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Applied Geochemistry 23 (2008) 955–976

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AppliedGeochemistry

Review

Leaching mechanisms of oxyanionic metalloid and metalspecies in alkaline solid wastes: A review

Geert Cornelis a,, C. Anette Johnson b, Tom Van Gerven a, Carlo Vandecasteele a

a Laboratory of Applied Physical Chemistry and Environmental Technology, K.U. Leuven, W. De Croylaan 46, B-3001 Leuven, Belgiumb Swiss Federal Institute for Environmental Science and Technology (EAWAG), Box 611, Ueberlandstrasse 133, CH-8600

Dubendorf, Switzerland

Received 13 November 2006; accepted 12 February 2008Editorial handling by R.N.J. Comans

Available online 10 March 2008

Abstract

An overview is presented on possible mechanisms that control the leaching behaviour of the oxyanion forming elementsAs, Cr, Mo, Sb, Se, V and W in cementituous systems and alkaline solid wastes, such as municipal solid waste incineratorbottom ash, fly ash and air pollution control residues, coal fly ash and metallurgical slags. Although the leachability ofthese elements generally depends on their redox state, speciation measurements are not common. Therefore, experimentalobservations available in the literature are combined with a summary of the thermal behaviour of these elements to assesspossible redox states in freshly produced alkaline wastes, given their origin at high temperature. Possible redox reactionsoccurring at room temperature, on the other hand, are reviewed because these may alter the initial redox state in alkalinewastes and their leachates. In many cases, precipitation of oxyanions as a pure metalate cannot provide a satisfactoryexplanation for their leaching behaviour. It is therefore highly likely that adsorption and solid solution formation withcommon minerals in alkaline waste and cement reduce the leachate concentration of oxyanions below pure-phasesolubility.� 2008 Elsevier Ltd. All rights reserved.

Contents

0d

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9562. Redox chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 957

883-2oi:10.

CorE-m

2.1. Possible redox states . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9572.2. Predominant redox states in alkaline wastes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9572.3. Oxidation and reduction processes in alkaline wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 961

3. Solubility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 963
3.1. Metalate precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9633.2. Adsorption and solid solution formation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 966
927/$ - see front matter � 2008 Elsevier Ltd. All rights reserved.1016/j.apgeochem.2008.02.001

responding author. Tel.: +32 16 322343; fax: +32 16 322991.ail address: [email protected] (G. Cornelis).

mailto:[email protected]

956 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976

4. Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 970References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 970

1. Introduction

Interest in the leaching behaviour of As, Cr, Mo,Sb, Se, V and W has been growing over the lastyears. Most of these elements are redox sensitiveand some oxidation states can form oxyanions(negatively charged species containing O) in solu-tion, forming a range of different species dependingboth on pH and redox potential. Past research ini-tiatives and the legislation in waste managementhave focused on contaminants of high concentra-tion and toxicity such as Cu, Cd, Hg, Pb and Znwhereas oxyanionic species have received consider-ably less attention because of their much lowertotal solid phase concentrations. However, theyare often found in relatively high concentrationsin leachates compared to the cationic species dueto their high solubility. Recently, the EuropeanDirective, 1999/31/EC on landfilling of waste hasbeen complemented by the Council Decision2003/33/EC establishing criteria and proceduresfor the acceptance of waste at landfills. Theseinclude limit values for As, Cr, Se, Mo, and Sb.In the Netherlands, V and W are also regulated(Van Gerven et al., 2005). In order to comply withlegislation, it is necessary to develop techniques tocontrol oxyanion leachability. Therefore, a betterunderstanding of the geochemical mechanisms thatcontrol their leaching behaviour, particularly ofMo, W, Sb and V, is required.

Most if not all alkaline wastes originate fromhigh temperature processes with thermal treatmentof waste, fossil fuel combustion (FFC) and ferrous/non-ferrous metal smelting being the most impor-tant ones in terms of waste production. Annually,about 100 million tons of municipal solid wasteincineration (MSWI) residues are produced in Eur-ope, USA and Japan together (Geysen, 2004; Mill-rath et al., 2004; Tanaka et al., 2004). Worldwidemore than 600 million tons of fossil fuel combus-tion residues are produced every year (Brennanet al., 2002; Kalyoncu, 2002). Ferrous and non-fer-rous industries also produce several hundreds ofmillion tons of residues annually (USGS, 2005).In addition, wastes containing high toxic metalconcentrations are often stabilised with cementi-

tuous matrices, creating a highly alkalineenvironment.

Although significant differences exist amongstthese waste types due to their different origin,some striking similarities exist. Freshly producedalkaline wastes have a narrow pH distribution(between 10 and 13) (van der Sloot et al., 1997)because the leachate pH is mainly controlled bydissolution of a limited set of minerals containingCa such as portlandite (Ca(OH)2), monosulphate(Ca4[Al(OH)6]2SO4 � 13H2O), hydrocalumite (Ca4-[Al(OH)6]2(OH)2 � 6H2O), ettringite (Ca6[Al(OH)6]2-(SO4)3 � 32H2O), Ca silicate hydrate (CSH) andcalcite (CaCO3), that can be found in all thesewastes (Warren and Dudas, 1985; Johnson et al.,1995; Meima and Comans, 1997; Barna et al.,2000b). The quantities of these minerals, however,may vary as reflected by the acid neutralisationcapacity (ANC). Table 1 shows a wide range ofANC values for metallurgical residues. Blast fur-nace and other ore-based slags are relatively inertmaterials and have a very low ANC (Barna et al.,2004), which is quickly depleted upon oxidation ofthe relatively large amount of sulphides present inthese slags (Fallman and Hartlen, 1994). Steel slagon the other hand contains an appreciable amountof portlandite and calcite (Ettler et al., 2004; Shenet al., 2004; Huijgen and Comans, 2006). It will beargued that in particular the minerals containingCa mentioned above exert control on oxyanionleaching. As hydrated ordinary Portland cement(OPC) is almost exclusively composed of theseminerals (Glasser, 1997), similar leaching trendsmay thus be expected.

Table 2 gives an overview of the total concen-tration of As, Cr, Se, Mo, Sb, V and W in alka-line waste types. Although the values vary widely,it can be recognised that some elements are morehighly enriched in particular waste types than oth-ers. This can evidently be explained by the sourceof a given waste. Fossil fuels, for instance, gener-ally contain much more Se (0.4–8 mg/kg) thanmunicipal solid waste and the scrap or ore fromwhich Fe or precious metals are produced,generally contain more Cr and V than MSWand fossil fuels. Even more important is the

Table 1Most important characteristics of alkaline wastes

MSWI bottom ash MSWI APC residues FFC fly ash Metallurgical residues

Temperature of formation (�C) 700–1100 �Ca 150–250 �Ca 500–600 �Ca >1100 �CANCpH7:5

u 0.5–1.2 equiv./kg bcd 3–13 equiv./kg akl 0.1–5.5 equiv./kgno 0.01–1.8 equiv./kgrs

Major reducing species

Non-combusted organic matter 1–5%cef NA NA NAFe0 9–15%ag 3–6%a NA 1–25%s

FeII-compounds 2–3%h NA NA NAAl0 1–2agi % 0.012–0.024%m 0.018 – 0.032 %pq 25–75%t

S�II 0.05–0.1%j NA NA 2–6%s

NA = data not available.a IAWG (1997).b van der Sloot et al. (2001).c Johnson et al. (1995).d Meima and Comans (1999).e Zevenbergen (1994).f Krebs et al. (1988).g Wiles (1996).h Zeltner (1998).i Calculated from Bertolini et al. (2004).j Redle (1992).

k Barna et al. (2000a).l Geysen (2004).

m Astrup et al. (2005).n Tiruta-Barna et al. (2006).o Lo and Liao (2007).p Aubert et al. (2004).q Calculated from Cai et al. (2003).r Barna et al. (2000b).s Barna et al. (2004).t Hazar et al. (2005).u Alkalinity or the average amount of acid to reach pH 7.5.

