Adsorption of arsenic(III) and arsenic(V) from groundwater using natural siderite as the adsorbent

Journal of Colloid and Interface Science 315 (2007) 47–53 www.elsevier.com/locate/jcis

Adsorption of arsenic(III) and arsenic(V) from groundwater using natural siderite as the adsorbent Huaming Guo a,∗ , Doris Stüben b , Zsolt Berner b a School of Water Resources and Environment, China University of Geosciences, Beijing 100083, People’s Republic of China b Institute for Mineralogy and Geochemistry, University Karlsruhe (TH), Karlsruhe 76131, Germany

Received 4 April 2007; accepted 19 June 2007 Available online 26 July 2007

Abstract Batch and column tests were performed utilizing natural siderite to remove As(V) and As(III) from water. One hundred milligrams of siderite was reacted at room temperature for up to 8 days with 50 mL of 1000 µg/L As(V) or As(III) in 0.01 M NaCl. Arsenic concentration decreased exponentially with time, and pseudoequilibrium was attained in 3 days. The estimated adsorption capacities were 520 and 1040 µg/g for As(V) and As(III), respectively. Column studies show that effluent As was below 1.0 µg/L after a throughput of 26,000 pore volumes of 500 µg/L As water, corresponding to about 2000 µg/g of As load in the filter. Results of scanning electron microscopy (SEM) and transmission electron microscopy (TEM) reveal that high As retention capacity of the filter arose from coprecipitation of Fe oxides with As and subsequently adsorption of As on the fresh Fe oxides/hydroxides. Arsenic adsorption in the filter from As-spiked tap water was relatively lower than that from artificial As solution because high HCO− 3 concentration restrained siderite dissolution and thus suppressed production of the fresh Fe oxides on the siderite grains. The TCLP (toxicity characteristic leaching procedure) results suggest that these spent adsorbents were inert and could be landfilled. © 2007 Elsevier Inc. All rights reserved. Keywords: Arsenate; Arsenite; Drinking water; Removal; Retention

1. Introduction Groundwater enriched with As species such as arsenate (As(V)), arsenite (As(III)), and organic arsenic has become one of the most serious problems in water environment, especially in the southeast of Asia, including West Bengal, India, Bangladesh, and China [1,2]. It is particularly worse when the groundwater is utilized as drinking water [3,4]. Although organic As species can be presented as a result of in situ biomethylation, inorganic As as As(III) and As(V) are generally considered to be the dominant species in natural water. The oxidation state of As depends primarily on pH and redox conditions, with As(V) being the most stable form under aerobic conditions as the pH-dependent deprotonated oxyan2− ions of arsenic acid (H2 AsO− 4 and HAsO4 ) and As(III) the chemically dominant forms in reducing environment as a neu* Corresponding author. Fax: +86 10 8232 1081.

E-mail address: [email protected] (H. Guo). 0021-9797/$ – see front matter © 2007 Elsevier Inc. All rights reserved. doi:10.1016/j.jcis.2007.06.035

tral species (i.e., pKa1 = 9.2) at natural pH. Therefore, the As(III) is more difficult to remove from water at neutral pH by means of adsorption and coprecipitation due to the lack of electrostatic attraction [5,6]. However, most As-enriched groundwater is generally dominated by As(III), up to 96% of total As [2,7]. Furthermore, As(III) is about 60 times more toxic than As(V) [8,9]. Thus, the removal of As(III) from drinking water has received significant attention, as well as As(V) [10–14], and it is of major concern to many water utilities and governmental agencies. Among a variety of adsorbents for As removal, natural geomaterials are promising because they are relatively cheap, readily available in different particle sizes, and therefore may easily be used as column fillings in small-scale treatment plants. Arsenic adsorption onto natural zeolite and volcanic stone [15], natural iron ores [16], oxisol [17], red mud [18], and ferruginous manganese ore [19] has been examined intensively. Contrastingly, few studies on As removal by natural siderite have been reported [20,21], where the combination of natural hematite and natural siderite has been utilized for As reten-

