nyre

ey,235

ciat

cal rnsisiour

2009 Elsevier Ltd. All rights reserved.

metallccur ng been

leishmaniasis (Vasquez et al., 2006), and in treating HIV (Fowlerand Goering, 1991). Arsenic is used industrially in the manufacture

oxides, hydroxides or oxoanions either in the 5 state in relativelyoxic environments (antimonates and arsenates) or in the 3 statein anoxic environments (antimonites and arsenites). Somecommon Sb and As species in the earth surface environment areshown in Table 1.

The toxicities of Sb and As in the environment strongly dependupon speciation (Gebel, 1997; Smith et al., 1998; Filella et al.,2002a). The general order of toxicity for Sb species is given as:

* Corresponding author. Tel.: 61 2 6773 2789; fax: 61 2 6773 3238.E-mail addresses: swilso24@une.edu.au (S.C. Wilson), peter.lockwood@une.

edu.au (P.V. Lockwood), pashley@une.edu.au (P.M. Ashley), mtighe2@une.edu.au

Contents lists availab

Environment

journal homepage: www.els

Environmental Pollution 158 (2010) 11691181(M. Tighe).between the two metalloids have prompted concerns over theenrichment of this metalloid in many environments (Krachler et al.,2001; Filella et al., 2002a). Often Sb is considered to behave similarlyto As, not always with justication (Casiot et al., 2007).

Antimony has awide range of uses including themanufacture ofsemiconductors, diodes, ameproof retardants, lead hardener,batteries, small arms, tracer bullets, automobile brake linings, andpigments (Filella et al., 2002a). Antimony also remains the treat-ment of choice for several tropical protozoan diseases, such as

geological mineralisation, but also anthropogenically (Ragaini et al.,1977; McLaren et al., 1998; Filella et al., 2002a; Wilson et al., 2004;Douay et al., 2008; Telford et al., 2009) although the majority of Sbcontamination appears to originate from mining and industrialemission sources, often smelting, co-occurring frequently with As(Telford et al., 2009).

Because of their identical s2p3 outer orbital electron congura-tion, Sb and As display the same range of oxidation states in envi-ronmental systems (3 to 5). Both most commonly occur aseffects. The understanding of its toxicity and environmental behav-iour ismuchmore limited (Filella et al., 2002a). Chemical similaritiestially harmful element, evidenced through its historic use asa popular poison (Azcue and Nriagu, 1994). Less is known about Sb

World production for Sb is considerably larger than As. In 2008estimated Sb mine output was 165,000 tonnes compared to40,500 tonnes for As (U.S. Geological Survey, 2009). Environmentalenrichment of both metalloids occurs naturally in areas ofAntimonyArsenicSoilsAdsorptionEnvironmental chemistry

1. Introduction

Antimony (Sb) andarsenic (As) are15 of the periodic table. They both oment at trace levels. Arsenic has lon0269-7491/$ see front matter 2009 Elsevier Ltd.doi:10.1016/j.envpol.2009.10.045oidsbelonging toGroupaturally in the environ-recognised as a poten-

of numerous products including glass, ceramics, electronics,cosmetics, and reworks (Smith et al., 1998). In the latter half of the20th century As was also widely used in pesticide and herbicideformulations and in wood preserving, although such use is nowdeclining (Azcue and Nriagu, 1994).Keywords:

risks associated with Sb in environmental soil systems, with particular reference to mobility andbioavailability.Accepted 29 October 2009tied and discussed in terms of the limitations this imposes on understanding the fate, behaviour andReview

The chemistry and behaviour of antimowith comparisons to arsenic: A critical

Susan C. Wilson*, Peter V. Lockwood, Paul M. AshlSchool of Environmental and Rural Science, University of New England, Armidale, NSW

A critical and comparative review of Sb and As chemistry and assounderstanding of risks.

a r t i c l e i n f o

Article history:Received 3 June 2009Received in revised form28 October 2009

a b s t r a c t

This article provides a criticomparing and contrastingreviewed, with an emphabetween Sb and As behavAll rights reserved.in the soil environmentview

Matthew Tighe1, Australia

ions in soil systems identies research directions needed for better

eview of the environmental chemistry of inorganic antimony (Sb) in soils,dings with those of arsenic (As). Characteristics of the Sb soil system areon speciation, sorption and phase associations, identifying differences. Knowledge gaps in environmentally relevant Sb data for soils are iden-

le at ScienceDirect

al Pollution

evier .com/locate/envpol

ck, 1973; Dodd et al., 1992; Jenkins et al., 1998; Bhattacharya et al., 2002; Filella et al.,

As

ula Name Formula

3 Arsenopyrite FeAsS

3 (orthorhombic) Orpiment As2S33 (cubic) Realgar AsS

4

H)5 Arsenic acid AsO(OH)3 (or H3AsO4)H)6

(or SbO3) Dihydrogen arsenate AsO2(OH)2

(or H2AsO4)

Monohydrogen arsenate AsO3OH2 (or HAsO4

2)

H)3 (or HSbO2) Arsenous acid As(OH)3 (or H3AsO3)H)2

Arsenite AsO(OH)2 (or H2AsO3

)H)4

HAsO32

2S4S4

42

Arsine AsH3H3)3

)SbO)2Sb)3Sb

tal Pollution 158 (2010) 11691181Organoantimonials (e.g. methylated species) < antimonates (Sb(V)) < antimonites (Sb (III)) (Gebel, 1997; He and Yang, 1999;Krachler et al., 2001; Filella et al., 2002a).