G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976 957

behaviour of the elements during thermal treat-ment. In MSWI and FFC residues, the elementsAs, Sb and Se form volatile compounds at rela-tively low temperatures and are more partitionedinto the fly ash and air pollution control (APC)residues than Cr, Mo, V and W (IAWG, 1997;Belevi and Moench, 2000).

2. Redox chemistry

2.1. Possible redox states

Table 3 shows the possible redox states of As,Cr, Se, Mo, Sb, V and W and their forms of occur-rence in alkaline conditions. The oxyanions show ahigh tendency to polymerise at high concentrationsand low pH (Baes and Mesmer, 1976). Waste poresolutions, however, are highly alkaline so for thesake of convenience, only the mononuclear speciesare discussed. Kinetic hindrance and heterogeneityof wastes cause redox disequilibria, and frequently

different redox states of the same element can befound together in waste leachates. Several authorshave hailed the importance of speciation measure-ments in determining leachability in solid wastesin realistic situations because the different redoxspecies can have a different geochemical behaviour(Dusing et al., 1992; Fallman and Hartlen, 1994;Kosson et al., 1996; van der Sloot et al., 1997; Sab-bas et al., 2003), but limited knowledge of the exactspeciation of oxyanion forming elements is one ofthe most important limiting factors for efficientmodelling of their leaching behaviour in solidwastes.

2.2. Predominant redox states in alkaline wastes

The relation between the speciation of an elementfound in waste pore waters on the one hand and thesolid phase on the other evidently depends on thedifferent geochemical behaviour of separate species.This literature review mainly focuses on more

Tab

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gin

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resi

du

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allm

anan

dH

artl

en,

1994

;IA

WG

,19

97;

van

der

Slo

ot

etal

.,19

97;

US

GS

,20

05),

FF

C-

resi

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ary

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.,19

90;

van

der

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ot

etal

.,19

97;

US

GS

,20

05)

and

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rgic

alre

sid

ues

(Fal

lman

and

Har

tlen

,19

94;

Par

son

set

al.,

2001

;P

iata

ket

al.,

2004

;S

hen

etal

.,20

04;

Alt

er,

2005

)

Ele

men

tL

ith

osp

her

eS

oil

sM

SW

Ire

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ues

FF

C-r

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ues

Met

allu

rgic

alsl

ags

Bo

tto

mas

hF

lyas

hA

PC

a

resi

du

esC

oal

bo

tto

mas

hC

oal

fly

ash

FG

Db

ash

Bla

stfu

rnac

esl

agS

teel

slag

No

n-f

erro

us

slag

s

As

51–

500.

1–20

040

–300

20–5

000.

02–2

002–

400

0.8–

50<

0.7

50.

2–2

Cr

200

1–10

0020

–300

010

0–10

0070

–700

0.2–

6000

4–90

02–

200

3080

00–3

0,00

020

–300

Mo

20.

2–5

2–30

015

–200

2–40

1–50

01–

100

1–50

<6.

020

3–50

Sb

0.2–

0.5

–10

–400

300–

1000

80–1

000

NA

NA

NA

NA

NA

<0.

1–0.

3S

e0.

090.

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0.05

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0.4–

300.

7–30

0.1–

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2–10

02–

200

NA

NA

2–6

V20

020

–500

20–1

0030

–200

8–90

10–5

0010

–100

08–

400

400

1000

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30N

AN

AN

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.


oxidised species (AsV, AsIII, CrVI, SeIV, SeVI, SbV,MoVI, VV and WVI) because they are found morefrequently in waste leachates (Table 4) whereas ele-ments occurring in their elemental state as well asmany hydroxides and oxides formed by reducedspecies at high pH (CrIII, SbIII, VIII and VIV) areonly slightly soluble (Wanty and Goldhaber, 1992;Beverskog and Puigdomenech, 1997; Seby et al.,2001; Filella and May, 2003).

Table 4 shows that speciation measurements arescarce, except in the case of coal fly ash leachates.Due to the indirect relationship between aqueousand solid speciation, leaching predictions may notbe based solely on pore water speciation measure-ments and knowledge of solid phase speciation isrequired but direct measurements are even scarcer.Table 5 lists the mineral compounds of oxyanionsexperimentally detected in wastes but the redoxstates in Table 5 give an incomplete picture of whatcan be found in alkaline wastes because trace oxya-nions are usually physically scattered and most oftheir compounds are therefore hard to detect. Gen-eral trends as summarised in Table 6 were thereforededuced from the thermochemical behaviour exhib-ited during formation of the considered wastes.

Nearly all As and Sb is reduced at temperatures>500 �C to volatile gaseous states that escape frombottom ash. If they do remain in the ashes their oxi-dation is catalysed by abundant metal oxides fol-lowed by absorption to form the correspondinginvolatile metal arsenate or antimonate (Mahuliet al., 1997; Paoletti et al., 2001; Paoletti, 2002; Ster-ling and Helble, 2003) and the highest oxidationstates of As and Sb therefore predominate in bot-tom ash and its leachates. VV and CrIII, on the otherhand, are known to be highly involatile (Belevi andMoench, 2000; Paoletti, 2002) and reaction withmetal oxides, without oxidation, lowers their vola-tility even further (Paoletti, 2002; Diaz-Somoanoet al., 2006). Volatile VIV compounds are oxidisedat temperatures >370 �C (Frandsen et al., 1994),whereas CrIII is relatively resistant to oxidation(Paoletti, 2002). Only a very small amount is oxi-dised to metal chromates at temperatures <850 �Cin the presence of metal oxides (Paoletti, 2002) orvolatilised as hexavalent CrO2Cl2(g) or CrO2(OH)2-(g) (Chen et al., 1998; Linak and Wendt, 1998). CrIII

compounds, however, are sparingly soluble andhence, predominantly CrVI is found in bottom ashleachates (Kersten et al., 1998). It has to be noted,however, that especially bottom ash is a highly het-erogeneous waste type because throughout the

Table 3Redox states of oxyanionic species and their forms of occurrence in alkaline leachates

Element Oxidation state

�II 0 +III +IV +V +VI

As As0 H2AsO�3 AsO3�4

H3AsO04 HAsO2�

4

Cr Cr0 CrðOHÞ�4 CrO2�4

Se HSe� Se0 SeO2�3 SeO2�

4

Mo Mo0 MoO2�4

Sb Sb0 SbðOHÞ�4 SbðOHÞ�6SbðOHÞ03

V V0 VðOHÞþ2 VOðOHÞþ2 VO3�4

W W0 WO2�4

Table 4Experimentally detected redox speciation of oxyanionic species in leachates of alkaline wastes

Waste type As Cr Se Sb W

MSWI bottom ash VI V VI Kersten et al. (1997, 1998) and Cornelis et al. (2006a).FFC fly ash V VI IV V Eary et al. (1990), van der Hoek et al. (1996), Goodarzi and Huggins (2001),

and Galbreath and Zygarlicke (2004), Lecuyer et al. (1996), Iwashita et al.(2005), Wadge and Hutton (1987), Rai and Szelmeczka (1990), and Miravetet al. (2006).

III VI IIIMetallurgical

residuesIII III

VIPillay et al. (2003), Ettler et al. (2004), and Huijgen and Comans (2006).


waste bed, highly oxidised conditions are alternatedwith smaller areas with low oxygen partial pressurewhere reduced gaseous oxides may condense.

