Arsenic re-mobilization in water treatment adsorbents under reducing conditions: Part II. XAS and modeling study

S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

a v a i l a b l e a t w w w. s c i e n c e d i r e c t . c o m

w w w. e l s e v i e r. c o m / l o c a t e / s c i t o t e n v

Arsenic re-mobilization in water treatment adsorbents under reducing conditions: Part II. XAS and modeling study Suqin Liu a , Chuanyong Jing a,b,⁎, Xiaoguang Meng a a

Center for Environmental Systems, Stevens Institute of Technology, Hoboken, NJ 07030, USA State Key Laboratory of Environmental Chemistry and Ecotoxicology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, China b

AR TIC LE I N FO

ABS TR ACT

Article history:

The mechanism of arsenic re-mobilization in spent adsorbents under reducing conditions

Received 27 July 2007

was studied using X-ray absorption spectroscopy and surface complexation model

Accepted 13 October 2007

calculations. X-ray absorption near edge structure (XANES) spectroscopy demonstrated that As(V) was partially reduced to As(III) in spent granular ferric hydroxide (GFH), titanium

Keywords:

dioxide (TiO2), activated alumina (AA) and modified activated alumina (MAA) adsorbents

Arsenic

after 2 years of anaerobic incubation. As(V) was completely reduced to As(III) in spent

Re-mobilization

granular ferric oxide (GFO) under 2-year incubation. The extended X-ray absorption fine

XANES

structure (EXAFS) spectroscopy analysis showed that As(III) formed bidentate binuclear

EXAFS

surface complexes on GFO as evidenced by an average As(III)–O bond distance of 1.78 Å and

CD-MUSIC model

As(III)–Fe distance of 3.34 Å. The release of As from the spent GFO and TiO2 was simulated using the charge distribution multi-site complexation (CD-MUSIC) model. The observed redox ranges for As release and sulfate mobility were described by model calculations. © 2007 Elsevier B.V. All rights reserved.

1.

Introduction

The As re-mobilization in water treatment residuals (WTRs) has caused great public concern because of the potential hazard posed with increasing amount of As-containing WTRs as a consequence of the more stringent drinking water standard. Researchers have observed enhanced As leachability in water treatment spent adsorbents under anaerobic conditions (Meng et al., 2001; Ghosh et al., 2004). More than 5 mg-As/L can be released from some spent adsorbents under reducing conditions though they are considered as non-hazardous materials according to the U.S. EPA leaching protocol (Jing et al., 2008). One fundamental difficulty in evaluating the leaching potential of As in the spent adsorbents is the accurate determination of its speciation in different solid phases and its possible presence as adsorbed surface complexes. X-ray absorption spectroscopy (XAS) including X-ray Absorption Near Edge

Structure (XANES) and Extended X-ray Absorption Fine Structure (EXAFS) was developed as a quantitative, short-range structural probe in the 1970's following the pioneering work of Sayers et al. (1971). XAS has since been applied widely in various fields of science and engineering to determine oxidation states and species of chemicals in solid, liquid, and biological samples. Among the most important applications of XAS in the environmental field is the study of As speciation and surface complexation in the aqueous phase and at liquid-mineral interfaces (Manning et al., 1998; La Force et al., 2000; Farquhar et al., 2002; Smith et al., 2005; Ona-Nguema et al., 2005). However, only limited reports have applied XAS in studying Ascontaining spent adsorbents (Jing et al., 2005a). XANES can be used to determine arsenic speciation under in situ conditions. The excitation energy for As(III) is well-separated from As(V) by more than 3 eV. The organic As species can also be distinguished from inorganic As(III) and As(V) (Jing et al., 2005c).