This is similar to As:Organoarsenicals (e.g. methylated species) < arsenates (As

(V)) < arsenites (As (III)) (Yamauchi and Fowler, 1994).It is debatable whether or not As is essential to human health,

but there is no known human requirement for Sb (Smith et al.,1998; Bhattacharya et al., 2002; Shotyk et al., 2005). Both metal-loids are clastogenic in the trivalent state, and have carcinogenicpotential (Gebel, 1997). Both have a strong afnity for thiol groupsand may substitute for P in biological reactions, which explainstheir inhibitory role in DNA replication and metabolic processes(ATSDR, 1992, 2000; Shotyk et al., 2005). There is evidence that As

Table 1Common Sb and As chemical species found in natural systems (Braman and Foreba2002a,b).

Sb

Name Form

Minerals Stibnite Sb2SValentinite Sb2OSenarmontite Sb2OCervantite Sb2O

Aqueous species (5 oxidation state) Antimonic acid Sb(OAntimonate Sb(O

Aqueous species (3 oxidation state) Antimonous acid Sb(OAntimonite Sb(O

Sb(O

Suldic complexes H2SbHSb2Sb2S

Gases Stibine SbH3Trimethylstibine Sb(C

Other methylated species Methylstibonic acid (MSA) (CH3Dimethylstibonic acid (DMSA) (CH3Trimethylstiboxide (CH3

S.C. Wilson et al. / Environmen1170is detoxied via methylation in biological systems, but lessevidence for the same process occurring for Sb (Gebel, 1997).Currently the World Health Organisation has set the AcceptableDaily Intake (ADI) for As at 2 mg kg1 day1 kg of body weight1

(WHO, 1989) and the Tolerable Daily Intake (TDI) for Sb at6 mg kg1 day1 kg of body weight1 (WHO, 2003).

This review initially compares and contrasts Sb and As specia-tion in natural systems. Characteristics and interactions of the Sbsoil system are then critically and comparatively reviewed withthose of As, with an emphasis on speciation, sorption and phaseassociations, focussing on inorganic species. Because this review isnot intended to be a review of As in soils (this has been undertakencomprehensively by other authors (Sadiq, 1997; Matera and LeHecho, 2001; Smith et al., 1998), important ndings for As aresummarised rst in individual sections and then comparativelydiscussed with Sb data available. Knowledge gaps and researchneeds for Sb in soil systems are identied.

2. Speciation in natural systems

2.1. Inorganic speciation in relation to pH and redox potential

System pH and redox largely determine metalloid oxidationstate and environmental reactions in soil systems (Sadiq, 1997;Smith et al., 1998) and some broad similarities have been identiedbetween As and Sb (Ashley et al., 2007). Arsenic and Sb speciespresent in any given system also depend upon the concentrationsof co-occurring reductants and oxidants. However, the metalloidproperties ensure their close association and strong bonding withoxygen and hydroxide in the environment with the oxyanionicform playing a central role in most environmental reactions.

Table 2 places some inorganic As and Sb redox reactions in thecontext of other common reactions that fall within the redox limitsof natural systems (Lindsay, 1979; Brookins, 1988). When water-logging causes soil or sediment to become anoxic, the reductionhalf reactions shown in Table 2 commonly occur in sequence asmicroorganisms are progressively forced to select less energeticterminal electron acceptors for respiration reactions. This suggeststhat pentavalent Sb and As will both be reduced to the trivalentstate after Fe (III) is reduced to Fe (II) but before sulfate is reduced to

Trimethylarsine As(CH3)3

(OH)2 Monomethylarsonic acid (MMAA) CH3AsO(OH)2(O)OH Dimethylarsinic acid (DMAA) (CH3)2As(O)OH

Dimethylarsine (CH3)3Assulde. Reduction of Sb (III) to elemental Sb or stibine and likewiseAs (III) to elemental As or arsine requires considerably lowerreduction potentials, at which soluble sulde may be present insufcient concentration for the metalloids to form suldecomplexes.

Table 2Some pertinent redox reactions in natural systems (adapted from Aylward andFindlay, 1994; Stumm and Morgan, 1996; Godfrey et al., 1998).

Reaction pe0a Eh0 (V)

O2 4H 4e 2H2O 20.8 1.232NO3

12H 10e N2 6H2O 21.1 1.25MnO2 4H 2e Mn2 2H2O 20.4 1.21Fe(OH)3 (amorphous) 3H e Fe2 3H2O 16.6 0.98a-FeOOH (goethite) 3H e Fe2 2H2O 13.1 0.77Sb(OH)6

3H 2e Sb(OH)3 3H2O 12.9 0.76Sb2O5 4H 4e Sb2O3 (valentinite) 2H2O 11.0 0.65Sb2O5 6H 4e 2Sb(OH)2 H2O 9.8 0.58H3AsO4 2H 2e H3AsO3 H2O 9.5 0.56SO4

2 10H 8e H2S 4H2O 5.3 0.31Sb(OH)2

2H 3e Sb 2H2O 3.6 0.21CO2 8H 8e CH4 2H2O 2.9 0.17Sb2O3 6H 6e 2Sb 3H2O 2.5 0.15H3AsO3 3H 3e As 3H2O 4.2 0.25Sb 3H 3e SbH3 8.6 0.51As 3H 3e AsH3 4.0 0.61a pe0 Eh0/0.059 V at 25 C and Eh0 is the equilibrium redox potential relative to

the oxidation of H2(g).

Inorganic species predominate over organic species for both Sband As in most environmental systems (Andreae et al., 1981, 1983;Sun et al., 1993; Ellwood and Maher, 2002). The acid ionisationconstants (pKa) of the oxoacids of Sb and As are listed in Table 3. Themost stable inorganic As (V) species in an Aswater system overa typical environmental pH range (49) are H2AsO4

and HAsO42.The arsenate species are directly analogous to correspondingphosphate species, having the same structure (tetrahedral) andsimilar pKa values (Aylward and Findlay, 1994). The dominant As(III) species is the uncharged pyramidal arsenous acid, As(OH)3,which, unlike its strongly acidic tetrahedral analogue phosphonicacid (H3PO3), is a very weak acid (pKa 9.22). The lack of charge onthe As (III) species compared with the successive deprotonation ofthe As (V) species implies less potential for charge dependent

Vink, 1996), while the scavenging of As by Fe suldes occurs atlower concentrations (Bowen, 1979).