Oxidative absorption of gaseous AsIII and SbIII

compounds by metal oxides in fly ash or APC par-ticles is less efficient at lower temperatures (Mahuliet al., 1997; Hirsch et al., 2000; Seames et al.,2002; Sterling and Helble, 2003). Both trivalentand pentavalent species can therefore condense(Huggins et al., 2000; Miravet et al., 2006) and arethus also found in waste pore waters (Table 4). Dur-ing cooling of flue gases, the VIV gaseous speciesthat managed to escape are probably oxidised(Diaz-Somoano et al., 2006) because Pavageauet al. (2004) found only VV in fly ash whereas Asspeciation was mixed. CrVI is easily reduced in fluegases by SO2 and speciation is thus mainly CrIII

(Huggins et al., 1999, 2000) but again, only CrVI isfound in leachates because some hexavalent Crhas been found in coal fly ash (Kingston et al.,2005). Possibly, the presence of O-rich organic mat-ter in coal leads to a higher percentage of chromatesbeing formed during combustion (Galbreath andZygarlicke, 2004). Selenium is almost completelyvolatilised from bottom ash and metallurgical resi-

dues. When volatile SeIV compounds are absorbedby metal oxides no oxidation occurs (Ghosh-Dasti-dar et al., 1996; Agnihotri et al., 1998), so it is not

surprising that SeIV is the dominant species in flyash and APC residues and their leachates. If SeVI

is found at all in fly ash leachates, it occurs in traceamounts (van der Hoek et al., 1996). Some traces ofSe0 or even Se�II may also be found due to reduc-tion of SeIV by SO2 at temperatures <150 �C (Table5, Yan et al., 2001), but these species are almostinsoluble (Seby et al., 2001) and are thus not foundin leachates.

During smelting, reducing conditions are neededto produce metals and metalloids from scrap or oreand much higher temperatures develop than duringMSWI and FFC. The absence of O2 prevents anyoxidation reactions from occurring. Arsenic, Sband Se are thus extremely volatile since oxidativeabsorption does not occur and they are found inhigh quantities in APC residues of metallurgicalprocesses where they occur in their reduced state(Dutre and Vandecasteele, 1998). In the slags them-selves, the elements are zerovalent or occur in morereduced valence states (Frandsen et al., 1994), veryoften incorporated in spinel structures if the

Table 5Compounds of toxic oxyanions experimentally found in alkaline wastes

MSWI bottom ash FFC fly ash Metallurgical residues Cement, concrete orcement-stabilised waste

AsV Ca3ðAsO4Þ2ghi Ca3AsO4 � xH2Ov

CaNaAsO4 � 7.5H2Owxy

AsIII NaAsO2a As2O3

j Ca–As–Ox

As0 FeAs, AsFeCu, CuAsSb, Cu3Asq

CrVI Cu11(OH)14(CrO4)bc Ca4[Al(OH)6]2CrO4.xH2Oz

PbCrO4b

Ca4[Al(OH)6]2CrO4 � 9H2Oc

Na2CrO4d

CrIII FeCr2O4ae FeCr2O4

kl FeCr2O4qr Ca2Cr(OH)7 � 3H2Oaa

Ca6[Cr(OH)6]2(SO4)3 � 26H2Ob,c Cr in spinelsqstu Ca2Cr2O5 � 6H2Oaa

Cr in spinelsc Ca2Cr2O5 � 8H2Oaa

Cr0 FeCr, Ni–Cr–Fer

Se0 Semn

Mo0

MoVI CaMoO4af CaMoO4

bb

Sb0 Sbq

SbV Pb2Sb2O7d

VV (Zn, Cu)PbVO4(OH)c Ca2V2O7 � 2H2Oo

V2O5p

VIII V in spinelst

WVI CaWO4l

a Zevenbergen (1994).b Freyssinet et al. (2002).c Piantone et al. (2004).d IAWG (1997).e Eusden et al. (1999).f Meima and Comans (1999).g Irgolic et al. (1999).h Huffman et al. (2000).i Galbreath and Zygarlicke (2004).j Turner (1981).

k Huggins et al. (2000).l Vassilev and Vassileva (1996).

m Yan et al. (2001).n Andren et al. (1975).o Jia et al. (2002).p Henry and Knapp (1980).q Piatak et al. (2004).r Shen et al. (2004).s Kucha et al. (1996).t Parsons et al. (2001).u Kuehn and Mudersbach (2004).v Mollah et al. (1998).w Akhter et al. (1997).x Moon et al. (2004).y Raposo et al. (2004).z Palmer (2000).

aa Kindness et al. (1994a).bb Kindness et al. (1994b).


trivalent oxidation state is stable as is the case forCrIII, SbIII and VIII (Table 5; Kim et al., 1998;Nivoix et al., 1999). Spinels are oxides of the form(M2+) (Fe3+)2O4 where M2+ and Fe3+ are the diva-lent and trivalent cations, respectively occupying

tetrahedral and octahedral interstitial positions inthe lattice formed by O2� ions. The elements arethus also leached as more reduced species comparedto other wastes (Table 4). Although Chaurand et al.(2006) found exclusively CrIII in steel slag, soluble

Table 6Predominant redox speciation of oxyanionic species in the solidphase of fresh alkaline wastes (see text)

Waste type As Cr Se Mo Sb V W

MSWI bottom ash III III IV VI III IV VIV (VI) V V

MSWI APC residues III III �II VI III V VIV (VI) 0 V

IV

FFC fly ash III III �II VI III V VIV (VI) 0 V

IV

Metallurgical residues 0 0 – 0 0 0 0III III VI III III VI

IV

Redox states between brackets occur in trace amounts (<5%).


Cr is almost always hexavalent because equilibriumwith insoluble Ca–CrIII minerals causes CrðOHÞ�4concentrations to be very low. Only oxidation ofCrIII can cause some Cr to leach. Soluble CrIII ishence only found in blast furnace slags, whosestrong reductive capacity and low Ca content causesome soluble CrðOHÞ�4 to persist (Kuehn andMudersbach, 2004).

Molybdenum and W are relatively involatile(Belevi and Moench, 2000; Paoletti, 2002) and theirhexavalent state is thermodynamically very stable inwastes and their leachates. Only in metallurgical res-idues, can they be reduced to their insoluble elemen-tal state.

2.3. Oxidation and reduction processes in alkaline

wastes

Table 1 shows that all alkaline wastes containreadily oxidisable compounds such as non-com-busted organic matter, ferrous and metallic Fe andAl or sulphides that buffer the redox potential ofcolumn test leachates to low values (Fallman andHartlen, 1994). Whereas bottom ash and especiallymetallurgical residues contain an appreciableamount of reducing agents, the overall reductivecapacity of APC residues and fly ash is compara-tively much lower. Pyrite (FeS2) has been detectedin FFC fly ash (Mattigod et al., 1990; Vassilev andVassileva, 1996) but most S is considered to occuras sulphates (Fruchter et al., 1990; IAWG, 1997).Cement, in contrast, does not contain a significantamount of readily oxidisable compounds and there-fore exhibits a positive redox potential during leach-

ing (Glasser, 1997). These compounds may, ofcourse, be furnished by blending agents such asblast furnace slags.

Fig. 1 shows Eh-pH diagrams of oxyanionic spe-cies, calculated with recent thermodynamic data(Wanty and Goldhaber, 1992; Beverskog and Puig-domenech, 1997; Cruywagen, 2000; Seby et al.,2001; Cruywagen et al., 2002; Filella and May,2003; Nordstrom and Archer, 2003). It appears thatfrom a thermodynamic point of view, AsV, CrVI,SeVI, SbV species should be reduced by H2, Fe(OH)2

and HS� in alkaline situations and O2, Fe(OH)3 andMnO2 may oxidize AsIII, SeIV and SbIII compounds.Table 7 shows the most important oxidizing andreducing agents that may alter the redox state ofoxyanions in natural systems in a realistic timescale(detectable in no more than 30 days). Although oxi-dation reactions, for instance, oxidation of CrIII bydissolved O2 (Petersen, 1998) and oxidation of SbIII

in the presence of Fe oxides (Leuz et al., 2006b),generally are faster at high pH, oxidation by dis-solved O2 at room temperature is either very slowor does not occur at all (van der Weijden and Reith,1982; Eary, 1987; Bissen and Frimmel, 2003; Leuzand Johnson, 2005). Adsorption to metal oxideshas, however, been shown to catalyse oxidation byO2, whereas MnO2 appears to be a powerful oxidantfor oxyanions.