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tion. In comparison with natural hematite, the natural siderite generally had a higher adsorption capacity for As species in batch tests, and preferentially adsorbed inorganic As in a relatively wider pH range [20]. Nevertheless, information of As adsorption on natural siderite is very limited and more testing is necessary. There is a need to study the kinetics and isotherms of the siderite in adsorbing As from water, and to evaluate the effectiveness of a siderite-packed column for As removal in achieving the permissible concentration in drinking water. The objectives of this study were to (1) evaluate the effectiveness of natural siderite in removing As(V) and As(III) from water; (2) examine the feasibility of the siderite as filter fillings for As removal from As-contaminated water; (3) evaluate the potential utilization of the siderite as permeable reactive barrier media in remediation of As contamination in groundwater. 2. Materials and methods 2.1. Materials Natural siderite, reported as SIO4 by Guo et al. [20], was used in this study. Detailed chemical and mineralogical components of the sample have been provided earlier in Guo et al. [20]. Prior to experiments, the mineral sample was ground and sieved into five main particle size fractions, including <0.04, 0.04–0.08, 0.08–0.10, 0.10–0.25, and 0.25–0.50 mm. All reagents used were of analytical grade. Stock solutions (100 mg/L As) were prepared from sodium arsenite (AsNaO2 ; >99.0%, Fluka Chemical) for As(III) and sodium arsenate (Na2 HAsO4 ·7H2 O; >98.5%, Fluka Chemical) for As(V). In order to maintain a relatively constant ionic strength, all artificial As solutions contained 0.01 M NaCl as background electrolyte. Tap water spiked with 100 µg/L As(V) and 100 µg/L As(III) was also used as inflow water to examine the effect of multicomponents on the adsorption performance of siderite-packed filters. Chemical composition of the tap water is shown in Table 1, indicating it was of Ca–HCO3 type. All glassware and sample bottles were washed with a detergent solution, rinsed with tap water, soaked with 1.0% subboiled HNO3 for at least 12 h, and finally rinsed with Milli-Q water three times. 2.2. Adsorption experiments Batch experiments were conducted to obtain rate and equilibrium data by reacting 50 mL of solution containing a defined As concentration with 0.1 g of sample material of known grain size range at neutral pH at room temperature of 20 ± 2 ◦ C in

100 mL light polyethylene bottles. Most experiments were, unless otherwise stated, performed with a sample of a grain size range of 0.10–0.25 mm. After the required reaction time, the suspension liquid was decanted and filtered through a 0.45 µm cellulose acetate filter. The filtered solution was analyzed for As species. To study the effect of contact time on As uptake, experiments were carried out with initial As(V) or As(III) concentrations of 1000 µg/L. Isotherm studies were conducted with initial As(V) or As(III) concentrations between 250 and 2000 µg/L, and a contact time of 72 h. To investigate the effect of the grain size range on As adsorption, batch tests were carried out with grain size ranges of <0.04, 0.04–0.08, 0.08–0.10, 0.10–0.25, and 0.25–0.50 mm, and initial As(V) or As(III) concentrations of 2000 µg/L. Column experiments were conducted with the siderite of 0.10–0.25 mm particle size fraction (SO), or with sideritecoated sand (SCS) prepared by homogeneously mixing 0.04– 0.08 mm siderite with 1.0% HNO3 -rinsed 0.10–0.25 mm quartz sand at a ratio of 1:5 in weight. The detailed procedure of SCS preparation was described in Guo et al. [21]. Plexiglas columns with an inner diameter of 30 mm and a height of 150 mm, as fixed-bed upflow reactor, were used in the column study, which yields a working volume of about 100 mL. Feedwater containing 250 µg/L As(V) and 250 µg/L As(III) was pumped through a SO-packed filter (SO-C1) with a peristaltic pump (Model 205S, Watson Marlow Company) at a flow rate 2.15 mL/min. In order to investigate the effect of multicomponent on As removal, another SO-packed column (SO-C2) was penetrated by As-spiked tap water containing 100 µg/L As(III) and 100 µg/L As(V). A SCS-packed column (SCS-C) was used to filter 500 µg/L As(V) solution at a flow rate of 1.48 mL/min to examine the capability of siderite-coated sand in removing As from water. Effluent solutions from the filters were collected at regular intervals and analyzed for residual As species. 2.3. Analytical methods A flow injection hydride generation system (FIAS 200, Perkin-Elmer) coupled with atomic absorption spectrometry (AAS 4001, Perkin-Elmer) was used for As speciation analyses. With regard to different As species, hydride generation was carried out in different acid media in the presence of reductant. After filling the sample loop the sample plug was pushed out by the acidic carrier solution and completely mixed with the selected acid in the mixing coil before reduction took place by entraining the potassium tetrahydroborate. The detailed analysis procedure was described by Rüde and Puchelt [22]. The detection limits of total As, As(III), and As(V) were 1, 0.2, and 0.3 µg/L, respectively. Major anions were determined by ion