Data for Sb are more limited but have been reviewed by Filellaet al. (2002b). The pepH diagram of the SbOH system proposedby Brookins (1988) indicated that even at activities as low as108 M, Sb would precipitate as oxides across a range of pepH,with a soluble form being limited to more oxidising conditions atlow pH. Modication by Vink (1996) who included the oxyanionicform SbO3

(Sb(OH)6) in the calculations showed that the solubilityof Sb (V) increased in mildly acidic to alkaline conditions (Fig. 1b).For Sb (III), Sb2O3 solubility is independent of pH (Filella et al.,2002b), consistent with the presence of uncharged aqueousSb(OH)3. When Sb(OH)3(aq) is included, solubility again increasescompared to earlier pepH diagrams. As a result of these modi-cations, Sb stability predictions more closely alignwith those for Asand suggest that Sb (III) may be more soluble over a wider pepHrange than previously suspected. This is supported by observationsof dissolved Sb (III) in aqueous systems (Filella et al., 2002b).

If sulfur is included in the modelling, pepH diagrams suggestthat like As, suldes are the predominant Sbminerals formed underreducing conditions (i.e. stibnite, Vink, 1996), and that soluble Sb(III) complexes such as SbS2

may dominate in alkaline reducingconditions (Takayanagi and Cossa,1997). Antimony (V) as SbS4

3 hasalso recently been observed under anoxic conditions (Filella et al.,2002b). Chloride complexes of both Sb (V) and Sb (III) have beenformed experimentally at relatively high chloride concentrations(Filella et al., 2002b), but their presence has not been conrmed attypical environmental concentrations. Chloride complexing is

2 3 4 5 6 7 8 9 10pH

Fig. 1. pepH diagram for the AsOH system (a), and SbOH system (b), assumingSAs and Sb 0.1 mM (c) crystalline, (g) gas. All other species are aqueous. Possibleoxide and hydroxide precipitates have been removed. Dashed lines indicate environ-mental redox limits imposed by the dissociation of water to H2 (g) and O2 (g).Computed using HYDRA (Puigdomenech, 2004) and MEDUSA (Puigdomenech, 2002)for windows.

S.C. Wilson et al. / Environmental Passociations with solid phases, such as clay minerals and oxy-hydroxides in soils. A generalisation made from this is that As (V)species are less mobile than As (III) species in many environmentalsystems (Bhattacharya et al., 2002).

The coordination of Sb (V) with oxygen is octahedral and differsto that of As (V) and P (V) due to a larger ionic radius and lowercharge density (Pauling, 1933). Thus, the deprotonated form ofantimonic acid is the octahedral antimonate ion, Sb(OH)6

. This isthe dominant Sb (V) form over most environmentally relevant pHvalues, in contrast to the successive deprotonation steps observedfor arsenic acid over a similar pH range (Table 3). The most abun-dant Sb (III) species in the Sbwater system is the unchargedantimonous acid Sb(OH)3 with pKa 11.9 (Zakaznova-Iakovlevaand Seward, 2000). So, like As, this indicates that the mobility of Sb(III) may be higher than Sb (V). While successive protonation withpH complicate arsenate binding to particulatematter in oxygenatedsystems, the inuence of pH-dependent binding of antimonate onsolid phases should be simpler, depending primarily on the prop-erties of the binding surface.

The interactive effects of redox and pH on Sb and As speciespresent at equilibrium can be illustrated using simple Eh(pe)pHdiagrams (Vink, 1996; Takayanagi and Cossa, 1997). Such diagramsare limited by the number of chemical elements and speciesincluded, and the availability of thermodynamic data. Fig. 1 illus-trates the simplied AsOH and SbOH systems under redox andpH conditions found in typical earth surface aquatic systems. Fig. 1ashows that under oxidising conditions As (V) is the predominant Asform. Under moderately reducing conditions across a wide pHrange, As exists as the uncharged aqueous species As(OH)3. Inslightly more complex (and realistic) pepH diagrams, As iscoprecipitated with Fe oxyhydroxides, for example, as the hydratediron arsenate mineral scorodite (FeAsO4$2H2O) (Mok and Wai,1990). In the presence of S and at sufciently high As concentra-tions, suldes of As (III), such as arsenopyrite, orpiment or realgarare the most stable form under reducing conditions (Bowen, 1979;

Table 3Equations and pKa values for inorganic As and Sb species (Smith et al., 1998;Zakaznova-Iakovleva and Seward, 2000; Puigdomenech, 2002).

Arsenic (V) pKa

H3AsO4(aq) H2O(l) H2AsO4(aq) H3O 2.20H2AsO4

(aq) H2O(l) HAsO42(aq) H3O 6.97

HAsO42

(aq) H2O(l) AsO43(aq) H3O 11.53Antimony (V)Sb(OH)5(aq) 2H2O(l) Sb(OH)6(aq) H3O 2.72Arsenic (III)H3AsO3 (aq) H2O(l) H2AsO3(aq) H3O 9.22H2AsO3

(aq) H2O(l) HAsO32(aq) H3O 12.13

HAsO32

(aq) H2O(l) AsO33(aq) H3O 13.4Antimony (III)

Sb(OH)3 (aq) 2H2O(l) Sb(OH)4(aq) H3O 11.9-17

-12

-7

-2

3

8

13

2 3 4 5 6 7 8 9 10pH

pe

H2AsO

3

-

As (c)

H3AsO

3

H3AsO

4

H2AsO

4

-

HAsO4

2-

AsH3 (g)

-17

-12

-7

-2

3

8

13

pe

Sb (c)

Sb(OH)3

Sb(OH)5

Sb(OH)6

-

a

b

ollution 158 (2010) 11691181 1171thought to play a minimal role in seawater (Strohal et al., 1975).

