Environ. Sci. Technol. 2005, 39, 1291-1298
Removal of Arsenic(III) from Groundwater by Nanoscale Zero-Valent Iron SUSHIL RAJ KANEL,† BRUCE MANNING,‡ LAURENT CHARLET,§ AND H E E C H U L C H O I * ,† Department of Environmental Science and Engineering, Gwangju Institute of Science and Technology (GIST), 1 Oryong-dong, Buk-gu, Gwangju 500-712, Korea, Department of Chemistry and Biochemistry, San Francisco State University, 1600 Holloway Avenue, San Francisco, California 94132, and Environmental Geochemistry Group, LGIT, University of Grenoble I/CNRS, BP 53, F-38041 Grenoble Cedex, France
Nanoscale zero-valent iron (NZVI) was synthesized and tested for the removal of As(III), which is a highly toxic, mobile, and predominant arsenic species in anoxic groundwater. We used SEM-EDX, AFM, and XRD to characterize particle size, surface morphology, and corrosion layers formed on pristine NZVI and As(III)-treated NZVI. AFM results showed that particle size ranged from 1 to 120 nm. XRD and SEM results revealed that NZVI gradually converted to magnetite/maghemite corrosion products mixed with lepidocrocite over 60 d. Arsenic(III) adsorption kinetics were rapid and occurred on a scale of minutes following a pseudo-first-order rate expression with observed reaction rate constants (kobs) of 0.07-1.3 min-1 (at varied NZVI concentration). These values are about 1000× higher than kobs literature values for As(III) adsorption on micron size ZVI. Batch experiments were performed to determine the feasibility of NZVI as an adsorbent for As(III) treatment in groundwater as affected by initial As(III) concentration and pH (pH 3-12). The maximum As(III) adsorption capacity in batch experiments calculated by Freundlich adsorption isotherm was 3.5 mg of As(III)/g of NZVI. Laser light scattering (electrophoretic mobility measurement) confirmed NZVI-As(III) inner-sphere surface complexation. The effects of competing anions showed HCO3-, H4SiO40, and H2PO42- are potential interferences in the As(III) adsorption reaction. Our results suggest that NZVI is a suitable candidate for both in-situ and ex-situ groundwater treatment due to its high reactivity.
Introduction Arsenic (As), a common constituent of the earth’s crust, is a well-known carcinogen (1). It is naturally present in water in different oxidation states and acid-base species depending on redox and pH conditions (2). Arsenic is introduced into the environment through a combination of natural processes (weathering reactions, biological activities, and volcanic * Corresponding author phone: +82-62-970-2441; fax: +82-62970-2434; e-mail:
[email protected]. † GIST. ‡ San Francisco State University. § University of Grenoble I/CNRS. 10.1021/es048991u CCC: $30.25 Published on Web 01/06/2005
2005 American Chemical Society
emissions) as well as anthropogenic activities (3). The natural occurrence of As in groundwater is of great concern due to the toxicity of As and the potential for chronic exposure (47). To address this problem, the World Health Organization (WHO) has set a maximum guideline concentration of 0.01 mg/L for As in drinking water (8). Arsenic exists in groundwater predominantly as inorganic arsenite, As(III) (H3AsO3, H2AsO31-, HAsO32-), and arsenate, As(V) (H3AsO4, H2AsO41-, HAsO42-) (2, 9). Greater attention is required for the removal of As(III) from groundwater due to its higher toxicity (10) and mobility (11, 12), which mainly arise from its neutral state (HAsO30) in groundwater as compared to the charged As(V) species (H2AsO4-, HAsO42-), which predominate near pH 6-9 (13, 14). This also correlates with the less efficient removal of As(III) by conventional water treatment processes (15). Many methods are currently in use for removing As from drinking water supplies including anion exchange, reverse osmosis, lime softening, microbial transformation, chemical precipitation, and adsorption (2, 16-18). Attention has recently focused on zero-valent iron (ZVI) for rapid As(III) and As(V) removal in the subsurface environment (9, 15, 19-22). The As(III) removal mechanism is mainly due to spontaneous adsorption and coprecipitation of As(III) with iron(II) and iron(III) oxides/hydroxides, which form in-situ during ZVI oxidation (corrosion) (9, 19-23). The oxidation of ZVI by water and oxygen produces ferrous iron (24):