The knowledge on redox reactions in alkalinesolid wastes is incomplete, but, except for As,Cr and Sb, redox transformations appear to belimited which implies a clear link between specia-tion in waste leachates and in the solid phase. Inmetallurgical residues, for example, reduced spe-cies may prevail longer than in other waste leach-ates, especially when an appreciable amount ofsulphide is present such as in blast furnace slagsor other ore-based slags. Steel slag, on the otherhand, is produced from scrap and its reducingproperties are determined not by sulphides butby ubiquitous zerovalent metal particles, whichare easily passivated by an oxide layer. Bonhoureet al. (2006) found that in blended cement sys-tems, despite redox potentials as low as�300 mV, no significant reduction of selenateoccurred. As can be seen from Fig. 1, it is unli-kely that redox transformations can play a signif-icant role in the geochemical behaviour of V.Furthermore, it is unlikely that MoIV compoundscan be formed by simple reduction of molybdateat room temperature and WO2�

4 oxyanions areeven more stable (Holm, 1987).

-0.8

-0.6

-0.4

-0.2

0

0.2

0.4

0.6

0.8

1

1.2

0 2 4 6 8 10 12 14pH

Eh

(V)

H3 A

sO4 ˚

HAsO42-

AsO43-

H2AsO4-

H2AsO3-

H3AsO3°

As(s)

MnO2

Fe2+

Fe(OH)2

HS-

H2S

Fe3+

Fe(OH)3

-0.8

-0.6

-0.4

-0.2

0

0.2

0.4

0.6

0.8

1

1.2

0 2 4 6 8 10 12 14pH

Eh

(V)

Eh

(V)

SeO42-

SeO32-

Se0

HSeO4-

H2SeO3°

HSeO3-

H2SeHSe-

Fe3

Fe2+

Fe(OH)2

Fe(OH)3

MnO2

HS-

H2S

-0.8

-0.6

-0.4

-0.2

0

0.2

0.4

0.6

0.8

1

1.2

0 2 4 6 8 10 12 14pH

Eh

(V)

CrO42-

Cr3+

CrO

H2+

HCrO42-

Cr(O

H)2 +

Cr(OH)4-

Fe2+

Fe(OH)2

MnO2

H2S

HS-

Fe3+

Fe(OH)3

-0.8

-0.6

-0.4

-0.2

0

0.2

0.4

0.6

0.8

1

1.2

0 2 4 6 8 10 12 14pH

Eh

(V)

Sb(OH)3°

Sb(O

H)5 ˚

Sb(O

H)4 -

Sb(OH)6-

Sb(O

H)2 +

Fe(OH)2

Fe2+

H2S

MnO2

HS-

Fe3+

Fe(OH)3

-0.8

-0.6

-0.4

-0.2

0

0.2

0.4

0.6

0.8

1

1.2

0 2 4 6 8 10 12 14pH

VO2+

H2VO4-

HVO42-

VO43-

VO2+

VO(OH)+

V3+

V(OH)2+

VOH2+

MnO2Fe2+

Fe3+

Fe(OH)3

Fe(OH)2H2S

HS-

°

Fig. 1. Eh-pH predominance diagrams for the soluble As, Cr, Se, Sb, V species at 25 �C. Region of occurrence of reduced Fe and Scompounds, and MnO2 are indicated (based on Brookins, 1988). Equilibrium with As0, Se0, solid Fe, Mn and S compounds was calculatedwith a soluble concentration of 10�6, 10�6, 10�5, 10�7 and 10�2 mol/L, respectively.


Table 7Possible oxidising and reducing agents of oxyanions in natural systems

Element Oxidizingagents

Reducing agents References

As MnO2 HS� Cherry et al. (1979), Eary et al. (1990), and Bissen and Frimmel (2003)O2(ads�Fe(OH)3)

Cr MnO2 Organic matter; Hattori et al. (1978), Johnson and Xyla (1991), Kindness et al. (1994a), Glasser(1997), Petersen (1998), Rodrigruez-Pinero et al. (1998), Abbas et al. (2001), Caiet al. (2003), Chen et al. (2003), Pillay et al. (2003), Halim et al. (2004), Li et al.(2004), and Cao and Zhang (2006)

O2(ads�CaO) Fe0, H2, Al0, FeII –compounds

O2

Se MnO2 Fe0, FeII –compounds

Scott and Morgan (1996), Refait et al. (2000), Scheidegger et al. (2003), and Zhanget al. (2005)

Sb MnO2, HS� Belzile et al. (2001), Filella et al. (2002), and Leuz et al. (2006a,b)O2ðads–FeðOHÞ3Þ

V O2ðads–Al2O3Þ Organic matter,HS�

Wehrli and Stumm (1988, 1989), Eary et al. (1990), and Wanty and Goldhaber(1992)

O2

The agents indicated in bold have been confirmed to oxidise or reduce oxyanions in alkaline solid wastes.


3. Solubility

3.1. Metalate precipitation

Although they are sometimes found in wastes(Table 5), solid oxides of AsV, AsIII, CrVI, SeIV, SeVI,SbV, MoVI,VV and WVI are generally not relevant inthe context of leaching due to their relatively high sol-ubility. Since Ca2+ will often be the most importantmultivalent cation in solution, particular attentionshould be given to Ca metalates (Table 8) but possi-bly also to Ba and Pb metalates because they have rel-atively low solubility products (Table 9). The pH-dependent leaching behaviour of these metalates inalkaline wastes of diverse origin is shown in Fig. 2together with saturation curves of Ca, Ba and Pbmetalates. Although wastes with different character-istics are considered, the oxyanion forming elementsstill exhibit similarities in their general leaching trend.For example, a leaching minimum is found round pH12 and sometimes also a second one at pH < 6. It canbe seen that most Ca and Ba metalates exhibit a min-imum in leaching at alkaline pH whereas Pb meta-lates are more soluble in alkaline situations.

In Pb slag, Pb arsenate appears to control arse-nate leaching but in most other cases, where Ca ismore abundant, hydrated Ca arsenates are the mostlikely to control arsenate leachability. Ca3(AsO4)2 �xH2O and CaNaAsO4 � 7.5H2O have been detectedexperimentally in cement (Table 5). Ca4(OH)2-

(AsO4)2 � 4H2O and arsenate apatite (Ca5(A-sO4)3(OH)) are less soluble but only occur whenthe available Ca amount is high (Bothe and Brown,1999a,b, 2002; Moon et al., 2004; Zhu et al., 2005).Furthermore, arsenate apatite has a very small crys-tal morphology which is distorted in the presence ofsmall amounts of Mg2+ ions in which case Ca3

(AsO4)2 � xH2O again occurs (Bothe and Brown,2002). Calcium arsenites are more soluble than Caarsenates. Solids that are several orders of magni-tude more soluble than the given pH-dependentleaching curves were not depicted.

The presence of Ca–CrIII compounds in alkalinesystems may explain the low solubility of CrIII,despite its amphoteric behaviour (Glasser, 1997).Solubility products for most of these compoundshave, however, not yet been determined. Jing et al.(2006) estimated a solubility product for Ca2Cr2O5.6H2O through fitting thermodynamic data to theCrIII solubility in a stabilised/solidified soil andPerkins (2000) calculated the solubility product ofCa2Cr2O5 � 8H2O from solubilities in a CaO–Cr2(SO4)3 mixture. Fig. 2 demonstrates the largedifference in solubility between CrIII and CrVI com-pounds. Much more than CaCrO4precipitation,BaCrO4 precipitation or Ba(S, Cr)O4 solid solutionformation are suggested to control chromate leach-ing, in for instance FFC fly ash (Eary et al., 1990;Fruchter et al., 1990), APC residues (Astrup et al.,2006) and steel slag (Fallman, 2000).