Lack of thermodynamic data on Sb complexes with chloride orsulde prevents predictions of speciation. Despite this, and as thepepH adaptations in Fig. 1 show, at environmentally relevantconcentrations and conditions Sb can exist as a soluble species innatural systems regardless of oxidation state (Takayanagi andCossa, 1997). As with As, the relatively high solubility of Sb speciessuggests that sorption on soil surfaces are potentially moreimportant than precipitation reactions in controlling environ-mental mobility.

2.2. Other inuences on inorganic speciation

Speciation in biogeochemical systems, typically not at equilib-rium, is however more complicated than a simple examination of

(III) oxidation to Sb (V) also decreases with pH (Belzile et al., 2001).Abiotic oxidants may also play a role (Quentel et al., 2006). Cutter

Attempts to understand Sb methylation have generally beeninconclusive or contradictory (Dodd et al., 1992; Gurleyuk et al.,1997). Despite this, biologicallymediated reduction andmethylationof Sb compounds in Pseudomonas uorescens bacterial cultures andsoil samples has been conrmed (Gurleyuk et al., 1997). The bio-logical production of trimethylantimony under reducing conditions(Jenkins et al.,1998) suggests that biomethylation of Sbmay occur inenvironments such as in waterlogged soils. Conversely, mono-,dimethyl and trimethyl Sb compounds have been detected in oxi-dised seawaters (Andreae et al.,1981) and urban soils (Duester et al.,2005), and Brannon and Patrick (1985) reported that unidentied Sbvolatiles could be lost from sediments regardless of oxygen status.The toxicity of these volatile Sb species is not yet understood andlittle is known about their environmental chemistry.

Methylation of both As and Sb may enhance mobility through

(Sb S ) is the primary Sb ore, with valentinite (Sb O , an oxidation

ise i

As Sb

ple

s, ra

assempd

S.C. Wilson et al. / Environmental Pollution 158 (2010) 116911811172

mineralisation, parent material differences, varying degrees ofanthropogenic inuence, and different sampling strategies.

3.2. Concentrations in contaminated soils

Antimony and As soil concentrations considered anthropogenicenrichment or contamination vary greatly because of variation inbackground concentrations (Table 4) and differences in contami-nation guideline values between countries and regions. Antimonysoil concentrations have been comprehensively reviewed up to2000 (Filella et al., 2002a) although not all the data included in thereview were considered anthropogenic enrichment. We provide inTable 5, in light of all the work since 2000 on Sb contaminated soils,a review of Sb concentrations reported for contaminated soils overthe last 10 years (prior to 2000 only if not reported by Filella et al.,2002a). Sites commonly associated with high As concentrations arenumerous and varied and we do not attempt to include a full list ofdata for these sites here. However, As concentrations have beenincluded in Table 5 where As co-occurs with Sb in the reviewedpublications. Table 5 demonstrates that Sb contamination has most

through studies with pure mineral phases (Huang, 1975; Thana-balasingam and Pickering, 1990; Fendorf et al., 1997).

4.1.1. Adsorption on pure phases4.1.1.1. Silicate clay minerals. Anion adsorption in clay minerals isassociated with broken clay particle edges, commonly throughsurface ligand exchange mechanisms.e.g. MOH H2AsO4 MH2AsO4 OH.

Extent of adsorption is strongly inuenced by pH, the mineralpoint of zero charge (PZC) (Matera and Le Hecho, 2001), and alsothe metalloid species. For example, As (V) has an environmentalsolubility minimum around pH 46 on kaolinite and montmoril-lonite (Mok and Wai, 1990) which corresponds to adsorptionmaxima (approximately pH 5) (Frost and Grifn, 1977; Goldbergand Glaubig, 1988). Arsenite adsorption by these minerals increaseswith increasing pH between 3 and 9 (with a possible maximumadsorption around a pH of 7) (Frost and Grifn,1977). Differences inmaximum adsorption pH for As (III) and As (V) are attributed toprotonation of As species in solution, and the greater As (V)adsorption to bonding strength differences related to soluble

men

oncmrizoprevlevets aanclasrizodplaSbosphrizoace smmSb45 c

S.C. Wilson et al. / Environmental Pollution 158 (2010) 11691181 1173frequently been reported on and around mining and smelting sitesoften co-occurring with As (Ashley et al., 2007) and that relativeconcentrations of bothmetalloids depend on contamination source.

4. Soil retention mechanisms

Both Sb and As can be strongly retained in soils (McLaren et al.,1998; Flynn et al., 2003). Obviously the extent of retention inu-ences the bioavailable and mobile fraction. Many factors impactretention, including soil characteristics and metalloid speciespresent. Understanding metalloid retention processes is funda-mental for understanding biogeochemical cycling and for accuraterisk assessment in different systems. In the following sections wecritically review data available on these processes and relate theinformation to experimentally known phase associations. In eachsection we rst summarise the important ndings for As and thencompare the information with that available for Sb.

4.1. Adsorption

Adsorption is one of the most important As and Sb retentionmechanisms in soil (Dudas, 1987; King, 1988; Peryea, 1998; Smithet al., 1998; Bhattacharya et al., 2002), albeit most often identied

Table 5Antimony and As concentrations in soils contaminated by anthropogenic activities.