Fe0 + 2H2O f 2Fe2+ + H2 + 2OH-
(1)
Fe0 + O2 + 2H2O f 2Fe2+ + 4OH-
(2)
Fe(II) further reacts to give magnetite (Fe3O4), ferrous hydroxide (Fe(OH)2), and ferric hydroxide (Fe(OH)3) depending upon redox conditions and pH:
6Fe2+ + O2 + 6H2O f 2Fe3O4(s) + 12H+
(3)
Fe2+ + 2OH-f Fe(OH)2(s)
(4)
6Fe(OH)2(s) + O2 f 2Fe3O4(s) + 6H2O
(5)
Fe3O4(s) + O2(aq) + 18H2O T 12Fe(OH)3(s)
(6)
Heterogeneous reactions at the corroding ZVI surface are complex and result in a variety of potential adsorption surfaces for As(III) and As(V). Despite this complexity, studies using X-ray absorption spectroscopy showed that the products after reaction of As(III) and As(V) with ZVI were innersphere As(III) and As(V) surface complexes on iron(III) oxides/ hydroxide corrosion products (9, 21). Recently, the versatility of nanometer-scale zero-valent iron (NZVI) material has been demonstrated for potential use in environmental engineering (25). Due to the extremely small particle size, large surface area, and high in-situ reactivity, these materials have great potential in a wide array of environmental applications such as soil, sediment, and groundwater remediation (24-26). In addition, due to small size and capacity to remain in suspension, NZVI can be transported effectively by groundwater (27) and can be injected as sub-colloidal metal particles into contaminated soils, sediments, and aquifers (26, 28). To the best of our knowledge, however, the application of NZVI to the removal of As from groundwater in developing countries such as Bangladesh and Nepal has not been reported. Arsenic VOL. 39, NO. 5, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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contamination in these countries has recently been detected not only in drinking water (29) but also in agricultural products such as rice (30). However the removal mechanism of As by NZVI has not been well understood, and very few studies have been reported on this issue (9, 27). The goal of this study is to investigate the application of laboratory-synthesized NZVI for the remediation of As(III) in groundwater. The main objectives are to (i) characterize NZVI and its reaction products using spectroscopic techniques, (ii) determine the rate and extent of As adsorption by NZVI, and (iii) investigate the effects of As(III) concentration and groundwater chemistry (pH, competing anions) on As(III) adsorption.
Experimental Section Materials and Chemicals. The chemical reagents used in the study (NaAsO2, HCl, NaOH, NaH2PO4, KI, and NaBH4) were reagent grade obtained from Aldrich Chemical Co. In some experiments, groundwater from Bangladesh was used (pH ) 6.5) with total alkalinity, dissolved organic carbon (DOC), iron (Fe2+), sulfate (SO42-), and phosphate (H2PO42-) levels of 209, 10.0, 7.1, 8.52, and 0.35 mg L-1, respectively (29). A groundwater sample from Nepal (pH ) 7.1) was also used with total alkalinity, DOC, Fe2+, SO42-, and H2PO42levels of 320, 1.0, 2.1, 8.5, and 0.2 mg L-1, respectively (31). The groundwater samples were collected in 50-mL polypropylene flasks and acidified with 1 mL of concentrated HNO3 for cation analyses on site. Replicate samples used for anion analyses were filtered with a 0.45-µm membrane and not acidified following the method reported by Berg et al. (32). All the experiments were performed in 0.01 M NaCl background solution of As(III). There was no arsenic in the groundwater obtained from Nepal whereas Bangladesh groundwater contained 20 µg/L. The concentration of Si2+ in Nepal and Bangladesh water was 41.0 and 18.1 mg/L, respectively. The NZVI material was synthesized by dropwise addition of 1.6 M NaBH4 aqueous solution to a Ne gas-purged 1 M FeCl3‚6H2O aqueous solution at ∼23 °C with magnetic stirring as described by Wang and Zhang (25). Ferric iron (Fe3+) was reduced according to the reaction (33):