Table 8Solubility products of calcium metalates at 25 �C

Calciummetalate logKsp Reference

AsV Ca3 (AsO4)2 � 3H2O ¼ 3Ca2þ þ 2AsO3�4 þ 3H2O �21.14 Zhu et al. (2005)

Ca3 (AsO4)2 � 2.25H2O ¼ 3Ca2þ þ 2AsO3�4 þ 2:25H2O �21.40 Zhu et al. (2005)

Ca3 (AsO4)2 � 3.66H2O ¼ 3Ca2þþ2AsO3�4 þ 3:66H2O �21.00 Bothe and Brown (1999b)

Ca3 (AsO4)2 � 4.25H2O ¼ 3Ca2þ þ 2AsO3�4 þ 4:25H2O �21.00 Bothe and Brown (1999b)

Ca3 (AsO4)2 � 10H2O ¼ 3Ca2þ þ 2AsO3�4 þ 10H2O �21.21 Raposo et al. (2004)

CaNaAsO4 � 7.5H2O ¼ Ca2þ þNaþ þAsO3�4 þ 7:5H2O �9.08 Raposo et al. (2004)

Ca5 (AsO4)3 (OH) ¼ 5Ca2þ þ 3AsO3�4 þOH� �40.12 Zhu et al. (2005)

�38.04 Bothe and Brown (1999b)Ca4 (OH)2 (AsO4)2 � 4H2O ¼ 4Ca2þ þ 2AsO3�

4 þ 2OH �27.49 Zhu et al. (2006)�29.20 Bothe and Brown (1999b)

AsIII CaHAsO3 ¼ Ca2þ þHAsO2�3 �6.52 Stronach et al. (1997)

�6.98 Dutre and Vandecasteele (1998)

CrVI CaCrO4 ¼ Ca2þ þ CrO2�4 �2.27 Allison et al. (1991)

CrIII Ca2Cr2O5.6H2O ¼ 2Ca2þ þ 2CrðOHÞþ2 þ 6OH� þH2O �46.5 Jing et al. (2006)Ca2Cr2O5.8H2O ¼ 2Ca2þ þ 2CrðOHÞþ2 þ 6OH� þ 3H2O �16.85 Perkins (2000)

SeVI CaSeO4 ¼ Ca2þ þ SeO2�4 �4.77 Essington (1988)

CaSeO4 � 2H2O ¼ Ca2þ þ SeO2�4 þ 2H2O �2.947 Parkhurst and Appelo (2005)

�3.02 Wagman et al. (1982)

SeIV CaSeO3 � H2O ¼ Ca2þ þ SeO2�3 þH2O �7.76 Sharmasarkar et al. (1996)

�6.84 Baur and Johnson (2003a)CaSeO3 ¼ Ca2þ þ SeO2�

3 �7.65 Essington (1988)CaSeO3 � 2H2O ¼ Ca2þ þ SeO2�

3 þ 2H2O �5.44 Elrashidi et al. (1987)

MoVI CaMoO4 ¼ Ca2þ þMoO2�4 �7.93 Felmy et al. (1992)

�7.94 Rai and Zachara (1984)

SbV Ca(Sb(OH)6)2 ¼ Ca2þ þ 2SbðOHÞ�6 �12.55 Johnson et al. (2005)Ca(Sb(OH)6)2 � 6H2O ¼ Ca2þ þ 2SbðOHÞ�6 �10.57 Amme (1999)V(V)

VV Ca3(VO4)2 ¼ 3Ca2þ þ 2VO3�4 �17.97 Allison et al. (1991)

Ca3(VO4)2 � 4H2O ¼ 3Ca2þ þ 2VO3�4 þ 4H2O �17.57 Allison et al. (1991)

Ca2V2O:7H2O ¼ 2Ca2þ þ 2HVO2�

4 �12.8 Allison et al. (1991)Ca2V2O7 � 2H2O ¼ 2Ca2þ þ 2HVO2�

4 þH2O �10.25 Allison et al. (1991)Ca(VO3)2 � 4H2O ¼ Ca2þ þ 2H2VO2�

4 þ 2H2O �17.97 Allison et al. (1991)WVI CaWO4 ¼ Ca2þ þWO2�

4 � 8.72 Wood (1992)


CaSeO4 � 2H2O was assumed to be the most sta-ble Ca selenate because sulphate and selenate havea similar chemistry and gypsum (CaSO4 � 2H2O) isthe most stable CaSO4 at ambient temperatures(Freyer and Voigt, 2003). Selenite and selenate com-pounds have a high solubility compared to othermetalates.

Fig. 2 suggests that Ca vanadates are moreimportant in the context of solubility control thanPb or Ba vanadates although at pH < 11, Pb vana-dates appear to be relevant phases for leachingcontrol in MSWI air pollution control residueseven though Ca is much more abundant than Pbin these wastes. Calcium vanadates are, however,not well studied. Fig. 3 shows the CaO–V2O5–H2O phase diagram that was based on model cal-

culations using PHREEQC, thermodynamic datafor vanadate hydrolysis from Cruywagen (2000)and solubility products from Table 8. In alkalinesystems, only Ca3 (VO4)2 and Ca2V2O7 are found(Schindler et al., 2000) but it appears thatCa2V2O7 is only formed when no residual portlan-dite is present or the pH is relatively low (Fig. 2).Similar phase diagrams exist for the CaO–As2O5–H2O system (Dumm and Brown, 1997), theCaO–As2O3–H2O system (Stronach et al., 1997)and the CaO–SeO2–H2O system (Bothe andBrown, 2002). Much less is known about Ca min-erals containing V of lower valence states. Evansand Garrels (1958) detected the mineral simplotite(CaVIV

4 O9:5H2O) in reduced zones of alkaline(pH > 8) ore deposits.

Table 9Lead and Ba metalate solubility products at 25 �C

Metalate logKsp Reference

Pb5(AsO4)3Cl ¼ 5Pb2þ þ 3AsO3�4 þ Cl� �83.51 Magalhaes and Silva (2003)

Pb3(AsO4)2 ¼ 3Pb2þ þ 2AsO3�4 �35.5 Allison et al. (1991)

Ba3(AsO4)3 ¼ Ba2þ þHAsO3�4 �23.53 Zhu et al. (2005)

PbCrO4 ¼ Pb2þ þ CrO2�4 �12.6 Allison et al. (1991)

BaCrO4 ¼ Ba2þ þ CrO2�4 �9.78 Rai et al. (1988)

PbSeO4 ¼ Pb2þ þ SeO2�4 �6.84 Seby et al. (2001)

BaSeO4 ¼ Ba2þ þ SeO2�4 �7.3 Seby et al. (2001)

PbSeO3 ¼ Pb2þ þ SeO2�3 �12.12 Seby et al. (2001)

BaSeO3 ¼ Ba2þ þ SeO2�3 �6.57 Seby et al. (2001)

PbMoO4 ¼ Pb2þ þMoO2�4 �15.80 Rai and Zachara (1984)

BaMoO4 ¼ Ba2þ þMoO2�4 �6.96 Bard et al. (1985)

PbV2O7 ¼ Pb2þ þ 2HVO2�4 �32.2 Allison et al. (1991)

BaV2O7.2H2O ¼ Ba2þ þ 2HVO2�4 þ 2H2O �15.93 Allison et al. (1991)

Pb3(VO4)2 ¼ 3Pb2þ þ 2VO3�4 �51.14 Allison et al. (1991)

Ba3(VO4)2 � 4H2O ¼ 3Ba2þ þ 2VO3�4 þ 4H2O �25.84 Allison et al. (1991)