Table 1 Chemical composition of tap water used in the column experiment (major ions in mg/L, trace elements in µg/L, and EC in µS/cm) pH 7.31

ECa 683

HCO− 3 333

Cl− 21.0

NO− 3 4.94

SO2− 4 76.2

HPO2− 4 <0.01

K+ 1.83

Na+ 10.7

Ca2+ 113

Mg2+ 12.5

Cr 0.13

Mn 4.35

Fe 8.40

Co 0.09

Ni 1.50

Cu 58.3

Zn 1437

As 0.35

Rb 0.77

Sr 357

Mo 0.46

Cd 1.95

Sb 0.37

Ba 74.2

Pb 2.73

a EC means electrical conductivity.

H. Guo et al. / Journal of Colloid and Interface Science 315 (2007) 47–53

chromatography (IC, DX-100, DIONEX), major cations (i.e., K+ , Ca2+ , Na+ , and Mg2+ ) and Mn and Fe by flame atomic absorption spectrometry (Model 1100B, Perkin-Elmer), and trace elements by high-resolution inductively coupled plasma mass spectrometry (HR-ICP-MS, Axiom, VG Elemental). Electrical conductivity (EC) and pH were monitored by a portable WTW EC meter (Model LF330, coupled with an EC probe of TetraCon 325) and pH meter (Model pH330, coupled with a pH probe of SenTix 43-1), respectively. The pH meter was calibrated prior to use. Morphological analysis of the pristine and spent SO grains was performed by scanning electron microscopy (SEM) using a LEO 1530 Gemini microscope (at 10 kV) with energydispersive X-ray analyses. For transmission electron microscopy (TEM), the spent SO grains were dispersed in glue and embedded in a hole at the center of a 3 mm metallic disk with a thickness of 500 µm. Each disk was mechanically grinded to a thickness between 100 and 150 µm. The grinding was followed by a dimpling process until the specimen thickness reached 10–50 µm in the thinnest regions. The final thinning procedure was performed by Ar+ ion bombardment with ion energy of 6 keV. The ion-thinning sections were observed with a Zeiss EM 912 TEM equipped with an X-ray spectrometer for elemental analysis. Energy dispersive spectroscopy (EDS) was performed at 120 kV with a beam current of 1 nA for ∼100 s. 2.4. The toxicity characteristic leaching procedure (TCLP) test The spent adsorbents generated in the column filters were tested with TCLP to determine the spent filling materials as inert or hazardous in terms of leachability of adsorbed As [23]. In the TCLP, the spent adsorbent was treated with acetate buffer (5.7 mL of glacial CH3 COOH added to 500 mL of Milli-Q water, plus 64.3 mL of 1 N NaOH and diluted to 1 L, pH 4.93), with a liquid/solid ratio of 20. The extraction was achieved by agitating the specimens for 18 h in a shaker, after which the liquid phase was separated off using 0.45 µm cellulose acetate filter and analyzed for total As by HR-ICP-MS. 3. Results and discussion 3.1. Effect of contact time The effect of contact time (0.5–194 h) on As adsorption in As(V) batches and As(III) batches is shown in Fig. 1. The results clearly demonstrate that adsorption efficiencies increased rapidly with an increase in contact time up to 24 h, and maximum removal capacities of 386 and 458 µg/g were achieved with a contact time of 194 h in the As(V) batch and the As(III) batch, respectively. Furthermore, the amount of adsorbed As from As(V) solution was a little greater than that from As(III) solution at a contact time between 0.5 and 2.5 h, while less than that from As(III) solution afterward. However, with the reaction condition of adsorbent dosage of 10 g/L and contact time of 24 h, the siderite with grain size fraction of 0.25–0.50 mm was previously found to remove As from As(V) solution (57.4 µg/g)