⁎ Corresponding author. State Key Laboratory of Environmental Chemistry and Ecotoxicology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, China. Tel.: +86 10 6284 9523; fax: +86 10 6284 9523. E-mail addresses: [email protected], [email protected] (C. Jing). 0048-9697/$ – see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2007.10.033

138

SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

An EXAFS study demonstrated that As(V) forms inner-sphere bidentate binuclear surface complexes on the surfaces of adsorptive media (Jing et al., 2005a). The primary structure of the As(III) surface complex on goethite was determined to be bidentate binuclear (Manning et al., 1998; Ona-Nguema et al., 2005). Geochemical modeling is urgently needed to improve our understanding of As leaching behavior under reducing conditions. Modeling As mobility in spent adsorbents is a challenge because of the complexity of As reactions including oxidation/ reduction, adsorption/desorption, and dissolution/precipitation. Recently, a multi-site/one-pK surface complexation model has been developed for ion binding to mineral surfaces (Hiemstra and van Riemsdijk, 1996, 1999). The model uses the Pauling concept of charge distribution (CD) and is an extension of the multi-site complexation (MUSIC) approach. In the CD-MUSIC model, surface complexes are not treated as point charges, but are considered as having a spatial charge distribution over their ligands, which are present in two different electrostatic planes. This model has been successfully applied to describe the As adsorption (Jing et al., 2005c; Stachowicz et al., 2006) and As(V) leachability in various spent adsorbents (Jing et al., 2005a). The objective of this study was to investigate As re-mobilization mechanisms in spent media under reducing conditions. XAS analysis and thermodynamic model calculations were

performed to determine the chemical species and reactions of As with the spent adsorbents. The findings of this study will shed further light on As speciation and local coordination environments on spent water treatment adsorbents under reducing landfill conditions, which will help to develop better prediction and treatment for long-term As release.

2.

Materials and methods

2.1.

XAS study

The spent adsorbents including GFH, GFO, TiO2, AA, and MAA, incubated under reducing conditions as described in our previous report (Jing et al., 2008), were used in the XANES and EXAFS study. The spent adsorbents were taken out from the BOD bottle, sealed between two layers of X-ray transparent Kapton tape, and immediately analyzed at the National Synchrotron Light Source (NSLS) in Brookhaven National Laboratory at Upton, NY. The As K-edge spectra were collected at beamline X19A and X18B for samples incubated for 27 days (d) and 2 years (yr), respectively. The Fe K-edge data acquisition was performed at Beamline X18B for 2-yr samples. Standard reference chemicals including NaAsO2, Na2HAsO4·7H2O, MMA, DMA, realgar (AsS), orpiment (As2S3), arsenopyrite (FeAsS),

Fig. 1 – Normalized As K-edge XANES spectra (solid line) and the LC fitting (dashed line) for As reference compounds and spent adsorbent incubated for 27-day (A) and 2-yr (B).

139

S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

pyrite (FeS2), Fe2O3, magnetite (Fe3O4) and FeSO4·7H2O, were also analyzed. Elemental arsenic (As0) and Fe metal foil was used to calibrate the beam energy at 11,868 and 7112 eV, respectively. An energy range of −200 to 200 eV from the K-edge of the target element was used to acquire the spectra, with a 0.5 eV step size from 20 eV below and 50 eV above the absorption edge. The spectra were taken under standard NSLS operation conditions (2.8 GeV and 150–280 mA) with a double-crystal Si (111) monochromator. The spectra were collected in fluorescence mode using a Passivated Implanted Planar Silicon (PIPS) detector at X19A at room temperature and a 13-element energy dispersive Ge detector at X18B at cryogenic temperature (77 K). The detector was positioned 90° to the incident beam. Four scans were collected from each sample, inspected for overall quality and averaged to improve the signal/noise ratio. Quantitative XAS data analysis was performed using the ATHENA and AETEMIS program in the IFEFFIT computer package (Newville, 2001; Ravel and Newville, 2005). The XANES analysis procedure was described in our previous study (Jing et al., 2005b). The raw data, measured in intensities, were converted to μ(E), and averaged spectra were used in the analysis. The spectra were background subtracted using a linear pre-edge fit and normalized to the atomic absorption. A linear combination fitting (LCF) was then conducted with respect to standard references to obtain the weighted percentage of each species in the sample over the relative energy range of −20 to 30 eV. No energy shift was allowed in the fitting procedure.