As-BA

-9

-7

-5

-3

2 6 10 14

Cr-BA-9

-7

-5

-3

2 6 10 14

Mo-BA

-9

-6

-3

0

2 6 10 14

BaM oO4

CaM oO4

PbM oO4

Sb-BA

-9

-7

-5

-3

2 6 10 14

Se-BA

-9

-6

-3

0

2 6 10 14

As-FA

-9

-7

-5

-3

2 6 10 14

Cr-APC

-9

-7

-5

-3

2 6 10 14

Mo-FA

-9

-7

-5

-3

2 6 10 14

Sb-FA

-9

-7

-5

-3

2 6 10 14

Ca[Sb(OH)6]2

Se-FA

-9

-6

-3

0

2 6 10 14

As-PbS

-9

-7

-5

-3

2 6 10 14

Cr-COPR

-7

-5

-3

-1

1

2 6 10 14

Cr(III)

Cr(VI)

Mo-PbS

-9

-6

-3

0

2 6 10 14

V-APC

-9

-6

-3

0

2 6 10 14

Se-PbS-9

-6

-3

0

2 6 10 14

As-leaching

←L

og(L

each

ing)

(m

ol/L

)→

Pb3(AsO4)2

Ba3(AsO4)2

Ca3(AsO4)2.xH2O

Ca4(AsO4)2

(OH)2.4H2OCa5(AsO4)3(OH)

Cr-leaching

Ca2Cr2O5.6H2O

BaCrO4

PbCrO4

CaCrO4

Cr(OH)3

Pb5(AsO4)3Cl

Cr-Cem

-9

-7

-5

-3

2 6 10 14

Mo-Cem

-8

-6

-4

-2

2 6 10 14

V-Cem

-9

-6

-3

0

2 6 10 14

V-leaching

Pb3(VO4)2

BaV2O7

Ba3(VO4)2

Ca3(VO4)2

CaV2O7

Se-leachingCaSeO3

PbSeO3

BaSeO3

PbV2O7

CaSeO4

←pH→

Fig. 2. pH-dependent leaching behaviour of oxyanion forming elements in MSWI bottom ash (BA) (Cornelis et al., 2006a), coal fly ash(FA) (De Groot et al., 1989), MSWI air pollution control residues (APC) (Astrup et al., 2006), Pb slag (PbS) (Saikia et al., unpublishedresults), Chromite ore processing residue (COPR) (Geelhoed et al., 2002) and an OPC-mortar (van der Sloot et al., 2001). Curves indicatehypothetical saturation of Ca–, Ba and Pb metalates calculated with PHREEQC (Parkhurst and Appelo, 2005) with the MINTEQA2database (Allison et al., 1991) updated with solubility data from Tables 8 and 9 using the Davies equation for activity corrections.


The minerals powellite (CaMoO4) and schee-lite (CaWO4) have been detected in alkalinewastes (Table 5) and it can be assumed thatno other Ca metalates precipitate in alkaline sys-tems. The solubility product of Ca(Sb(OH)6)2

was obtained at near-neutral pH values (Johnsonet al., 2005), and the mineral has not yet beendetected in alkaline wastes. Calcium antimonateprecipitation at high pH is a subject of ongoingresearch.

0

0.2

0.4

0.6

0.8

1

0 0.2 0.4 0.6 0.8 1(Molar ratio CaO)1/5

(Mol

ar ra

tio V

2O5)

1/5

Ca(OH)2

V2O5

A: Ca3(VO4)2

C: Ca(VO3)2

B: Ca2V2O7

A

C

B

H2O

Fig. 3. Modelled occurrence of Ca vanadates in the CaO–V2O5–H2O system at 25 �C.


Fig. 2 suggests that solubility control is possiblefor CrIII, AsV, MoVI, VV and WVI, which has beenproposed also in a limited amount of model studies(Table 10). Molybdate concentrations in leachatesare often found in equilibrium with powellite. Itcan therefore be assumed that tungstate regularlycan be found in equilibrium with scheelite due to asimilar chemistry. Most model studies however sug-gest that in alkaline wastes oxyanion concentrationsare below saturation levels with respect to metalatesolubility, especially in the case of chromate, sele-nite, selenate and antimonate (e.g. Fruchter et al.,1990; van der Hoek et al., 1994; Duchesne andReardon, 1999; Johnson et al., 1999; Palmer, 2000;Ochs et al., 2002; Astrup et al., 2006; Corneliset al., 2006a,b).

3.2. Adsorption and solid solution formation

Surface adsorption and solid solution formationwith major minerals can reduce the leachate concen-

Table 10Metalates suggested to exert solubility control on oxyanions in model

Element Mineral Waste type Referenc

AsIII As2O3 FFC fly ash Turner (AsV Ca3(AsO4)2 � xH2O S/S stabilised waste VandecaCrIII CaCr2O5 � 6H52O Cement paste Jing et a

CaCr2O5 � 8H52O Cement paste PerkinsMoVI CaMoO4 MSWI bottom ash Kersten

CaMoO4 FFC fly ash Eary etCaMoO4 S/S stabilised waste Baur et

VV Pb3 (VO4)2, Pb2V2O7 MSWI APC residues Astrup ePb3 (VO4)2, Pb2V2O7 MSWI fly ash Astrup eCa3(VO4)2 Steel slag Huijgen

WVI CaWO4 MSWI bottom ash KerstenCaWO4 S/S stabilised waste Baur et

tration of oxyanions below pure-phase saturationlevels in alkaline wastes. The interaction can beeither specific or non-specific in the case of surfaceadsorption. Solid solution formation is always spe-cific because incorporation of ions depends onparameters such as size, charge and geometry. Sur-faces in alkaline wastes are, however, predomi-nantly negatively charged (Cocke and Mollah,1993; Kirby and Rimstidt, 1993). Probably, fewminerals develop a permanent positive charge andpH-dependent charges are predominantly negativeat high pH. Furthermore, trace oxyanions have tocompete with ubiquitous anions such as SO2�

4 ,CO2�

3 , silicate and Cl�. Significant immobilisationof trace oxyanions therefore occurs through specificinteractions only.

3.2.1. Iron oxides

IronIII is preferentially precipitated as Fe oxides,due to their thermodynamic stability. Therefore,FeIII–oxyanion compounds are seldom detected inalkaline wastes. Much more important is the surfaceadsorption complexation with Fe oxides, which formost oxyanions is well established in the literature.Amorphous Fe oxides, also called hydrous ferricoxides (HFO), and amorphous Al oxides are ubiqui-tous in many alkaline wastes (Warren and Dudas,1985; van der Hoek et al., 1996; Meima andComans, 1997; Freyssinet et al., 2002; Piantoneet al., 2004). They usually are more important inthe context of adsorption than crystalline oxidesdue to their much higher specific surface area.McKenzie (1983), for instance, found a linear rela-tionship between the extent of MoO2�

4 adsorptionand specific surface area of different amorphousand crystalline Fe oxides. During weathering,freshly precipitated HFO are, however, progres-

studies

e

1981)steele et al. (2002), Halim et al. (2005), and Phenrat et al. (2005)l. (2006)(2000)et al. (1997), IAWG (1997), and Meima and Comans (1999)al. (1990)al. (2001)t al. (2006)t al. (2006)and Comans (2006)et al. (1997)al. (2001)


sively transformed into crystalline oxides (Fordet al., 1997; Meima, 1997), which reduces theiradsorptive capability.