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Fig. 1. Effect of contact time on As adsorption from As(V) solution and As(III) solution by the siderite, with reaction conditions: ionic strength = 0.01 M NaCl, adsorbent dosage = 2 g/L, initial As concentration = 1000 µg/L, grain size fraction = 0.10–0.25 mm.

much more than from As(III) solution (22.1 µg/g) at neutral pH [20]. This difference suggests that the reaction conditions and the grain size fraction influenced As adsorption on the siderite. The kinetics data also indicate that the As removal from As(V) solutions and As(III) solutions mainly occurred within 72 h and there was no significant change in residual As concentrations after this time up to 194 h, which means that a pseudoequilibrium of As adsorption was roughly attained after 72 h. The pseudoequilibrium time of 72 h was greater than that for As(V) adsorption on natural hematite (0.25–0.50 mm) reported as 24 h [20]. 3.2. Effect of initial As concentration The As loadings on the adsorbents (Qe , µg/g) were calculated from the pseudoequilibrium As concentrations (Ce , µg/L) using mass balance. The adsorption isotherm data (Qe vs Ce ) were fitted to Langmuir and Freundlich isotherm models [Eqs. (1) and (2), respectively] [24,25]. Nonlinear regression was performed with Statistica for Windows using the QuasiNewton method. The Langmuir isotherm model assumes a monolayer surface coverage limiting the adsorption due to the surface saturation, while the Freundlich isotherm model is an empirical model allowing for multilayer adsorption [26]. Q0 bCe , 1 + bCe Qe = KF Cen , Qe =

(1) (2)

where Ce is the pseudoequilibrium concentration in the solution (µg/L), Qe is the amount adsorbed on the adsorbent at pseudoequilibrium (µg/g), Q0 is the maximum adsorption capacity (µg/g), b is a constant related to the adsorption energy (L/µg), KF is the Freundlich constant denoting the adsorption capacity of the adsorbent [(µg/g)(L/µg)n ], and n is the adsorption intensity parameter. Values of 0.1 < n < 1 show favorable adsorption of As onto adsorbents [27].

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All the adsorption data obtained were fitted to both models as shown in Fig. 2. However, with the adsorption of both As(V) and As(III), calculated correlation coefficients for the Freundlich isotherm model were a little higher compared to those for the Langmuir (Table 2). Therefore, the Freundlich isotherm yielded a better fit to the experimental data with regard to the siderite with a grain size range of 0.10–0.25 mm. In contrast, the As(V) adsorption on the same siderite with a grain size range of 0.25–0.50 mm followed Langmuir isotherm more closely in comparison with the Freundlich isotherm [20]. Because the Freundlich isotherm model [Eq. (2)] is an empirical model allowing for multilayer adsorption, it could be speculated that the multilayer adsorption would be involved in the process of As removal by the fine grains (0.10–0.25 mm), whereas the monolayer adsorption by the coarse grains (0.25–0.50 mm), due to the greater surface area of the fine grains in comparison with the coarse ones [20]. 3.3. Effect of grain size The results demonstrate that the As removal from both As(III) solution and As(V) solution increased with the decrease in grain size fractions (see Fig. S1 in supporting information). The siderite with grain size fractions between 0.04–0.08 and

Fig. 2. Langmuir and Freundlich plots for As adsorption from As(V) solution and As(III) solution on the siderite, with reaction conditions: ionic strength = 0.01 M NaCl; adsorbent dosage = 2 g/L; initial concentrations = 250–2000 µg/L; contact time = 72 h; grain size fraction = 0.10–0.25 mm. Table 2 Correlation coefficients and isotherm parameters of both Langmuir and Freundlich models (reaction conditions: ionic strength = 0.01 M NaCl; adsorbent dosage = 2 g/L; initial concentrations = 250–2000 µg/L; contact time = 72 h; grain size fraction = 0.10–0.25 mm) Langmuir model