Table 1 – Relative contents of As(V) and As(III) in incubated samples determined using XANES LC fitting Incubation time

Adsorbent

As(V), wt.

As(III), wt.

Rfactor

χ2

GFH GFO TiO2 AA MAA GFH GFO TiO2 AA MAA

42 ± 3 31 ± 2 63 ± 3 84 ± 4 43 ± 4 18 ± 5 0 28 ± 4 65 ± 5 27 ± 5

58 ± 3 70 ± 2 37 ± 3 16 ± 4 57 ± 4 82 ± 5 100 72 ± 4 35 ± 5 73 ± 5

0.0097 0.0042 0.0100 0.0154 0.0126 0.0105 0.0157 0.0073 0.0103 0.0124

1.08 0.45 1.12 1.97 1.43 1.20 1.70 0.79 1.29 1.43

27-d

2-yr

The EXAFS signal χ(k) was extracted from the measured data using the AUTOBK algorithm (Newville et al., 1993) where k is the photoelectron wave number. The primary quantity for EXAFS is χ(k), the oscillations as a function of photoelectron wave number. χ(k) was weighted by k3 to account for the dampening of oscillations with increasing k. The different frequencies in the oscillations in χ(k) correspond to different near neighbor coordination shells, which can be described and modeled according to the EXAFS equation vðkÞ ¼

X Nj fj ðkÞe2k

2 r2 j

j

kR2j

  sin 2kRj þ dj ðkÞ

where f(k) and δ(k) represent the photoelectron backscattering amplitude and phase shift, respectively, N is the number of neighboring atoms, R is the distance to the neighboring atom, and the σ2 is the Debye–Waller factor representing the disorder in the neighbor distance. The k3 weighted EXAFS in K-space (Å− 1) was Fourier transformed (FT) to produce the radial structure function (RSF) in R-space (Å). The experimental spectra were fitted with single-scattering theoretical phase shift and amplitude functions calculated with the ab initio computer code FEFF6 (Deleon et al., 1991) using atomic clusters generated from the crystal structure of scorodite (FeAsO4·2H2O). The many-body amplitude reduction factor (S20) was fixed at 0.9. The spectrum was fit by first isolating and fitting the first-shell As–O to estimate ΔE0, the difference in threshold energy between theory and experiment, and by fixing the Debye– Waller parameter (σ2) at 0 and coordination number (CN) at 3 for As(III). Then ΔE0 was fixed to the best-fit value from first-shell fitting. The CN of As–O and As–Fe was initially fixed to obtain estimated values for interatomic distances (R) and σ2. Finally, the spectrum was fitted using estimated values for CN, R, and σ2 as starting values. The goodness-of-fit parameters were also calculated and compared including χ2 and R-factor, the relative error of the fit and data. Good fits occur for R b 0.05.

2.2.

Fig. 2 – Normalized Fe K-edge XANES spectra (solid line) and the LC fitting (dashed line) for Fe reference compounds and iron-based adsorbent incubated for 2-yr.

Surface complexation modeling

The charge distribution multi-site complexation (CD-MUSIC) model with the triple plane option was used to describe incubation experimental data for GFO and TiO2 adsorbents. The basic principles of the model have been well documented in the literature (Hiemstra and van Riemsdijk, 1996, 1999). In the CD-MUSIC model, the oxygens on the surface are divided

140

SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

Table 2 – Relative contents of Fe species in 2-yr incubation samples determined using XANES LC fitting Adsorbent FeS2, Fe3O4, Fe2O3, FeSO4, FeAsS, Rwt.% wt.% wt.% wt.% wt.% factor GFH GFO

