Catalytic Coal Gasification Process Simulation with Alkaline ...
Leaching mechanisms of oxyanionic metalloid and metal species in alkaline solid wastes: A review
Transcript of Leaching mechanisms of oxyanionic metalloid and metal species in alkaline solid wastes: A review
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Applied Geochemistry 23 (2008) 955–976
www.elsevier.com/locate/apgeochem
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 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 966927/$ - 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).
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
le2
Ran
ges
of
tota
lco
nte
nt
of
oxy
anio
nfo
rmin
gel
emen
tsex
pre
ssed
inm
g/k
gin
MS
WI
resi
du
es(F
allm
anan
dH
artl
en,
1994
;IA
WG
,19
97;
van
der
Slo
ot
etal
.,19
97;
US
GS
,20
05),
FF
C-
resi
du
es(E
ary
etal
.,19
90;
van
der
Slo
ot
etal
.,19
97;
US
GS
,20
05)
and
met
allu
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
sid
ues
FF
C-r
esid
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.
1–2
0.05
–10
0.4–
300.
7–30
0.1–
100.
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
–10,
000
40–4
00W
––
30N
AN
AN
AN
AN
A<
1340
00.
4–3
NA
=n
ot
avai
lab
le.
aA
irp
oll
uti
on
con
tro
l.b
Flu
ega
sd
esu
lph
uri
sati
on
.
958 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976
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).
G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976 959
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).
960 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976
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%).
G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976 961
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.
962 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976
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.
G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976 963
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)
964 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976
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.
G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976 965
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.
966 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976
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)
G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976 967
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).
968 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976
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.
G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976 969
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).
970 G. Cornelis et al. / Applied Geochemistry 23 (2008) 955–976
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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