Most oxyanions can form inner-sphere com-plexes with Fe oxide surfaces (Goldberg et al.,1996; Su and Suarez, 2000; Weerasooriya andTobschall, 2000; Dixit and Hering, 2003; Peacockand Sherman, 2004). Conversely, SeO2�

4 sorbsnon-specifically through an outer-sphere complexat pH > 6 (Peak and Sparks, 2002). Adsorption ofoxyanions on HFO at alkaline pH modelled accord-ing to the diffuse layer approach (Dzombak andMorel, 1990) is shown in Fig. 4. The strength ofadsorption varies and hence also the pH at whichoxyanions are desorbed from HFO surfaces. Thethermodynamic data for adsorption of SbV inFig. 4 is based on linear free energy relations(Dzombak and Morel, 1990) because no experimen-tally based data exists for adsorption of Sb specieson HFO. Leuz et al. (2006b) compared adsorptionof Sb species on goethite (FeOOH) with As speciesand found that although all As and Sb oxyanionsform inner-sphere complexes on goethite, antimo-nate is desorbed at much lower pH values than arse-nate. The adsorption of SbIII, on the other hand,remains high, independent of pH, and exceeds thatof AsIII at pH > 10. Hence, in alkaline wastes withpH > 12, adsorption is probably only significant inthe case of AsIII, CrIII and SbIII and oxyanion leach-ability is reduced below pure-phase saturation byadsorption or solid solution formation with otherabundant minerals in alkaline wastes as discussedbelow. During weathering, however, the leachatepH can be lowered to as low as 8. At that pH,adsorption by Fe oxides is also relevant for SbV

0%

25%

50%

75%

100%

7 8 9 10 11 12 13 14pH

% a

dsor

bed

CrVI VVSeIVMoVI CrIIIWVI

AsV

AsIIISbV

SeVI

Fig. 4. Modelling of adsorption of oxyanions on hydrous ferricoxides using the diffuse layer model (Dzombak and Morel, 1990)updated with intrinsic constants from Gustafsson (2003) forMoVI and WVI and from Dixit and Hering (2003) for AsV andAsIII. Oxyanion concentration was 10�6 mol/L, HFO concentra-tion was 1 mmol/L and ionic strength 0.01 mol/L.

and MoVI in bottom ash (Meima and Comans,1998a,b; Cornelis et al., 2006a,b) and probably alsofor AsV, CrVI, SeIV, VV and WVI oxyanions. Signif-icant desorption can, however, be induced by thepresence of competing anions that also forminner-sphere complexes. Chloride, NO�3 , SO2�

4

adsorb non-specifically (e.g. Wilkie and Hering,1996), whereas silicate and carbonate form stronginner-sphere surface complexes (Carlson and Schw-ertman, 1981; Tejedor-Tejedor and Anderson, 1990;Vempati et al., 1990; Su and Suarez, 1997; Wijnjaand Schultess, 2001; Hiemstra et al., 2004).

3.2.2. Aluminium oxides

In spite of their abundance, adsorption to amor-phous Al oxides is much less considered than toHFO. Compared to HFO, thermodynamic data isgenerally lacking but it has been found that at leastsome oxyanions adsorb less strongly on Al oxides.Arsenate, arsenite and selenate show a much morespecific interaction with Fe oxides than with Al oxi-des (Wijnja and Schultess, 2000; Goldberg andJohnston, 2001). Amorphous Al oxides are, how-ever, much less prone to mineral transformationsand may therefore control oxyanion leaching inweathered residues (Meima, 1997).

3.2.3. EttringiteSolid solution formation with ettringite in cement

and some types of alkaline wastes is likely for manyoxyanions because it is a common mineral (Glasser,1997; Piantone et al., 2004) and it has been shownthat partial or full replacement of SO4 (Fig. 5) ispossible in the case of AsO3�

4 , CrO2�4 , SeO2�

4 ,SeO2�

3 , MoO2�4 , SbðOHÞ�6 andVO3�

4 (Kumarathasan

AsO43-

SO42-

Columns of {Ca6[Al(OH)6]4.24H2O}6+

Fig. 5. Arsenic incorporation in ettringite (after Myneni et al.,1997).


et al., 1990; Pollman et al., 1993; Myneni et al.,1997; Perkins, 2000; Perkins and Palmer, 2000; Ochset al., 2002; Baur and Johnson, 2003a,b; Zhang andReardon, 2003; Cornelis et al., 2006a,b). Cr3+ canalso replace Al3+ (Perkins, 2000). Particle surfacesof ettringite exhibit a net negative charge (Myneniet al., 1997) so incorporation of anions in the bulkis probably much more important than surfaceadsorption (Myneni et al., 1997; Ochs et al., 2002;Cornelis et al., 2006a).

The extent of solid solution with ettringite isthought to be inversely proportional to the differ-ence in size and electronegativity of the oxyanioncompared to SO2�

4 (Zhang and Reardon, 2003). Aconsistent observation is that uptake of MoO2�

4 , islow or non-existent (Kumarathasan et al., 1990;Kindness et al., 1994b; Zhang and Reardon, 2003;Saikia et al., 2006) which can be attributed to itslarge size (Table 11). However, solid solution isprobably determined by more factors than just sizeand electronegativity. Ettringite shows an equalpreference for CrO2�

4 andSeO2�4 (Klemm and Bhatty,

2002), which is reflected in a similar solubility of thecorresponding ettringite analogues (Table 12), andthe ettringite structure is destroyed in the presence

Table 11Ionic radii of different redox species of oxyanion formingelements (Greenwood and Earnshaw, 1984)

Element Elektronegativity (redox state) Ionic radius

S 2.5 (VI) 0.29 AAs 2.0 (V) 0.47 A (III) 0.69 ACr 1.6 (VI) 0.52 A (III) 0.69 ASe 2.4 (VI) 0.42 A (IV) 0.50 AMo 1.8 (VI) 0.62 ASb 1.9 (V) 0.62 A (III) 0.76 AV 1.6 (V) 0.59 A (IV) 0.61 A

Table 12Solubility of Aft and AFm analogues at 25 �C

Aft-phase

Ca6[Al(OH)6]2(SO4)3 � 32H2O + 12H+ ¼ 6Ca2þ þ 2Al3þ þ 3SOCa6[Al(OH)6]2(CrO4)3 � 26H2O + 12H+ ¼ 6Ca2þ þ 2Al3þ þ 3CrCa6[Cr(OH)6]2(SO4)3 � 26H2O + 12H+ ¼ 6Ca2þ þ 2Cr3þ þ 3SOCa6[Al(OH)6]2(SeO4)3 � 37.5H2O + 12H+ ¼ 6Ca2þ þ 2Al3þ þ 3SeO

Afm phase

Ca4[Al(OH)6]2SO4 � 13H2O + 12H+ ¼ 4Ca2þ þ 2Al3þ þ SO24

Ca4[Al(OH)6]2CrO4 � 9H2O + 12H+ ¼ 4Ca2þ þ 2Al3þ þ CrOCa4[Al(OH)6]2MoO4 � 10H2O + 12H+ ¼ 4Ca2þ þ 2Al3þ þMoOCa4[Al(OH)6]2SeO4 � xH2O + 12H+ ¼ 4Ca2þ þ 2Al3þ þ SeO

a Value calculated from solubility data of Kindness et al. (1994b) usi

of high AsO3�4 concentrations (Myneni et al.,

1997; Saikia et al., 2006) despite the less beneficialsize and electronegativity of CrO2�

4 compared toSeO2�

4 and AsO3�4 . Selenite incorporation in ettring-

ite is unlikely (Baur and Johnson, 2003b) whereasuptake of antimonate and vanadate by ettringitecan be significant (Kumarathasan et al., 1990; Corn-elis et al., 2006a,b) despite their large size.

Solid solution formation in ettringite is fre-quently believed to be a controlling mechanism foroxyanion leaching, for example for CrVI in cement(Ochs et al., 2002; Rose et al., 2003) or concrete(Palmer, 2000), SeVI in cement (Ochs et al., 2002)or SbV in MSWI bottom ash (Meima and Comans,1998b; Johnson et al., 1999; Cornelis et al.,2006a,b). It has, however, only been modelled inone case (Ochs et al., 2002) because thermodynamicdata is lacking or incomplete. Only in the case of thechromate–ettringite solid solution, solubility ofsolid solution end-members as well as non-idealityparameters are available (Perkins, 2000).