41 ± 2 27 ± 4

3±1 56 ± 6

6±2 0

50 ± 4 17 ± 5

0 0

0.0036 0.0071

χ2 0.77 0.97

into singly (SO), doubly (S2O), and triply (S3O) coordinated groups to account for the different types of sites. The charge of the surface oxygen is affected by the adsorption of protons and the positive charge of metal ions present in the solid. Application of Pauling's bond valence concept leads to the definition of surface groups and their corresponding charge. Singly coordinated surface hydroxyl groups can form innersphere complexes with As (Hiemstra and van Riemsdijk, 1996; Stachowicz et al., 2006). The charge of the adsorbed arsenic is spatially distributed over two electrostatic planes. The first plane (0-plane) contains the surface groups with the oxygens shared between the surface and arsenic. The second plane (1plane) holds the solution-oriented oxygen groups of the adsorbed arsenic. The outermost plane (2-plane) is the head end of the Gouy–Chapman diffuse double layer (DDL). The surface acid–base reactions, As(V) surface complexation reactions, formation of outer-sphere complexes with background electrolytes, and their corresponding adsorption constants were obtained from our previous report (Jing et al., 2005a). The As(III) surface complexation constants were optimized by fitting model-calculated values to the experimental data. The constants were varied systematically until the difference P ( pH ðexperimental adsorption  calculated adsorptionÞ2 ) between the calculated and observed values reached a minimum. To test the sensitivity of the constant, models were also performed with 10% above and below the best-fit constant. The calculation was performed using the chemical equilibrium program Visual MINTEQ (http://www.lwr.kth.se/English/OurSoftware/ vminteq/index.htm) with the 1-pk TPM adsorption option. In addition to adsorption reactions, the model included aqueous and precipitation reactions with the standard database in − MINTEQ. The redox reactions of As(V)/As(III), SO2− 4 /HS , Fe(III)/ Fe(II) were considered in the model with sweep of Eh from 400 mV to −700 mV under pH = 7 and I = 0.1 M as NaCl.

3.

Results and discussion

3.1.

Redox transformation of As and Fe in the solid phase

Normalized As and Fe K-edge XANES spectra for reference compounds and 2-yr incubation samples are shown in Figs. 1 and 2, respectively. The dominant feature of the XANES spectra is the sharp increase in absorption edge that occurs over a 50 eV interval around the edge position. The excitation potential shifts to higher energy levels for increasing oxidation states (Smith et al., 2005). Although XANES cannot distinguish between the excitation energies of the two arsenic-sulfur compounds, realgar and orpiment (Hansel et al., 2002; Smith et al., 2005), the distinct peak positions of other species allow for As speciation in incubated samples through a robust linear combination fitting (LCF) using a combination of reference

compounds. This technique has been successfully employed to identify and quantify the As species in solids (La Force et al., 2000; Jing et al., 2005b). The LCF results and statistical parameters are shown in Tables 1 and 2. Our previous EXAFS study demonstrated that As(V) is the only arsenic species in the five spent media samples before incubation (Jing et al., 2005a). When spent media were incubated in BOD bottles for 27 days, reduction of As(V) to As(III) was observed. About 58 and 70% of the total As was converted to As (III) in GFH and GFO, respectively (Table 1). The content of As(III)

Fig. 3 – Normalized K3-weighted observed (dashed line) and model-calculated (solid line) As K-edge EXAFS spectra (A) and the corresponding magnitude (dotted line) and real parts (dashed line) of Fourier transform (B, x-axis not corrected for phase shifts) for GFO incubated 2 yr. Numerical fit results (inset) are true interatomic distances (R) corrected for backscatterer phase shifts; CN is average coordination number of backscatterers, and σ2 is the Debye–Waller disorder parameter. The goodness-of-fit parameters: R-factor = 0.0097, χ2 = 9.60.

141

S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

Table 3 – Surface parameters and species used in the CD-MUSIC modeling Species

P0⁎

P1⁎

P2⁎

SOH

S2O

S3O

H

AsO4

H3AsO3

Na

Cl

log K GFO

− 1/2

FeOH FeOH+2 1/2 FeOHNa+ 1/2 FeOH2Cl− 1/2 Fe3O− 1/2 Fe3OH+ 1/2 Fe3ONa+ 1/2 Fe3OHCl− 1/2 Fe2O2AsO−2 2 Fe2O2AsO− 2 TiOH− 1/3 TiOH+2 2/3 TiOHNa+ 2/3 TiOH2Cl− 1/3 Ti2O− 2/3 Ti2OH+ 1/3 Ti2ONa+ 1/3 Ti2OHCl− 2/3 Ti2O2AsO−2 5/3 Ti2O2AsO− 5/3 Surface site density (nm− 2) SOH S 2O S 3O Surface area (m2 g− 1) Inner-sphere capacitance C1 (F m− 2) Outer-sphere capacitance C2 (F m− 2) Adsorbent concentration (g L− 1) Total As concentration (mM) Total sulfate concentration (mM) Total iron concentration (mM)

