that mercury was trans-formed from Hg(0) to relatively soluble Hg(II)oxides and chlorides (3-4%), to organic/inor-ganic sorption complexes, to amalgamations(75-87%) and to highly insoluble phases ofHgS (6-20%). Despite the relatively minor con-tribution from HgS to the total amount of Hg,column experiments combined with EXAFSspectroscopic analysis showed that colloids ofcinnabar represented the dominant mobile Hgphase that could be transported down stream.

1

New idria calcines(colloid generation)36% cinnabar (HgS)46% metacinnabar (HgS)22% montroydite (HgO)

Sulphur Bank calcines(pre-equlibration)

Sulphur Bank waste rock(colloid generation)

100% metacinnabar (HgS)

49% cinnabar (HgS)51% metacinnabar (HgS)

0.5

0

–0.5

–1

–1.5

3

2

1

0

–1

–2

1.51

0.50

–0.5–1

–1.5–2

2 3 4 5 6 7 8

2 3 4 5 6 7 8 9 10 11

2 3 4 5 6 7 8

c(k

) ×

k3

c(k

) ×

k3

c(k

) ×

k3

k (Å–1)

k (Å–1)

k (Å–1)

FIGURE 13.16 Linear combination fits ofHg LIII-edge EXAFS data for mine tailing col-loids generated in the effluent of leachingcolumns. The quantitative estimates have anerror of approximately �10%. Black line exper-imental data and dashed line calculated fits.From Lowry et al. (2004). Printed with permis-sion from American Chemical Society.


2.6. Comparison of Hg EXAFS with WetChemical Extraction and ThermoDesorption

A major limitation of Hg EXAFS spectro-scopic analysis of environmental solid samplesis that concentrations of Hg, even when usingthe brightest third-generation SR sources, needto be at least 50 mg kg�1 to yield reasonablesignal to noise ratios. This level of concentra-tion is encountered mainly in soils and sedi-ments severely contaminated by local pointsources, or in mine tailings. With a focusedbeam (m-EXAFS), single particles with high

concentrations of Hg may be investigated inan otherwise less contaminated matrix. In mostcases, however, EXAFS analysis of less con-taminated samples requires additions of Hg(or MeHg). Obviously, there is a need forcomplementary methods that can be used fora speciation analysis of Hg at lower concentra-tions in intact soils and sediments. SCE andsequential thermo desorption (TD) analysisare two methods in use. Both these methodshave been compared with Hg EXAFS analysis.

Kim et al. (2003) compared the outcome ofthe SCE method of Bloom et al. (2003) with


Hg EXAFS determinations on samples fromgold mine tailings and marine sediments, andHg adsorbed to kaolinite and copper smelterfly ash. The Hg concentration in the samplesvaried between 130 and 7500 mg Hg kg�1. It wasconcluded that Hg dissolved by aqua regia,the fifth and last step of the SCE (1, water;2, HCl/HOAc at pH 2.0; 3, 1 M KOH; 4, 12 MHNO3) was in good agreement with HgS (thesum of cinnabar and metacinnabar) as deter-mined by Hg EXAFS. A similar result wasobtained in calcine and waste Hg ore (Lowryet al., 2004). Differences in Hg speciationobtained by the two methods were associatedwith the more soluble Hg phases HgCl2 andHgO and were explained by factors affectingthe SCE method like; encapsulation of Hgphases into less soluble particles, correlation ofsolubility with particle size and crystallinityand secondary effects with the mineral matrix.It was recommended that several independentmethods, including spectroscopic methods likeSEM, TEM, and XRD, are combined with SCEand EXAFS at low concentrations of Hg.