Soil concentration range(mg kg1)

Contamination source Com

As Sb

2220 80200 Smelter site From4.465.1 2.5175 Urban soils impacted by lead zinc smelter 0250371 01090 Area surrounding a goldantimony mine A ho

(in 11.9710 Previous mining activities Low 27.715100 Previous mining activities Shoo

116019138600 2.5237 Previous mining activities Soils11651 315986 Abandoned mine area B ho1.840 0.139.4 Antimony mine 300 km upstream Floo

with59176 180554 Antimony mine area approx 2 km upstream Rhiz3713040 7.413610 Previous mining activities A ho 261150 Mining contaminated area, S France Surf114 3517500 Shooting range Switzerland

tal Pabout pH 4 (Pierce and Moore, 1982). Both arsenite and arsenateare adsorbed quickly (90% of added As in less than 2 h) and surfacesorption site densities and binding constants on amorphous ironoxide, goethite and magnetite at optimal pH are similar (Dixit andHering, 2003). Isotherms indicate more than one type of surfacebinding site for both species including inner-sphere surface com-plexing with monodentate and bidentate binuclear and mono-nuclear complexes reported depending on iron surface(Waychunas et al., 1993; Manceau, 1995; Ona-Nguema et al., 2005;Fendorf et al., 1997).

Oxides and hydroxides are also known to be important for Sbadsorption in soil (Gleyzes et al., 2002; Chen et al., 2003; Butleret al., 2005; Manaka, 2006; Mitsunobu et al., 2006). Antimony hasbeen positively correlated with the soil iron oxide component (Galet al., 2006; Denys et al., 2008) and also shown to sorb onto Mnoxides in preference to Fe oxides in lake sediments (Muller et al.,2002). Antimony (III) sorbs strongly to hydrous Mn oxides andsorption decreases in the order MnOOH > Al(OH)3 > FeOOH, withthe amount sorbed decreasing gradually as pH increases above 6.Below a pH of 6, over 80% of added Sb (III) is retained by all 3phases (Thanabalasingam and Pickering, 1990). This is in agree-ment with more recent work where >80% of Sb (III) was adsorbedon goethite between pH 112 (Leuz et al., 2006b). Both specic andnon-specic sorption processes operate similar to As. Maximum Sb(III) adsorption by specic processes has been reported arounda pH of 78 on Al and Fe oxides (Thanabalasingam and Pickering,1990) indicating that non-specic adsorption decreased theadsorption maximum by a pH of 12 units. On goethite, Sb (III),similar to arsenite, forms bidentate, corner-sharing inner-spherecomplexes but is bound more strongly than arsenite for the sameinitial anion and goethite concentrations (Leuz et al., 2006b).Despite the stronger sorption onto Mn hydrous oxides, the highabundance of Fe and Al hydrous oxides in environmental samplesindicates these minerals could determine dissolved Sb (III) innatural systems (Casiot et al., 2007). The oxidative affect of thesesorbents on metalloid speciation and mobility is considered laterin this text.

Sb (V) sorption by Mn oxyhydroxides is not yet understood.Adsorption to Fe hydroxides, however, appears to be strong butmore dependent on pH than that of Sb (III) with maximumadsorption below pH 7 (Leuz et al., 2006b). Tighe and Lockwood(2007) reported 95% Sb (V) sorption by a non-crystalline Fehydroxide across a pH range of 2.57 with a sorption maximum atabout pH 4. Sorption of Sb (V) on hematite is also strong, witha similar sorption maxima (wpH 4) involving a combination ofsurface binding sites, similar to As (Pierce and Moore, 1982; Ambe,1987). Antimonate, however, is less strongly sorbed to goethitethan As (V) over a wider pH range (Leuz et al., 2006b) and this mayimpact importantly on mobility and bioavailability in specicenvironments. Specic Sb (V) adsorption on Fe hydroxides in theform of mixed inner-sphere complex formations has beenconrmed in contaminated shooting range soils and on goethite(Leuz et al., 2006b; Scheinost et al., 2006).

4.1.1.3. Organic matter. Although sorption on natural organicmatter is considered responsible for As retention in several studies,these are limited and the percentage retained low (Jones et al.,1997; Lintschinger et al., 1998). The only comparative study of As(III) and As (V) sorption by humic acid (Thanabalasingam andPickering, 1986) reported maximum adsorption of both As (III) and(V) at around pH 5.5. Adsorption decreased with additional pHincreases (some experimental discrepancies raise doubts about thecalculated quantitative maximum adsorption reported of 70 and90 mmol kg1 for As (III) and As (V), respectively). Enhanced

S.C. Wilson et al. / Environmen1174deprotonation and competition with OH at higher pH wasproposed to explain results, although physical loss of binding siteswith humic acid dissolution at higher pH values may also haveoccurred. Electrostatic attractionwas also proposed as an operatingmechanism with displacement of adsorbed As by a range ofother anions.

Antimony associationwith soil organic matter in environmentalsamples has only been conrmed in recent years (Clemente et al.,2008; Ceriotti and Amarasiriwardena, 2009). Neutral Sb (III) specieslike Sb(OH)3 will readily bind to humic acids, with up to 30% of totalSb (III) bound at environmentally relevant conditions (Filella et al.,2002b; Buschmann and Sigg, 2004). The Sb (III) binds comparativelymore strongly thanAs (III). In highly organic acid soils under specicconditions Sb (III) was found to account for up to 34% of total Sbpresent (Ettler et al., 2007) and interactionwith the humic materialcould signicantly impact mobility. Adsorption maxima for Sb (III)on humic acid of 23 mmol Sb g1 and 53 mmol Sb g1 have beenreported using inorganic (Sb(OH)3) and organic (antimonyl tartrate)solutions, respectively, at pH 4 (Pilarski et al., 1995). Differenceswere attributed to the added effect of tartrate in chelating andionising Sb (III) as an anionic complex at this pH. For the inorganic Sb(Sb(OH)3), adsorption was reduced from a maximum of 70% at pH3.8 to 55% by pH increases between 3.1 and 5.4 due to humic acidprotonation effects, competition with other aqueous species, oraqueous complexing of Sb reducing its afnity for the surface athigher pH values (i.e. by organic complexes). Several mechanismsfor Sb (III) binding with humic matter phenolic, carboxylic andhydroxyl-carboxylic groups have been proposed including ligandexchangewith the Sb centre, and formation of negatively charged Sbcomplexes with carboxylic groups (Buschmann and Sigg, 2004;Tella and Pokrovski, 2009). Chelation, H-bridges or cationic metalsmay stabilise the Sb (III) binding.