1 1 −1

1

1 1 1 1

1 1 1 1 1 1 1

1 1 0.5

1 −1 − 1.5 −1

1 1 −1

1

2 2 1 1 1 1

1 1 0.5

− 1.5 −1

1 1 1 2 −1

1 1 1

9.2 0.2 9 9.2 0.2 9 27.2 4.0a

1

5.8 −0.6 4.7

1 1 1 1 1 1

1

1

1

1

5.8 −0.6 4.7 24.8 2.6a

1 1 2 −1

2 2

TiO2

1 1 1 3.45 2.7 140 1.1 5 24.0 1.871 35 0.75

9.0 9.0 180 2.36 5 24.9 1.179 35 0.75

P0⁎ = exp(−FΨ0 / RT), P1⁎ = exp(− FΨ1 / RT), P2⁎ = exp(− FΨ2 / RT); F, the Faraday constant (C mol− 1); R, the gas constant (J mol− 1 K− 1); T, the absolute temperature (K); Ψ0, Ψ1, Ψ2, the electrostatic potential (V) of 0-, 1-, and 2-plane, respectively. a Best-fit value.

in the TiO2, AA and MAA samples was 37, 16, and 57%, respectively. When the samples were incubated for 2-yr, the As(III) content was increased to 82, 100, 72, 35, and 73% for GFH, GFO, TiO2, AA and MAA, respectively. The XANES results indicated that the incubated samples consisted of only two inorganic As species, As(III) and As(V). No arsenopyrite, realgar, orpiment, or organic arsenic compounds were detected. It is interesting to note that the percentage of As(III) in MAA was more than double of that in AA at the same incubation time. MAA was made of iron-impregnated AA and contained ferric (hydr)oxides. This observation showed the effect of ironprecipitating bacteria on the reduction of As(V) in solids. Fe (III)-respiring bacteria could release As(V) by reduction of Fe(III) to Fe(II), which could subsequently be reduced to As(III) (Cummings et al., 1999). It has also been discovered that certain bacteria could use both ferric iron and As(V) as electron acceptors, and adsorbed As(V) could be reduced to As(III) without the reduction of Fe(III) (Zobrist et al., 2000). The LCF results of Fe XANES spectra show that pyrite was 41 and 27% of total Fe in GFH and GFO 2-yr samples, respectively (Fig. 2 and Table 2). Iron sulfide is responsible for trace metal uptake including As in anoxic sediments through adsorption and co-precipitation. Arsenic-rich pyrite (arsenian pyrite) is commonly found in arsenic-rich aquifers and natural deposits

(Savage et al., 2000; Farquhar et al., 2002). The arsenic content in arsenian pyrite ranges from less than 0.5 to 10 wt.% (Simon et al., 1999). However, a recent XANES study using field sediment samples indicates pyrite accounts for a small percentage (b20%) of the total arsenic and is a relatively unimportant host for arsenic in the reduced sediments (Wilkin and Ford, 2006). Bostick and Fendorf (2003) reported that adsorption of As(III) on iron sulfides including FeS and FeS2 can shift the As XANES edge relative to arsenite and orpiment, revealing reduction of As(III). Although pyrite was detected in GFH and GFO samples in this study, the majority of As was presumably unassociated with pyrite. The conclusion was supported by the fact that no reduction of As(III) was observed in the XANES analysis (Fig. 1 and Table 1). With the strong affinity of As(III) to iron oxides surfaces (Dixit and Hering, 2003), As(III) could be stabilized by adsorption on spent adsorbents.