Extinction A

1.4

1.2

Hg�

HgCl2

Hg bound tohumic acids

1.0

0.8

0.6

0.4

0.2

0.00 100 200 300

�C

FIGURE 13.17 Mercury release curves for standard Hg cScholz (1997). Printed with permission from American Chemic

Another useful method for Hg speciation oflow contaminated samples is TD (Biester andScholz, 1997). The temperature is increased ata rate of 0.5 �C s�1 and the result is depictedas a release of Hg versus temperature. As canbe seen in Fig. 13.17, pure Hg compounds havedifferent maximum temperatures at which theyare released. Mercury TD curves of pure com-pounds may be used to identify Hg phases innatural samples. The method has a detectionlimit of approximately 0.2 mg Hg kg�1 and ithas been used by the group of Biester to finger-print the composition of Hg in mining residues(Biester et al., 2000) and soils (Biester andScholz, 1997; Bollen et al., 2008). Sladek et al.(2002) compared the TD method with HgEXAFS in mine waste. The two methods arecomplementary because of the lower detectionlimit of TD. Elemental Hg(0) is most easilydetected by TD. It is also possible to separateand quantify HgCl2 from Hg(II) bound to thesoil matrix. A definite separation of Hg(II)bound to organic matter (Hg-thiols), from non-crystalline HgS phases like metacinnabar is

HgS

Standard Hg-compounds

400 500 600 700

ompounds as a function of temperature. From Biester andal Society.


more difficult and at this point the TD methodmay be at best semiquantitative for differenttypes of Hg–S compounds.

12,270 12,280 12,290 12,300Energy (eV)

0

0.2

0.4

0.6

0.8

Nor

mal

ised

abs

orba

nce

roots

shoots

FIGURE 13.18 Experimental Hg LIII-edges ofE. crassipes roots (thick, solid line) and shoots (thin, solidline) and their first derivatives (roots: longer broken lines;shoots: shorter broken lines). From Riddle et al. (2002).Printed with permission from American Chemical Society.

2.7. Use of Hg EXAFS to MonitorRemediation of Mercury

Different methods have been suggested for aremoval of mercury from waste and ground-water. Zero-valent iron has been proposed asa reactive media for remediation of variousmetalloids like, for example, As and Hg. In afield column study, zero-valent iron was usedto trap and transform Hg(II) from a contami-nated groundwater (Weisener et al., 2005).Groundwater was pumped over a 42-dayperiod into a column packed with Fe(0) andhoused in a mobile trailer. The concentrationof Hg(II) in the influent to the column rangedfrom 18 to 42.5 mg L�1 and pH was 7.8-9.5. Asa consequence of reactions between compo-nents in the groundwater and Fe(0), pH roseto 9-10 in the effluent and the Hg(II) concentra-tion was lowered to 0.17-0.02 mg L�1. Theremoval of Hg was inversely related to the flowrate. SEM combined with energy dispersive X-ray spectroscopy (SEM-EDX) revealed a forma-tion of a HgS(s) phase in the lower half of thecolumn. The finding was confirmed using m-SXRF and m-EXAFS with a spatial resolutionof a 1-5 mm, and indicated a formation ofmainly metacinnabar. It was suggested thatiron was oxidized to ferrous iron and sulfatein the groundwater was reduced to sulfideaccording to the reaction:

Hg2þ þ 4Fe0ðsÞ þ SO42� þ 4H2O

¼ HgSðsÞ þ 4Fe2þ þ 8OH� ð13:3Þ

It should be noted that only first shell EXAFSspectroscopic data were extracted, and there-fore processes like adsorption of Hg(II) ontoFeS(s) surfaces or the possibility of mixed Hg/FeS(s) phases could not be examined (see dis-cussion above).

The strong bond formation between Hg(II)and S(-II) has resulted in a construction of avariety of thiol/mercaptan functionalizedtrapping material for Hg(II) remediation ofcontaminated water. Several of these materialslike silicates (Chen et al., 2004) and thiacrowns(Ito et al., 2005) have been characterized usingHg LIII-edge EXAFS spectroscopy. Some ofthese studies are reviewed by Andrews (2006).