As(V) As(III)

Freundlich model

Q0 (µg/g)

b

r2 (%)

KF (µg/g)(L/µg)n

n

r2 (%)

516 1040

0.0066 0.0019

92.5 98.2

64.9 13.1

0.281 0.585

99.7 99.3

0.25–0.50 mm removed As from As(III) solution more than from As(V) solution, while with the grain size fraction of <0.04 mm slightly less than from As(V) solution. It was also found that Mn was released into solution during adsorption reaction, the concentration of which in the As(III) batch was greater than that in the As(V) batch with regard to each grain size fraction investigated (see Fig. S2 in supporting information). The higher Mn concentration possibly means more Mn compounds had been reduced in the As(III) batch, in comparison with the As(V) batch. It has been known that Mn compounds oxidize As(III) readily [28,29]. The processes of As(III) removal by the siderite were not fully understood, but generally believed that the higher efficiency of As removal from As(III) solution possibly arose from the additional adsorption sites on the fresh surface from which Mn compounds were introduced into the solution. From the results of the kinetic study (Fig. 1), it is deduced that the enhanced adsorption may have taken effect when the adsorption proceeded for 2.5 h, after which the amount of As removed from As(III) solution exceeded that from As(V) solution. 3.4. Column studies Column experiments were conducted to evaluate the feasibility of the siderite as filter filling in removing As from contaminated water. Total As in the SO-C1 effluent was below 1.0 µg/L after a throughput of 26,000 pore volumes (PV: the volume of water required to replace water in a certain volume of saturated porous media; it is about 30 mL in the studied column) of As solution containing 250 µg/L As(III) and 250 µg/L As(V) (Fig. 3A). With a column packed with 0.25–0.50 mm siderite in the lower half and 0.25–0.50 mm hematite in the upper half, effluent As was close to 1.00 µg/L after a throughput of 7200 PV of the same As solution (data not shown). It is believed that adsorption sites were still available for As retention in the filter, and more As water was able to be remediated before the European drinking water standard (10 µg/L As) was exceeded. A mass balance calculation indicates that total As load in the filter was about 2000 µg/g before 1 µg/L of As breakthrough was observed, which was higher than the maximum As(III) adsorption of 1040 µg/g calculated from the Langmuir isotherm (Table 2). Bang et al. also found that the As load on the TiO2 in the filter was higher than that predicted by the batch adsorption isotherm [30]. They suggested that it should have resulted from the long contact time of As with the adsorbent in the filter. In our study, the higher As load of the filter likely attributed to the freshly formed Fe hydroxides in the column during its operation. A detailed explanation is provided later. The natural siderite as a filling material had a higher adsorption capacity, in comparison with other natural geomaterials (Table 3). In order to examine the effect of multicomponents on column performance, SO-C2 was used to treat As-spiked tap water. Breakthrough curve of the filter is shown in Fig. 3B. In comparison with artificial As solution, removal efficiency from As-spiked tap water was relatively lower, showing that the pres2− − ence of competing anions (i.e., NO− 3 , SO4 , and HCO3 ) in

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Table 3 Arsenic removal by natural solids reported in the literature at ambient temperature

(A)

(B)

Adsorbent

Arsenic species

Adsorption Removal References maxima efficiency (µg/g) (%)

Natural iron ores Ferruginous manganese ore Natural feldspar Natural hematite Natural manganese oxide Clinoptilotile-rich tuffs

As(V) As(III)

400 537

99 72

Zhang et al. [16] Chakravarty et al. [19]

As(V) As(V) As(V)

208 219 200

97 100 –

Singh et al. [31] Singh et al. [31] Ouvrard et al. [32]

As(V)