3.2.

EXAFS study

EXAFS spectroscopy was employed to determine the arsenic local coordination environment for the GFO sample. The k3 weighted As K-edge EXAFS spectra are shown in Fig. 3-A, and the corresponding radial structure functions (RSF) are shown in Fig. 3-B as the magnitude and real part of the Fourier

142

SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

transformation (FT) vs. radial distance. The optimal parameters inserted in Fig. 3-B were obtained by fitting the theoretical curves to the experimental spectra. The FT of EXAFS spectra can isolate the contributions of different

coordination shells, in which the peak positions correspond to the interatomic distances. However, these peak positions are uncorrected for phase shift so that they are shifted from the true distance by 0.3–0.5 Å. Based on the fit of the theoretical to the experimental spectra, the first and strongest peak in the FT curve was contributed by 3.2 oxygen atoms at an average distance of 1.78 Å. The As–O interatomic distance suggested the AsO3 triangular pyramid geometry on the surface of GFO. The second shell can be fitted with 1.8 Fe atoms at 3.34 Å. The distances and coordination numbers (CN) of As–O and As–Fe are in good agreement with previously published data (Manning et al., 1998; Farquhar et al., 2002; Ona-Nguema et al., 2005). The EXAFS results show that As (III) formed bidentate binuclear corner-sharing inner-sphere complexes on the surfaces of GFO. The finding of the As(III) surface structure is consistent with previous reports of the association of As(III) with iron oxides (Manning et al., 1998; Farquhar et al., 2002; Ona-Nguema et al., 2005). This stable structure can be used to explain the high affinity of As(III) to GFO. As(V) was reduced to As(III) in GFO samples incubated for 2-yr; however, onlyb10% As(III) released to the aqueous phase due to the strong As(III)–Fe surface complexes.

3.3.

CD-MUSIC modeling

In the present study, the bidentate binuclear As(III) complexes determined with EXAFS were incorporated in the CD-MUSIC model to describe As concentrations in incubation for spent GFO and TiO2. These two adsorbents, representing Fe- and TiO2based adsorbent, were selected as model adsorbents because of their high As content and widespread application. The CDMUSIC model has been employed to describe the As(V) adsorption behaviors on various adsorbents including GFO and TiO2 (Jing et al., 2005a). The surface acid–base reactions, As(V) surface complexation reactions, formation of outer-sphere complexes with background electrolytes, and their corresponding adsorption constants were obtained from our previous study and listed in Table 3. Only singly coordinated surface groups were responsible for the As(III) adsorption in the CD-MUSIC modeling. The charge distribution (CD) value expressed the fraction f of the charge of the central As(III) ion attributed to the surface plane, and the remaining part (1−f) was attributed to the other ligands of the complex, which were located in the 1-plane. The assumption for the CD factor f was based on the theoretical value from the Pauling rule of equal distribution over the surrounding ligands. For bidentate As(III) surface complexes, two ligands were placed in the 0-plane. Thus, f was set to 0.67 (2/3) since As(III) has three ligands. The only adjustable parameter in the model calculation is the formation constant of the inner-sphere As(III) complexes. The optimized adsorption constant of the S2O2AsO complex was 4.0 and 2.6 for Fig. 4 – Model calculated (lines) and experimentally observed soluble concentrations of arsenite (▲), arsenate (◊) (A), sulfate (□) (B), and percentage of As distribution (C) as a function of pe during incubation of GFO sample in a closed reducing system. The solid line is the model result with the best-fit As(III) adsorption constant. The dashed (dotted) line is the model result with 10% above (below) the best-fit As(III) adsorption constant.