Retention of Sb (V) by humic acid at Sb concentrations less than10 mmol L1 and at pH values 3.15.4 was not detected by Pilarskiet al. (1995). Conversely, Steely et al. (2007) recently reported thathumic acids in contaminated shooting range soils had a highcapacity for complexing and tightly binding Sb (present predomi-nantly as Sb (V)), trapping most in the soil organic layer. Sb (III) wasoxidised to Sb (V) by humic acid. Tighe et al. (2005b) also reportedsignicant sorption of Sb (V) (56% of added Sb) on humic acid ata concentration of 0.23 mmol L1 and pH 4. Similar mechanisms tothose proposed by Buschmann and Sigg (2004) could be operatingfor the antimonate ion although this has not yet been conrmed.

4.1.2. Adsorption on whole soilsAdsorption in whole soils is less studied than on pure phases.

Arsenic (III) sorption in soils is largely irreversible, and governed bythe presence and abundance of Fe oxides (Elkhatib et al., 1984). Thedominant sorptionmechanism is binuclear bridging complexes andother inner-sphere complexing with oxides and hydroxides asdiscussed previously (Fendorf et al., 1997). Similarly, Wauchope(1975) reported that arsenate binding by alluvial soils was stronglycorrelated with clay and Fe oxide content of the soils but wasindependent of organic matter. In four Australian soils with pHvalues ranging between 4.97 and 6.90 As (V) sorbed was greaterthan As (III) (Smith and Huyck, 1999). Arsenate sorption increasedas pH decreased, while arsenite sorption increased as the pHapproached neutral. Crystalline and non-crystalline Fe oxides wereimportant for arsenate adsorption, and the proportions of both As(V) and As (III) sorbed approached 100% in oxidic soils at optimumpH values for adsorption. Oxidation of As (III) to As (V) by Fe or Mnminerals was proposed as a possible mechanism affecting the As(III) adsorption proles although this was not conrmedexperimentally.

Few studies on whole soils exist for Sb. In thirteen U.S. surface

ollution 158 (2010) 11691181and subsoils, 50100% of added Sb (III) (26.4 mM as antimony

tal Ppotassium tartrate) was sorbed and, like As (III), a high proportion(5799%) of the sorbed component was non-exchangeable, asdetermined by a 1 M KCl extract (King, 1988). Sorption increased asthe silt and clay fraction increased, although no relationshipbetween sorbed Sb and the non-crystalline or crystalline oxidephases in the soils was detected. The relative sorption of Sb washigher in mineral soils compared to organic rich soils. Tighe et al.(2005b), however, reported over 75% sorption of Sb (as Sb (V))added to two acidic, organic rich soils (33.6% and 26.2% organicmatter) with 80100% sorbed at pH < 6.5. Maximum sorptionreected the PZC differences between the soils.

To summarise, based on the discussion above, Sb associationwith silicate clays may be important in certain environmentsdepending on the Sb source but requires further study. In manynatural environments adsorption on the Mn and Fe oxyhydroxidesis responsible for retention of a high proportion of the soil boundSb. However, the extent of adsorption may be greater or less thanthat of As in the same environment depending on species presentand sorptive surface. This will impact bioavailability and mobility.Furthermore, the retention of Sb by organic matter may be moreimportant than originally thought, based on As data. Mechanismsof Sb interaction with organic phases are still largely unexploredand the effects of protonation and specic adsorption in acidic soilsnot assessed. Further Sb sorption studies on a range of differentsoils are needed.

4.2. Effects of co-oxidants/co-reductants

Oxidation of Sb (III) to Sb (V) in the presence of some hydrox-ides, similar to that for As, has been reported. The predominance ofSb (V) in soils is partly attributed to this co-oxidation mechanism(Leuz et al., 2006b) and also the association of Sb (V) with Fehydroxides even under reducing conditions (Mitsunobu et al.,2006). The rapid oxidation of Sb (III) to Sb (V) by amorphous formsof Fe and Mn oxyhydroxides in the pH range 510 suggests that thepresence of these materials in soils may play a detoxifying role forsoil adsorbed Sb (Belzile et al., 2001; Leuz et al., 2006b). Ironmediated oxidation of Sb (III) is approximately 10 times faster thanthat of As (III) (Leuz et al., 2006a). However, whether the Sb (V)remains adsorbed to the mineral surface or is mobilised has beenlittle studied. The one study reported to date, shows that althoughoxidation of Sb (III) on goethite did not mobilise Sb within 35 daysbelow pH values of 7, at pH 9.9, 30% of adsorbed Sb (III) wasreleased into solution within the same time (Leuz et al., 2006b).Therefore, in alkaline soils the co-oxidation process could in factcontribute to mobilisation of Sb species.

Green rusts, (layered Fe (II)Fe (III) hydroxides) are formed insuboxic and mildly alkaline environments, (likely to be animportant Fe mineral in immature sediments and soils) and havea high afnity for Sb (V). It is known that green rusts do notreduce As (V) to As (III). However, Sb (V) is reduced to Sb (III)despite the high stability of the Sb (V) species (Mitsunobu et al.,2008) which could inuence Sb mobility in suboxic environments.To fully understand Sb geochemical cycling there is a need tobetter understand the effects of co-oxidants and reductants inspecic environments.