2.8. Mercury Speciation in OrganismsUsing XANES and EXAFS

Recent mercury speciation analysis of organ-isms has taken advantage of XAS techniques.Riddle et al. (2002) investigated the speciationof Hg taken up by water hyacinth (Eichhorniacrassipes) experimentally exposed to 1 mg L�1

solution of Hg-nitrate. After 28 days of expo-sure the concentration of Hg in roots reached15 mg kg�1 and in shoots 0.2 mg kg�1. Becausethese concentrations are too low for a conven-tional EXAFS analysis, Riddle et al. used thefirst derivatives of the Hg XANES to determinethe speciation of Hg. More specifically, theyused the distance between the first two inflec-tion points. As illustrated by Fig. 13.18, the


longer distance between the two first derivativepeaks in the root, as compared to the shoot,reflects a shift from coordination with mainlyS atoms (thiols) in shoot to coordination withO/N atoms in the root. Similar shifts havereported for the shift from thiols to carboxyls/amino groups in NOM (Skyllberg et al., 2006,supporting information).

In fish, mercury mainly occurs in the form ofMeHg. The first spectroscopic evidence wasobtained using Hg XANES spectroscopic analy-sis on samples of swordfish (Xiphias gladius),orange roughy (Hoplostethus atlanticus), andsand sole (Psetttichys melanostictus), Harriset al. (2003). Similar to the study of Riddleet al. (2002) concentrations of Hg in the sampleswere too low for conventional EXAFS analysis.A qualitative comparison of the XANES regionwith model compounds showed a good corre-spondence with MeHg complexed by one thiolgroup in cysteine.

Arai et al. (2004) combined Hg LII-edgeEXAFS and Se K-edge EXAFS to determinethe speciation of mercury in liver of northernfur seal (Callorhinus ursinus) and black-footedalbatross (Diomedea nigripes). Concentrations ofHg were in the range of 110-2100 mg kg�1

(dry mass), allowing for a first coordinationshell analysis. They concluded that HgSe wasthe major form in the liver of the seal and thata mixture of HgS and HgSe was found in thealbatross liver. This finding was in agreementwith previous studies using TEM and X-raymicroanalysis.

3. CONCLUSIONS

Applications of XAS the last decade haveled to a significant improvement of our under-standing of mercury biogeochemistry in soils,sediments, and surface waters. We can con-clude that Hg and MeHg associate strongly

to low and high molecular weight thiols underoxic conditions, both in aqueous solution andin association to organic particle surfaces.Also in mineral soils and sediments is the con-centration of free LMW thiols, or thiol func-tional groups associated to HMW organicmatter, normally in large excess of Hg andMeHg. This holds also for moderately contami-nated conditions. Thus, under oxic conditionsassociation of Hg and MeHg to mineral phasesis typically limited to a secondary effect ofNOM adsorption. Under sub- and anoxic con-ditions, however, Hg and MeHg form strongcomplexes with inorganic sulfides (S2�), bisul-fides (HS�) and polysulfides in aqueous solu-tion as well as with sulfides at the solution-particle interface. XPS and EXAFS studieshas shown that metacinnabar (b-HgS, cubic)is the dominant HgS-phase in soils andsediments, as well as in colloids transportedfrom mine tailings. Adsorption of Hg(II) toFeS(s) surfaces, or possible coprecipitatates ofHgFeS(s), have been shown in a couple of stud-ies but these processes and chemical formsneed to be studied more in detail. The adsorp-tion of MeHg to FeS(s), HgS(s) and to othersolid phases and surfaces remains to bedescribed. In mine tailings and highly contami-nated soils comparatively soluble phases likecorderoite (Hg3S2Cl2), montroydite (HgO),and mercuric chloride (HgCl2) may form.Under such conditions, colloidal transport ofHg(II) adsorbed to surfaces of iron, aluminum,and manganese oxy/hydroxides may play arole. Because of the complexity of mercury bio-geochemistry and its tight links to the biogeo-chemistry of carbon, iron, and sulfur, amultidisciplinary approach using bulk andmicrofocused applications of Hg, S, and FeXAFS, isotope-labeled bioassay experiments incombination with other spectroscopic techni-ques will be required for further improvementof our understanding of mercury transforma-tions in the environment.

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