S CIE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

GFO and TiO2, respectively, obtained by fitting incubation experimental data for each spent adsorbent. The model calculations considered adsorption of As(III) and As(V) species, reduction of

143

− As(V) to As(III), SO2− 4 to HS , and Fe(III) to Fe(II), and precipitation of orpiment (As2S3), realgar (AsS), arsenopyrite (FeAsS), and pyrite (FeS2). The reactions and corresponding constants are the same as in Table 1 in part I of this series study. The model calculation and experimental observation for concentrations of As and sulfate in GFO and TiO2 samples are shown in Figs. 4 and 5, respectively. The model predicted three redox zones for the release of arsenic in spent adsorbents, which was in agreement with our previous study using Ascontaining water treatment sludge (Meng et al., 2001). Due to their high affinity, both As(V) and As(III) were adsorbed on spent adsorbent surfaces when pe N −3 (Figs. 4–5). In addition to the best-fit As(III) adsorption constant, the model employed values at ±10% from the best-fit constant to test the sensitivity of the model and to examine the contribution of adsorption to the As redox zonation. The model results indicated that varying the adsorption constant only changed the magnitude of soluble As concentrations, but could not change the redox zonation for As release (Figs. 4A and 5A). The model fitted the experimental data well for GFO (Fig. 4). Over 90% of the As was adsorbed on the GFO until pe b − 7, where precipitation as realgar would be the dominant As form (Fig. 4C). The calculated sulfate concentrations suggested that sulfate was reduced to sulfide in a narrow pe range of −3 to − 4. However, the calculated redox zones for As release and sulfate reduction in TiO2 were shifted by about 3 pe units. The shift of redox zonation may be due to the kinetics of the redox reactions (Meng et al., 2001; Jing et al., 2008).

4.

Conclusions

When spent adsorbents were incubated in reducing conditions, reduction of As(V) to As(III), Fe(III) to Fe(II), and sulfate to sulfide occurred. During two years incubation, As(V) was partially reduced to As(III) in spent GFH, AA, MAA and TiO2 adsorbent samples. On the other hand, all As(V) was converted to As(III) and formed inner-sphere bidentate binuclear surface complexes on GFO. The As immobilization in spent adsorbents under reducing conditions was primarily attributed to adsorption. The CD-MUSIC model coupled with redox and precipitation reactions could describe sulfate reduction, As redox zonation, and leaching behavior for GFO samples.

Acknowledgements We acknowledge the staff on beamline X18B and X19A at the National Synchrotron Light Source (NSLS) for their assistance with XAS data collection.

Fig. 5 – Model calculated (lines) and experimentally observed soluble concentrations of arsenite (▲), arsenate (◊) (A), sulfate (□) (B), and percentage of As distribution (C) as a function of pe during incubation of TiO2 sample in a closed reducing system. The solid line is the model result with the best-fit As(III) adsorption constant. The dashed (dotted) line is the model result with 10% above (below) the best-fit As(III) adsorption constant.

144

SC IE N CE OF T H E TOT AL E N V I RO N ME N T 3 9 2 ( 2 00 8 ) 1 3 7–1 44

REFERENCES Bostick BC, Fendorf S. Arsenite sorption on troilite (FeS) and pyrite (FeS2). Geochim Cosmochim Acta 2003;67:909–21. Cummings DE, Caccavo Jr F, Fendorf S, Rosenzweig RF. Arsenic mobilization by the dissimilatory Fe(III)-reducing bacterium Shewanella alga BrY. Environ Sci Technol 1999;33:723–9. Deleon JM, Rehr JJ, Zabinsky SI, Albers RC. Ab initio curved-wave X-ray absorption fine structure. Phys Rev, B 1991;44:4146–56. Dixit S, Hering JG. Comparison of arsenic(V) and arsenic(III) sorption onto iron oxide minerals: implications for arsenic mobility. Environ Sci Technol 2003;37:4182–9. Farquhar ML, Charnock JM, Livens FR, Vaughan DJ. Mechanisms of arsenic uptake from aqueous solution by interaction with goethite, lepidocrocite, mackinawite, and pyrite: an X-ray absorption spectroscopy study. Environ Sci Technol 2002;36:1757–62. Ghosh A, Mukiibi M, Ela W. TCLP underestimates leaching of arsenic from solid residuals under landfill conditions. Environ Sci Technol 2004;38:4677–82. Hansel CM, La Force MJ, Fendorf S, Sutton S. Spatial and temporal association of As and Fe species on aquatic plant roots. Environ Sci Technol 2002;36:1988–94. Hiemstra T, van Riemsdijk WH. A surface structural approach to ion adsorption: the charge distribution (CD) model. J Colloid Interface Sci 1996;179:488–508. Hiemstra T, van Riemsdijk WH. Surface structural ion adsorption modeling of competitive binding of oxyanions by metal (hydr) oxides. J Colloid Interface Sci 1999;210:182–93. Jing C, Liu S, Patel M, Meng X. Arsenic leachability in water treatment adsorbents. Environ Sci Technol 2005a;39:5481–7. Jing C, Liu S, Meng X. Arsenic leachability and speciation in cement immobilized water treatment sludge. Chemosphere 2005b;59:1241–7. Jing C, Meng X, Liu S, Baidas S, Patraju R, Christodoulatos C, Korfiatis GP. Surface complexation of organic arsenic on nanocrystalline titanium oxide. J Colloid Interface Sci 2005c;290:14–21. Jing C, Liu S, Meng X. Arsenic remobilization in water treatment adsorbents under reducing conditions: Part I. Incubation study. Sci Total Environ 2008;389:188–94. La Force MJ, Hansel CM, Fendorf S. Arsenic speciation, seasonal transformations, and co-distribution with iron in a mine