Fe(H2O)63+ + 3BH4- + 3H2O ) Fe0V + 3B(OH)3 + 10.5H2 (7) The solution was stirred for 20 min and centrifuged at 6000g for 2 min, and the supernatant solution was replaced by acetone. Acetone-washing prevented the immediate rusting of NZVI during purification leading to a fine black powder product after freeze-drying. A ZVI electrolytic powder sample from Kanto Chemical Co., Inc. (Japan, 98% purity) was used in this study for comparison with NZVI. All Fe materials were stored in a N2-purged desiccator. Solid-Phase Characterization of NZVI and NZVI Corrosion Products. The NZVI material was characterized by powder X-ray diffraction (XRD) using a Rigaku diffractrometer and monochromatized Cu KR radiation (generator tension ) 40 kV, current ) 40 mA). Diffractograms were recorded from 5 to 85° (2θ) with a step size of 0.02° and a count time of 5 s per step. Morphological analysis of the samples was performed by field emission scanning electron microscopy (FE-SEM) using a Hitachi 4700 microscope (at 15 kV) with energy-dispersive X-ray (EDX) analyses. Powder samples were prepared by mounting on carbon tape followed by platinum coating. The morphology and size of NZVI material was analyzed by atomic force microscopy (AFM) using a Digital Instrument-Nanoscope IIIA (Santa Barbara, CA) in contact mode. A silicon nitride tip with 20 nm radius of curvature was used (spring constant (k) of 0.58 N/m). The sample was 1292
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prepared by spin coating the aqueous NZVI dispersion on a silicon wafer. The specific surface area (SBET) of NZVI was measured by Brunauer-Emmett-Teller (BET) N2 method, and pore size was calculated by the BJH (Barrett, Joyner, and Halenda) N2 adsorption/desorption isotherm method at 77 K using the automatic analyzer (ASAP 2020, Micromeritics, USA). Batch Equilibrium As(III) Adsorption/Desorption Methods. Adsorption of As(III) on NZVI. Stock solutions of 1000 mg L-1 As(III) and As(V) were prepared by dissolving NaAsO2 and Na2HAsO4‚7H2O, respectively in deionized (DI) water. Batch adsorption of As(III) was studied in 40-mL polypropylene copolymer centrifuge tubes containing 20 mg of NZVI in 20 mL of As(III) reaction solution in 0.01 M NaCl. Adsorption isotherms were produced by varying the As(III) starting concentration, and the adsorption envelopes were generated with a fixed As(III) concentration (1 mg L-1) and varying pH. The solution pH was adjusted using 1 M NaOH or HCl to the desired pH value, measured by Orion (model 250A) pH meter. The pH of the solution was controlled from the beginning of the reaction only if needed and monitored during the experiments. Reactions proceeded for 12 h at 25° C (unless otherwise specified) in a shaken water bath (185 rpm) kept in the dark by covering with aluminum foil. The 12 h reaction time was shown to be adequate by preliminary experiments (Figure 5, inset) for equilibrium to be attained. The concentration of dissolved oxygen was not measured during our experiments; however, a NZVI-free control experiment with identical As(III) concentration, pH, and reaction conditions was performed, and As(III) was observed not to be oxidized to As(V). After 12 h, the supernatant solution was filtered through a 0.45-µm membrane filter (Millipore), with a disposable syringe, and analyzed by Hydride Generation atomic absorption spectrophotometry (HGAAS; Perkin-Elmer 5100 PC) with a detection limit of 1 µg L-1 As. As(III) was the initial form of As in all experiments (unless otherwise specified), and total As (AsT) was measured by HGAAS after reaction with NZVI. As(III) and AsT concentrations were also measured in certain experiments using an anion-exchange cartridge method for speciation of As(III) and As(V). The cation and anion-exchange cartridge retained organically complexed As and As(V) whereas As(III) was unretained. Retained As(V) was eluted with 1 M HCl to separate it from organically complexed As (34). All experiments were performed in triplicate, and solutions analyzed for As by HGAAS were prepared fresh at the time of analysis. Competing Anions. Batch tests were performed