100

–

Elizalde-González et al. [33]

tap water depressed As adsorption [20,34]. Although phosphate strongly competes the adsorption site with As [20], it would have few effects on As adsorption with a concentration of less than 0.01 mg/L in the As-spiked tap water. The As-drinking water standard of 10 µg/L was exceeded after 7000 PV total throughputs. Arsenic(V) was the dominant As species in the effluents, with the ratio of As(V)/total As of around 0.80. It means that the siderite column reactor removed As(III) more efficiently than As(V). Column experiments were also carried out to test the retention ability of SCS as a column filling. Fig. 3C shows development of As concentration in the effluent from SCS-C filter treating 500 µg/L As(V) solution. The breakthrough curve demonstrates an efficient elimination of As(V). The As concentration remained below the European drinking water standard of 10 µg/L up to a total throughput of 11,600 PV. The As load in the filter was 1090 µg/g, which exceeded the adsorption capacity of 0.04–0.08 mm siderite as a coating material observed as 940 µg/g in the batch test. The good performance of the SCS-C filter possibly indicates that the fine siderite was coated on the quartz sand acting as a good scavenger for As. In the TCLP test, the As released from the used filling materials into the extraction solution (1:20 in weight) at pH 4.93 did not exceed 400 µg/L. It is much less than the established U.S. EPA standard of 5 mg/L, indicating that these spent adsorbents were inert and could be landfilled. 3.5. Mechanisms of As removal

(C) Fig. 3. Development of As concentration in the effluent from SO-C1 (A), SO-C2 (B), and SCS-C (C). (A) As solution containing 250 µg/L As(V) and 250 µg/L As(III) at a flow rate of 2.15 mL/min; (B) As-spiked tap water containing 100 µg/L As(V) and 100 µg/L As(III) at a flow rate of 2.15 mL/min; (C) 500 µg/L As(V) solution at a flow rate of 1.48 mL/min.

The structural differences between the pristine and the used siderite from SO-C1 were studied by SEM and are presented in Fig. 4. It can be noted that there are many small particles (∼10 nm diameter) presented on the surface of the pristine siderite, whereas the used siderite is mainly covered by needlelike goethite (or ferrihydrite) (∼50 nm diameter). The goethite or ferrihydrite loosely nested on the siderite matrix with two main directions (Fig. 4B), which drastically enhanced the specific surface area. The coating of the fresh Fe oxides was also verified by TEM images of the used siderite (Fig. 4C). Siderite matrix (dark) is located on the top left of Fig. 4C, while the fresh Fe oxides (light) are on the bottom right. It can be seen that the fresh Fe oxides were attached to the surface of the

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Fig. 4. SEM images of the pristine siderite (A) and the used siderite (B), and TEM image of the used siderite (C) and EDS spectra for the coating (D). (In the TEM image the dark spot is siderite, while the light spot is fresh Fe-oxide coating.)

siderite, and the coating was about 300 nm in thickness. The EDS analysis (Fig. 4D) of the coating shown in Fig. 4C indicates that the fresh mineral was dominated by the Fe oxides and contained high As. The high As concentration presented in the Fe-oxide coating possibly implies that the fresh Fe-oxide coating had a very high affinity for As [35]. In comparison with the relatively well aerated flowthrough column, the development of fresh Fe(III)-oxide coatings was less significant in the tightly closed bottles of the batch experiments. Guo et al. observed that the thickness of Fe(III)-oxide coatings on the siderite particles taken from the column was much greater than that from the batch experiment by means of the µ-synchrotron X-ray fluorescence analysis (μ-XRFA) [20]. The high As load in the filter possibly arose from both coprecipitation of the Fe oxides with As and subsequent adsorption of As on the fresh Fe oxides/hydroxides. Adsorption from Asspiked tap water in SO-C2 was lower than that from artificial As solution in SO-C1, which likely attributed to the limited fresh Fe oxides developing in SO-C2. It is observed that the appearance of the filling in SO-C2 remained relatively stable during the entire experiment, while the siderite in SO-C1 turned red after the column run for 4 days. Prior to the formation of the Fe oxides, the dissolution of siderite may have occurred to provide Fe2+ . Because the As-spiked tap water contained high HCO− 3 (Table 1), it would restrain the dissolution of FeCO3 and consequently suppress the production of the fresh Fe oxide. Therefore, the formation of the fresh Fe oxide was the key factor in improving As removal efficiency of the column filling. Preliminary μ-XANES investigations were carried out at the SUL beamline of the ANKA synchrotron facility, Karlsruhe,