4.3. Effects of ionic strength and competition

Ionic strength effects on anion adsorption depend on the soilPZC, and whether the adsorption mechanism is specic (inner-sphere complexing) or non-specic (outer-sphere complexing).Specic adsorption is often assumed when there is little change insorption with ionic strength (Leuz et al., 2006b). If outer-sphere

S.C. Wilson et al. / Environmencomplexes are involved then increasing ionic strength decreasesthe negative charge of the surface at higher pH values andadsorption would increase as observed for As (V) (Smith et al.,1999). Below the PZC, increasing ionic strength decreases positiveelectrostatic potential close to the surface resulting in lessadsorption (van Olphen, 1963; Bowden et al., 1980). However,explanations are not always this simple. For example, increasingionic strength (using 0.01 and 0.1MKClO4 solutions) lowered Sb (V)adsorption on goethite above pH 6. The results could not beexplained by outer-sphere modelling nor a combination of bothinner-sphere and outer-sphere but were explained by inner-spherecomplexing and formation of ion pairs of KSb(OH)6 in solution(Leuz et al., 2006b). Ionic strength had no inuence on Sb (III)sorption over the wide pH range of 112 where inner-spherecomplexing was conrmed. However, addition of acetate bufferreduced Sb (III) sorption on hydrous Al, Mn and Fe oxides and wasattributed to a decrease in surface ion exchange (Thanabalasingamand Pickering, 1990). The effect of reduced ion exchange onuncharged Sb(OH)3 was not considered, nor the effect of Sb(OH)3aqueous complexing with the buffer. Specic and non-specicadsorption mechanisms could in fact co-occur and the effect of onenegate the effect of the other.

Competition with other inorganic or organic anions may alsobe important. Whilst chloride, nitrate and sulfate show little effecton As (V) sorption at concentrations typical of saline soils (Liveseyand Huang, 1981), phosphate (and sulphate in some studies)competitively desorb or reduce As (V) adsorption by soils(Alvarez-Benedi et al., 2005), pure mineral phases and Fe(OH)3(Livesey and Huang, 1981; Manning and Goldberg, 1996; Smithet al., 2002; Waltham and Eick, 2002). This competitive effect hasalso been observed for As (III) on iron oxides (Dixit and Hering,2003). The competition is reduced in soils with high free Fe oxidecontents (a Vertisol and Oxisol), but this retardation is lessobvious as P additions increase (Smith et al., 2002). Competitionappeared to be limited by the sorption sites available (Walthamand Eick, 2002).

Less is known about competitive effects on Sb adsorptionalthough Kilgour et al. (2008) recently reported signicant releaseof both As and Sb from contaminated ring range soils where P wasadded as superphosphate. Soil-solution distribution coefcients forSb (as spiked equilibrated 124Sb (V)) on 110 Japanese agriculturalsoils (Andosols, Fluvisols, Cambisols and Regosols 59 upland soils,51 paddy soils) decreased with increasing phosphate concentrationand was attributed to direct competition for specic binding sites(Nakamaru et al., 2006). Only 2040% of sorbed Sb was in a phos-phate exchangeable form. In a later study again using Japanese soilsonly 0.21.3% of total native Sb was in a phosphate exchangeableform (Nakamaru and Sekine, 2008). Sb mobilised by phosphateadditions seems to depend on the form present with the extent ofligand exchangeable Sb being important. Sb (III) was less extract-able than Sb (V).

Competition with organic acids has not been reported for Sb.However, As (V) adsorption on ferrihydrite decreased in thepresence of citric acid across a wide pH range (311), but not withfulvic acid, while both citric and fulvic acids decreased As (III)adsorption (Grafe et al., 2002). Humic acid had no effect on eitherspecies. Differences were attributed to the low afnity of dissolvedorganic carbon for ferrihydrite surface sites, and the increasedrelative binding of carboxyl groups in the lower molecular weightcitric and fulvic acids. Conversely, Saada et al. (2003) found thathumic acid coatings on kaolinite at a pH of 7 increased As (V)adsorption at low initial As concentrations and concluded thehumic acid had a higher afnity for As than the kaolinite, with theprotonated amine groups being responsible for the increased Asadsorption, similar to the anion exchange mechanism proposed for

ollution 158 (2010) 11691181 1175As sorption by humic acids (Thanabalasingam and Pickering, 1986).

tal PSaada et al. (2003) have proposed that the soil organic mattercontent may govern As adsorption if the hydrous Fe oxide contentis low. The interaction of dissolved As with organics such ashumates and fulvates remains unclear. In light of these ndings,understanding competition interactions remains an area ofresearch need for Sb.

4.4. Precipitation, co-precipitation and dissolution

Metalloid mobilisation in mineralised areas and areas affectedby mining and smelting frequently occurs through the oxidationand dissolution of As and Sb bearing primary minerals (Ashley andLottermoser, 1999; Ashley et al., 2003, 2006; Baron et al., 2006).Dissolution, however, is often accompanied by an abrupt change inenvironmental conditions leading to removal of the mobilemetalloids from solution. For example, scorodite (FeAsO4$2H2O) isknown to precipitate following dissolution of As bearing suldes, ifconditions are acidic and oxidising (Vink, 1996). Subsequentdissolution of scorodite releases As, the dissolved concentrationthen being controlled by sorption on freshly formed hydroxidessuch as Fe(OH)3 (Zhu and Merkel, 2001). Similarly, dissolution ofstibnite (Sb2S3) has been shown to produce up to 55 mg L

1 of Sb insolution at active mining sites, although some attenuation occursdue to adsorption and co-precipitationwith amorphous hydroxides(Ashley et al., 2003). Adsorption or precipitation of As with Caphases in neutral to alkaline calcareous soils is also important(Matera and Le Hecho, 2001). Similarly, in a highly contaminatedalkaline soil, Johnson et al. (2005) found the solid Ca(Sb(OH)6)2 wascontrolling the concentration of dissolved Sb (V). In general,precipitation and co-precipitation processes are only important inareas of gross contamination where As and Sb concentrations arehigh enough to initiate precipitation, secondary As or Sb mineralsoccur and control dissolved concentrations. In soils with lower totalmetalloid concentrations, adsorption mechanisms as described inprevious sections of this review govern mobility (Matera and LeHecho, 2001).