with competitive anions using solutions of 1 mg L-1 As(III) in 1, 10, and 100 mM solutions of NaH2PO4, Na2SiO3, Na2SO4, NaNO3, or NaHCO3. After a 12-h reaction time, the suspension was centrifuged, filtered through a 0.45-µm membrane filter (Millipore), and analyzed for AsT as described above (20). Desorption of As from NZVI Corrosion Product. A range of As(III) (1-10 mg L-1) was reacted with NZVI (1 g L-1) in 0.01 M NaCl. After 12 h, the As(III)-treated NZVI was separated and gently washed with water to remove aqueous As. The As(III)-treated NZVI was agitated for 18 h with 20 mL 0.10 M NaH2PO4 similar to previous studies (35, 9). Aliquots of supernatant solution were filtered (0.45 µm), eluted through cartridges for As(III)/(V) speciation, and analyzed for AsT by HGAAS. All samples were analyzed for AsT immediately without storing before analysis. Kinetic Investigation of As(III) Adsorption by NZVI. The short-term (0-1 h) time dependence of the As(III) adsorption reaction with NZVI was analyzed by reacting 1 mg L-1 As(III) with varying NZVI solid concentrations (0.5, 1, 2.5, 5, 7.5, and 10 g L-1) with AsT analyzed by HGAAS at 0, 0.5, 1, 3, 5, 7, 10, 30, and 60 min. The long-term As sorption capacity was also evaluated for NZVI. Triplicate samples of 1.0 g of
FIGURE 1. SEM image of pristine NZVI (a) and As(III) sorbed on NZVI for 7 (b), 30 (c), and 60 d (d), respectively. Reaction conditions: 100 mg L-1 As(III) adsorbed on 50 g L-1 NZVI in 0.01 M NaCl at pH 7, 25 °C. NZVI were reacted in 40-mL polypropylene bottles containing 20 mL of 100 mg L-1 As(III) in 0.01 M NaCl in the dark with analyses performed at 1, 7, 30, and 60 d to test the sorption capacity of NZVI over time. After completion of the given reaction time, the tubes were centrifuged, and supernatant solutions were filtered and analyzed by HGAAS. Selected solids from these experiments were also investigated by SEM, AFM, and XRD. Electrophoretic Mobility. Electrophoretic mobility (ζpotential) of NZVI in aqueous solution was measured by light scattering (ELS-8000, Photal, Otsuka Electronics, Japan). The reaction conditions were 1 g L-1 NZVI in solutions containing 0.01 M NaCl and either 0.1 or 1.0 mg L-1 As(III). The pH was adjusted between 5 and 9 by dropwise addition of 1 M HCl or NaOH, and the mixtures were shaken for 12 h prior to analysis.
Results and Discussion Characterization of NZVI and NZVI-As(III) Products. SEM Images and EDX Spectra. Solid samples collected from pristine NZVI and 100 mg/L As(III)-treated NZVI after 7, 30, and 60 d and imaged by SEM are shown in Figure 1a-d. Synthetic NZVI particles were in the size range of 10-100 nm as measured by SEM and had a pore size of ∼20 nm. Adsorption of As and time causes increases in particle aggregation due to iron(III) oxide/hydroxide precipitation. SEM pictures clearly show a growth of a fine needle-like crystallite, which transform to an apparent amorphous phase. The thin crystallites (about 100 nm long by 20 nm wide) are energetically unstable and disappear to be replaced by more stable phases according to the Gay-Lussac-Oswald ripening rule. Quantitative SEM-EDX peak area analysis shows As(III) adsorbed on NZVI up to 1.11, 1.42, and 2.27% (w:w) over 7, 30, and 60 d, respectively (Figure 2). The EDX spectra were collected on samples treated with a higher As(III) concentration (100 mg L-1) than the batch equilibrium and kinetic work. The EDX peak area analysis provides an approximate surface corrosion product percent Fe:O:As elemental composition of 65:33:2 at 60 d. The Fe percent of 65 compares
FIGURE 2. SEM-energy dispersive X-ray analysis of pristine NZVI (a) and 100 mg L-1 As(III) adsorbed on 50 g L-1 NZVI in 0.01 M NaCl for 7 (b), 30 (c), and 60 d (d), respectively. with FeOOH ) 53% Fe and an Fe2O3 ) 70% Fe (w:w). The high As content of 2.27% (w:w) measured using the data shown in Figure 2 (inset) suggests an enriched surface layer VOL. 39, NO. 5, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 3. Contact mode AFM analysis of pristine NZVI: 3-D image (a), cross-sectional diagram (b), and tip inflection (c).