Germany, on the speciation of As and Fe in the used siderite [36]. The Fe μ-XANES spectra show a superposition of siderite and Fe oxyhydroxides happened near the surface, which provided more evidence to prove the formation of the Fe-oxide coating. Furthermore, the speciation of adsorbed As indicates that both As(III) and As(V) were present in the rim area of the grains, which means that the siderite filter directly adsorbed both As(III) and As(V) from aqueous solution. After completion of As adsorption, the coexistence of As(V) and As(III) was also found on the Fe0 surface [37]. Since both As(III) and As(V) have been found in high As groundwater, natural siderite provides a promising material for the permeable reactive barrier technology for in situ remediation of As-contaminated groundwater. Literature is lacking regarding As removal mechanisms by natural siderite. It appears that electrostatic interaction, surface complexation, and specific adsorption are important mechanisms for As removal by both siderite and Fe oxides which were continuously developed on the surface of the siderite. Research on the interactions between As and Fe oxides thus should have direct implications for the As adsorption. Spectroscopic methods such as EXAFS and IR have provided information on the molecular structures of As complexes on Fe oxides [38–44]. Spectroscopic evidence suggests that As(V) predominantly forms inner sphere bidentate surface complexes with goethite [43,44]. A pressure-jump relaxation kinetics study proposed a two-step process for As(V) adsorption by goethite [39]. A FTIR spectroscopic study supported inner sphere complexation of both As(V) and As(III) on goethite [38].

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4. Conclusions This study confirms that the natural siderite is effective in removing both As(V) and As(III) from solution. Arsenic adsorption increased exponentially with contact time in the batch test, and generally reached pseudoequilibrium at a contact time of 3 days. Arsenic removal from As(III) solution was much higher than from As(V) solution. Since As-enriched groundwater is generally dominated by As(III) [2,7,45,46], the siderite can be utilized to efficiently remove As from groundwater without preoxidation of As(III) to As(V). Column studies show that both the siderite filter and the sand-siderite filter had a high As retention capacity, with As loads of 2000 and 1090 µg/g, respectively. The high efficiency for As removal was attributed to the adsorption of As on the pristine siderite and afterward on the fresh Fe-oxide coatings. The availability of the natural siderite at a low cost makes it a promising material for in situ remediation of As-contaminated groundwater resources. The positive effects of residence time may be advantageous to potential permeable reactive barriers (PRB) composed of the natural siderite to remediate As in the field. More studies are needed along this path to optimize the remediation efforts. Factors influencing the removal efficiency should be among the areas of future research. These factors include temperature, dissolved organic constituents, redox condition, and microorganisms in groundwater. Acknowledgments H.M.G. is grateful to the Alexander von Humboldt Foundation, Germany, for providing a Research Fellowship to carry out this research. Funding for this research has also been provided by the Natural Science Foundation of China (No. 40572145). The authors express their appreciation for analyses support of our colleagues M. Fotouhi (TEM), V. Zibat (SEM), C. Moessner (HR-ICP-MS), T. Neumann (FI-AAS), and G. Preuss (GFAAS). Supporting information Supporting information data for this article may be found in the online version at DOI: 10.1016/j.jcis.2007.06.035. References [1] P.L. Smedley, D.G. Kinniburgh, Appl. Geochem. 17 (2002) 517. [2] H.M. Guo, Y.X. Wang, G.M. Shpeizer, S. Yan, J. Environ. Sci. Health A 38 (2003) 2565. [3] National Research Council, Arsenic in Drinking Water: 2001 Update, National Academy Press, Washington, DC, 2001. [4] J.O. Nriagu, in: W.T. Frankenberger (Ed.), Environmental Chemistry of Arsenic, Dekker, New York, 2002, p. 1. [5] E.O. Kartinen, J.M. Martin, Desalination 103 (1995) 79. [6] M. Bissen, F.H. Frimmel, Acta Hydrochim. Hydrobiol. 31 (2003) 97.

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