3.2.4. Monosulphate and hydrocalumite

When SO2�4 becomes limited, for example during

the hardening of cement, ettringite is converted tomonosulphate or to its hydroxide analogue hydro-calumite (Ca4[Al(OH)6]2(OH)2 � 6H2O) (Gougaret al., 1996) which are both more stable at highpH (Chrysochoou and Dermatas, 2006). Althougheven more scarcely studied, solid solution formationwith these minerals is suspected to cause a largerreduction in oxyanion mobility as compared toettringite but the order of preference is different(Perkins and Palmer, 2001; Zhang and Reardon,2003; Chrysochoou and Dermatas, 2006). The solu-bility product of chromate and molybdate ana-logues, for instance, is lower than the ones of

logKsp Reference

2�4 þ 44H2O 57.45 Damidot and Glasser (1993)O2�

4 þ 38H2O 60.54 Perkins and Palmer (2000)2�4 þ 38H2O 114.2 Perkins (2000)

2�4 þ 49:5H2O 61.29 Baur and Johnson (2003b)

� þ 25H2O 72.57 Damidot and Glasser (1993)2�4 þ 21H2O 71.62 Perkins and Palmer (2001)

2�4 þ 22H2O 71.66a Kindness et al. (1994b)

2�4 þ ðxþ 12ÞH2O 73.40 Baur and Johnson (2003b)

ng Phreeqc.


selenate and sulphate (Table 12). Solid solution for-mation in monosulphate has therefore been sug-gested as a possible controlling mineral forchromate and molybdate in cement (Kindnesset al., 1994a,b; Rose et al., 2003). Selenite, as withettringite, interacts through a surface adsorptionmechanism (Baur and Johnson, 2003a,b).

3.2.5. Hydrotalcite-like minerals

The mineral hydrotalcite (Mg6[Al(OH)6]2-

CO3 � 4H2O) is structurally very similar to hydrocal-umite but has a permanent positive charge. It showsaffinity for oxyanions such as arsenate (Dousovaet al., 2003), chromate (Misra and Perrotta, 1992;Chatelet et al., 1996), molybdate (Misra and Perrot-ta, 1992), vanadate (Ulibarri et al., 1994) and possi-bly also selenite and selenate (You et al., 2001).ChromiumIII-hydrotalcite has been detected incement but it is stable at lower pH values thanettringite (Rose et al., 2003). Few studies considerit as a possible absorbent for oxyanions. It has beenpredicted that hydrotalcite is quantitatively lessimportant than ettringite and monosulphate incement (Lothenbach and Winnifeld, 2006) and highconcentrations of soluble carbonate prevent anionexchange (Dousova et al., 2003).

3.2.6. Portlanditevan der Hoek et al. (1994) found that arsenate

showed some affinity for portlandite whichdeclined as pH was increased further above pH12.5. It was thought that this mechanism alsoreduced selenite mobility in a coal fly ash, althoughto a lesser extent than arsenate. Molybdate andantimonate possibly also adsorb to portlandite(Cornelis et al., 2006a,b).

3.2.7. C–S–H

Silica is present in most alkaline wastes andcauses the precipitation of Ca silicate hydrate,which comprises approximately 50–60 mol% ofmost cement pastes (Glasser, 1997) but has alsobeen found in bottom ash (Speiser et al., 2000).Since it is amorphous in nature, no real crystal sub-stitution reactions can occur. However, its disor-dered stacking of layers creates a large volume ofmicropores and a vast specific surface area, avail-able for sorption. At Ca/Si ratios higher than 1.2,the CSH surfaces are positively charged (Jonssonet al., 2004). CSH has thus been shown to exhibitan adsorption potential for arsenate (Phenratet al., 2005), arsenite (Stronach et al., 1997), selenite

(Baur and Johnson, 2003a) and to a limited extentalso chromate (Omotoso et al., 1998). In waste sys-tems there are, however, always vast amounts ofcompeting anions (OH�, SO2�

4 , CO2�3 ). In addition,

adsorption is not the only mechanism by whichCSH immobilizes metals. There are indications thatboth CrO2�

4 andAsO3�4 can substitute silicate in

CSH structures (Fowler et al., 1995; Halim et al.,2004).

3.2.8. Calcite

In spite of its abundance and capacity to scav-enge arsenite and selenite up to pH 12 (Goldbergand Glaubig, 1988a,b; Roman-Ross et al., 2002),calcite is seldom considered as a possible sink foroxyanions in wastes. The potential determining ionsfor the calcite surface are Ca2+ and CO2�

3 (or HCO�3or H2CO3) and not H+ and �OH as is the case withmetal oxides (Foxall et al., 1979). The point of zerocharge of calcite can thus be found at pCa = 4.4with the surface exhibiting positive charge at Caconcentrations above this value (Stipp, 1999). Oxya-nions with coordination number 3, such asAsO3�

3 and SeO3�3 , show a particular affinity for cal-

cite surfaces because they have non-bonding valenceshell electron pairs and assume a pyramidal trigonalshape like CO2�

3 (Cheng et al., 1997, 1999). Molyb-date and selenate are not trigonal and are henceonly marginally retained by calcite (Reardon et al.,1993; Goldberg et al., 1996). High concentrationsof SO4 will depress incorporation of trace oxyanionsdespite its tetragonal coordination (Cowan et al.,1990).

3.2.9. Gypsum

Most oxyanions do not exhibit a particular affin-ity for gypsum although it appears that HAsO2�

4 ,CrO2�

4 andSeO2�4 can form a solid solution with this

mineral (Freyer and Voigt, 2003; Roman-Rosset al., 2002). Fernandez-Gonzalez et al. (2004) thor-oughly studied the (CaSeO4–CaSO4) � 2H2O solidsolution and found that, although of all oxyanions,the structure, charge and geometry of the selenateoxyanion most resembles that of SO4, a miscibilitygap occurred between molar S/Se ratios of 0.23and 0.77. This suggests that solid solution formationwith gypsum will be limited for other oxyanions.Molybdate as well as antimonate showed no affinityfor gypsum, which explains their increased leach-ability when gypsum is formed at the expense ofAFt and AFm phases during weathering (Corneliset al., 2006a,b).


4. Conclusions

The main conclusions arising from the precedingdiscussion are:

� As, Cr, Se, Mo, Sb, V and W are redox-sensitiveelements but only AsV, AsIII, CrVI, SeVI, SeIV,MoVI, SbV, VV and WVI are relevant for alkalinewastes in the context of leaching. The leachingbehaviour of As, Cr, Se, Sb and V is highlydependent on the redox state at which they occurin the solid whereas Mo and W are redoxinsensitive.� Several oxidation states of one element can fre-

quently be found together due to heterogeneousconditions during formation but also due toredox disequilibria because redox transforma-tions are generally kinetically impaired.� The leachability of oxyanions usually is lower than

what is expected on the basis of pure-phase solubil-ity in alkaline wastes or cement. Possible alterna-tive mechanisms are surface adsorption and solidsolution formation with minerals containing Cathat can be found in all alkaline waste types.� In most cases, adsorption by Fe and Al oxides

will only be significant in weathered wastes wherethese oxides have obtained sufficient positivecharge.� To date, no quantitative models have been

developed to assess the relative importance ofadsorption to portlandite, CSH, calcite andsolid solution formation with gypsum. However,the high oxyanion uptake noted in most studieson ettringite, monosulphate and hydrocalumitesuggests that in highly alkaline systems, solidsolution formation with these solid phases ismost likely to control oxyanion leaching. Thedevelopment of efficient models is, however,hampered by a general lack of thermodynamicparameters.

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