FIGURE 4. X-ray diffraction analysis of NZVI, commercial ZVI (Kanto Chemical Co.) (a), pristine NZVI (b), and 100 mg L-1 As (III) sorbed on 50 g L-1 NZVI in 0.01 M NaCl for 1 (c), 7 (d), 30 (e), and 60 d (f), respectively. Peaks are due to magnetite/maghemite (M) (Fe3O4/ γ-Fe2O3), lepidocrocite (γ-FeOOH) (L), and NZVI (Fe0), respectively. of adsorbed As with an As elemental composition 10× in excess of the bulk. AFM Analysis. Figure 3 shows AFM images of pristine NZVI in 3-D and a cross-sectional analysis detected by tip inflection. AFM cross-section analysis shows that the size distribution of NZVI particles are in the range of 1-120 nm, in which more than 60% are less than 50 nm. Surfaces of the nanometer-scale domains were of two types: (i) a small size of less than 8 nm and (ii) a larger size of about 200 nm that includes aggregates. X-ray Diffractograms. The XRD analysis of NZVI, commercial ZVI (Kanto Chemical Co.) and As(III)-treated NZVI samples (1, 7, 30, and 60 d) is shown in Figure 4. The zero valence state and crystalline structure of NZVI were confirmed by X-ray diffraction analysis when comparing with Kanto Chemical Co. ZVI material (Figure 4). X-ray diffractograms 1294
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demonstrate that the NZVI corrosion products are a mixture of amorphous iron(III) oxide/hydroxide, magnetite (Fe3O4), and/or maghemite (γ-Fe2O3), and lepidocrocite (γ-FeOOH). These Fe(II)/(III) and Fe(III) corrosion products indicate that Fe(II) formation is an intermediate step in the NZVI corrosion process. In the 24-h reaction product, an amorphous domain is seen among magnetite, lepidocrocite peaks, and a predominant ZVI (Fe0) peak. Amorphous products were replaced by magnetite and lepidocrocite over a 2-month period (Figure 4). After 60 d, the As(III)-NZVI corrosion product had predominantly magnetite and lepidocrocite crystalline composition. Similar results were reported by Manning et al. (9) in corrosion products from ZVI powder and by Richmond et al. (36) in As removal via ferrihydrite crystallization control. Surface Area. Unreacted NZVI had a specific surface area of 24.4 m2/g as measured by BET surface area analyzer. The BET surface area measurement was also performed on corroded NZVI and found to be 37.2 m2/g. Though modest when compared to high surface area sorbents, in-situ formation of iron(III) oxide/hydroxide coatings on NZVI and the resulting As adsorption and occlusion within the precipitate is a reactive surface that may not be accounted for by the BET surface area measurement. Arsenic(III) Adsorption on NZVI Corrosion Products. Kinetics of As(III) Adsorption. The influence of NZVI concentrations (0.5, 2.5, 5, 7.5, and 10 g L-1) on the rate of adsorption of the As(III) was investigated using 1 mg L-1 of the As(III) at pH 7 (Figure 5 and inset). For all treatments (except the 0.5 g/L treatment), greater than 80% total As (AsT) was adsorbed within 7 min and ∼99.9% within 60 min. The As(III) adsorption kinetics data were examined using a pseudo-first-order reaction kinetics expression:
rate ) -
d[AsT] ) kobs[NZVI] dt
(8)
where AsT ) As(III) + As(V) and is the concentration of As (mg L-1) at time t (min), [NZVI] is the concentration of NZVI
TABLE 1. Pseudo-First-Order Rate Constants (kobs) and Their Surface Area Normalized Rate Constants (ksa) for Arsenic Removal by NZVIa
a
NZVI (g/L)
kobs (min-1)
ksa (L m-2 min-1)
R2
0.5 1.0 2.5 5.0 7.5 10.0
0.07 0.28 0.32 0.68 0.76 1.30
0.0057 0.0115 0.0052 0.0056 0.0042 0.0053
0.85 0.97 0.95 0.95 0.77 0.77
As(III): 1 mg/L in 0.01 M NaCl, at pH 7, 25 °C.
TABLE 2. Related Parameters for the Adsorption of AsT on NZVI at Different Temperaturesa Freundlich constants