5. Phase associations

Single or sequential chemical soil extractions are frequentlyused to infer As or Sb association with specic phases such asminerals or organic matter. Extractants become progressivelystronger for assessment of more residual forms. The informationcan provide useful information on potential metalloid availabilityor mobility under specic environmental conditions. The ongoingdifculties and controversies associated with the use of suchchemical extractions will not be discussed here, but are summar-ised succinctly in several reviews (Hirner, 1991; Gleyzes et al.,2002; Hill et al., 2002; Butler et al., 2005). While the use ofsequential extraction procedures has limitations compared todirect speciation measurements, the additional phase informationand lower cost has seen their continued use and development(Butler et al., 2005). In this section we review several single andsequential extractions used for As and Sb to understand betterphase associations in whole soil systems. The single extractions aresummarised in Table 6. Table 7 describes two common sequentialextraction procedures used for As and Table 8 details sequentialextraction schemes used for Sb (or both metalloids together)where the proportion of metalloid extracted in each fraction hasbeen reported. All experiments cited have used eld unspiked soilsexcept where indicated. It is important to recognise the difcultiesand complexity of comparing data obtained from different exper-iments with different soils extracted using different chemicalextractants as illustrated by the work reported by Tighe and

S.C. Wilson et al. / Environmen1176Lockwood (2007) on moderately contaminated oodplain soils.Nevertheless, the data in Tables 68 provide useful information oncommonly used extractants for As and Sb and phase associationpatterns.

5.1. Dissolved, soluble and labile concentrations

Dissolved or soluble As and Sb concentrations can be high ingrossly contaminated soils, but typically contribute a smallpercentage of the total metalloid concentration in the soil (99% of water-soluble Sb (up to 5 mg L1) fromcontaminated shooting range soils (up to 13.8 g kg1 soil Sb) andLintschinger et al. (1998) detected only Sb (V) in the water-solublefraction of oxidised soils. A recent comparative study using Japa-nese mine contaminated soils showed, however, that the depen-dence of dissolved concentration with depth or Eh was verydifferent for both metalloids with redox controlling those concen-trations (Mitsunobu et al., 2006). Sb (V) was the stable oxidationstate present even under reducing conditions (Eh 180 mV, pH8), (conrming this as the stable Sb form in the environment), butAs was reduced from As (V) to As (III) as reducing conditionsdeveloped. As a result dissolved Sb concentrations did not increasewith decreasing Eh as observed for As.

Exchangeable or weakly bound As and Sb (often determinedwith a mild extract such as ammonium nitrate) also typicallycontributes only a small proportion to the total soil concentration(Table 6). Freely exchangeable and labile As is typically

Table 6Metalloid fractions in soil extracted using single extraction procedures (all As and Sb concentrations are eld values except where indicated).

Targeting Extract Metalloid Total soilconcentration(mg kg1)

Fractionextracted (%)

Soil details & As/Sb source Reference

Dissolved Extracted soil solution As 66 0.005 pH 56, SwedishCambisol

Tyler and Olsson (2001)

545 0.61.8 pH 4.9 (spiked) Onken and Hossner (1996)Sb 0.40

Table

8Sequ

ential

extraction

proceduresforSb

insoil(dataforAshas

beenincluded

when

also

reported)(soilsnot

spiked

).

Pool

targeted

Den

yset

al.(2008)

Urban

soilscontaminated

bymining

TotalSb

26.161.8

mgkg

1

pH7.17.8

He(2007)

Nineminecontaminated

soils

TotalSb

97.14489.9

mgkg

1

pH4.657.14

Tigh

ean

dLockwoo

d(2007)

Twooo

dplain

soils

TotalSb

2326.9

mgkg

1

TotalAs22.530.5

mgkg

1

pH3.555.40

Hou

etal.(2006)

Five

Japan

esesoils

TotalSb

0.351.2

mgkg

1

pH5.26.58

Extract

%Sb

extracted

Extract

%Sb

extracted

Extract

%extracted

Extract

%Sb

extracted

SbAs

Soluble

Water

511

Water

0.092.5

Freely

exchan

geable

Mg(NO3) 2(1

M)

615

MgC

l 2(1

M)

0.432.11

NaH

CO3(0.5

M)

27

713

NH4NO3(1

M)

approx.5

Acid-soluble

(carbon

ates)

Sodium

acetate/aceticacid

820

Sodium

acetate/aceticacid

(1M)

0.191.85

CH3COONH4CH3

COOH(1

M)

6.1

Man

ganeseox

ides

Hyd

roxylammon

ium

chloride

615

Hyd

roxylammon

ium

chloride

(0.4

M)in

25%aceticacid

0.523.39

NH2OH(0.1M)easily

reducible

metalox

ide

6.8

Amorphou

siron

oxides

Ammon

ium

oxalate(0.2

M)/

oxalicacid

(0.2M)

615

Included

withabov

eextract

Oxalicacid

(0.2

M)/

ammon

ium

oxalate

3047

5071

(NH4) 2C2O4(0.2M)

16

Crystallineiron

oxides

Ammon

ium

oxalate(0.2

M)/

oxalicacid

(0.2M)/

ascorbicacid

(0.1M)

410

Included

withabov

eextract

(NH4) 2C2O4(0.2M)/

ascorbicacid

(0.1

M)

8.9

Chem

isorbed

toFe

andAl

NaO

H(0.1M)

1216

1322

Organ

icmatter

H2O235%

24

HNO3(0.02M)in

30%

H2O2

pH2,N

H4OAc(3.2M)in

20%HNO3

0.914.59

H2O230%

8.7

Na 4P 2O7(0.1M)For

M-com

plexes

18

Sulphide

Nitricacid

7N

24

Included

withabov

eextract

Residual

orhighly

recalcitrant

HFHNO3

1145

HNO3HClO

4HF

88.297.9

HClHNO3

3749