waste-influenced palustrine emergent wetland. Environ Sci Technol 2000;33:3937–43. Manning BA, Fendorf SE, Goldberg S. Surface structures and stability of arsenic(III) on goethite: spectroscopic evidence for inner-sphere complexes. Environ Sci Technol 1998;32:2383–8. Meng X, Korfiatis GP, Jing C, Christodoulatos C. Redox transformations of arsenic and iron in water treatment sludge during aging and TCLP extraction. Environ Sci Technol 2001;35:3476–81. Newville M. IFEFFIT: interactive XAFS analysis and FEFF fitting. J Synchrotron Radiat 2001;8:322–4. Newville M, Livins P, Yacoby Y, Rehr JJ, Stern EA. Near-edge X-ray absorption fine structure of Pb — a comparison of theory and experiment. Phys Rev, B 1993;47:14126–31. Ona-Nguema G, Morin G, Juillot F, Calas G, Brown GE. EXAFS analysis of arsenite adsorption onto two-line ferrihydrite, hematite, goethite, and lepidocrocite. Environ Sci Technol 2005;39:9147–55. Ravel B, Newville M. ATHENA, ARTEMIS, HEPHAESTUS: data analysis for X-ray absorption spectroscopy using IFEFFIT. J Synchrotron Radiat 2005;12:537–41. Savage KS, Tingle TN, O'Day PA, Waychunas GA, Bird DK. Arsenic speciation in pyrite and secondary weathering phases, Mother Lode Gold District, Tuolumne County, California. Appl Geochem 2000;15:1219–44. Sayers DE, Stern EA, Lytle FW. New technique for investigating noncrystalline structures — Fourier analysis of extended X-ray absorption fine structure. Phys Rev Lett 1971;27:1204–7. Simon G, Huang H, Penner-Hahn JE, Kesler SE, Kao LS. Oxidation state of gold and arsenic in gold-bearing arsenian pyrite. Am Mineral 1999;84:1071–9. Smith PG, Koch I, Gordon RA, Mandoli DF, Chapman BD, Reimer KJ. X-ray absorption near-edge structure analysis of arsenic species for application to biological environmental samples. Environ Sci Technol 2005;39:248–54. Stachowicz M, Hiemstra T, van Riemsdijk WH. Surface speciation of As(III) and As(V) in relation to charge distribution. J Colloid Interface Sci 2006;302:62–75. Wilkin RT, Ford RG. Arsenic solid-phase partitioning in reducing sediments of a contaminated wetland. Chem Geol 2006;228:156–74. Zobrist J, Dowdle PR, Davis JA, Oremland RS. Mobilization of arsenite by dissimilatory reduction of adsorbed arsenate. Environ Sci Technol 2000;34:4747–53.