FIGURE 5. Kinetics of As(III) adsorption. The reaction is pseudofirst-order with respect to the total NZVI concentrations. Reaction conditions: 1 mg L-1 As(III) adsorbed on 0.5-10.0 g L-1 NZVI concentration in 0.01 M NaCl at pH 7, 25 °C. Upper inset shows As(III) adsorption with respect to time, initial As(III): 1 mg/L in 0.01 M NaCl, pH 7, at 25 °C. (g L-1), and kobs is the pseudo-first-order rate constant of As (min-1). Oxidation of As(III) to As(V) during the in-situ reaction with NZVI required the use of AsT to describe the overall kinetics of As(III) adsorption. For 20 mL of 1 mg/L initial As concentration and 0.5-10.0 g/L NZVI, the values of initial rate constants were 0.07-1.3 min-1 (Table 1). An initial faster rate of As(III) disappearance from aqueous solution (∼80% adsorption of AsT) took place within 7 min followed by a slower uptake reaction. This sequence of reactivity was consistent with the results of pseudo-firstorder rate of As(V) adsorption by ZVI (21, 37) and chromium(VI) and lead(II) adsorption by NZVI (24, 38). In our experiment, the surface normalized rate constant (ksa) for As(III) was 342, 690, 312, 336, 252, and 318 mL m-2 h-1 at 0.5, 1.0, 2.5, 5.0, 7.5, and 10.0 g/L NZVI, respectively. Su and Puls (19, 20) reported adsorption of As(III) by micron size ZVI and found for 24 g L-1 ZVI (Fisher, Peerless, Master Builders and Aldrich), ksa for As(III) was 68.2, 0.59, 0.44, and 1.56 mL m-2 h-1 for each Fisher, Peerless, Master Builders, and Aldrich ZVI, respectively. Hence, the ksa for NZVI was 1-3 orders of magnitude higher than micron size ZVI. Arsenic(III) Adsorption Isotherms. The equilibrium adsorption isotherm data at varying temperature were analyzed using Freundlich (eq 9) and Langmuir (eq 10) adsorption expressions:
q ) KFC1/n q)
qmKLC 1 + K LC
(9) (10)
where qm (mg g-1) is the maximum adsorption capacity, q (mg g-1) is the amount of adsorbed As, C (mg L-1) is the equilibrium solute (As) concentration, KF and n are the Freundlich constants, and KL is the Langmuir constant. The Freundlich and Langmuir parameters were obtained by nonlinear least-squares regression analysis (39). The results of fitting Freundlich and Langmuir equations to isotherm curves are summarized in Table 2. The As adsorption on NZVI follows Freundlich as well as Langmuir isotherms; however, only the Freundlich curves are shown in Figure 6 based on the lower square of residual (R) value. The maximum adsorption of As on NZVI calculated by Freundlich isotherm was 3.5 mg of As(III)/g of NZVI at 25 °C. Thermodynamic
Langmuir constants
temp (°C)
KFb
nc
R
KL d
qmaxe
R
25 35 45
3.50 5.06 2.56
3.27 2.02 1.71
0.183 0.036 0.037
135.00 2.04 7.94
1.80 2.47 1.56
0.197 0.041 0.049
a Experimental conditions for NZVI: 1 g/L in 0.01 M NaCl, pH 7. b K F ) Freundlich constant related to adsorption capacity. c n ) Freundlich adsorption intensity. d KL ) Langmuir constant related to the Langmuir maximum adsorption capacity. e qmax ) Langmuir constant related to energy of adsorption where R (square of residuals) was calculated as R ) 1/2∑i)1(Ccal - Cexp)2. Ccal ) calculated concentration; Cexpt ) experimental value of concentrations.
FIGURE 6. Adsorption isotherm plots for the adsorption of As(III) by NZVI at 25, 35, and 45 °C. Reaction conditions for NZVI: 1.0 g L-1 in 0.01 M NaCl, at pH 7, reaction time 12 h. parameters were also calculated based on the eqs 11-13 and reported in Table 3 (40):
∆G ) -RT ln K
(11)
∆H +C RT
(12)
ln K )
∆G ) ∆H - T∆S
(13)
where ∆G is the Gibbs free energy change, R is the ideal gas constant (4.187 J mol-1 K-1), T is temperature (K), K is the Langmuir or Freundlich isotherm constant, ∆H is the enthalpy change, and ∆S is the entropy change. The calculated values of n (eq 9) lie between 1 and 10, indicating a spontaneous adsorption reaction (41). In addition, the negative free energy value and the measured decrease in As(III) adsorption with increasing temperature are evidence of an exothermic adsorption reaction (42) (Figure 6). Effect of pH on As(III) Adsorption. The effect of pH on As(III) adsorption on NZVI is presented in Figure 7. The extent of removal was 88.6-99.9% in the pH range 4-10 and VOL. 39, NO. 5, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 3. Thermodynamic Parameters for Adsorption of AsT on NZVI at Different Temperatures (Langmuir and Freundlich Isotherm)a temp (°C)
ln KF
25 35 45
1.25 1.62 0.94
temp (°C)
ln KL
25 35 45
4.91 3.02 2.07
Freundlich Constants -∆Gb ∆ Sc (kJ/mol) (KJ K-1 mol-1) -1.57 -2.09 -1.25
-0.01 -0.01 -0.01
Langmuir Constants -∆Gb ∆ Sc (kJ/mol) (KJ K-1 mol-1) -6.1 -3.8 -2.7
∆Hd (kJ/mol)
-0.17 -0.17 -0.17
a Initial As(III): 1 mg/L; NZVI: 1 g/L in 0.01 M NaCl, pH 7. energy. c ∆S ) entropy of activation. d ∆H ) enthalpy.
-5.9
∆Hd (kJ/mol) -56.41
b
∆G ) free
FIGURE 8. Electrophoretic mobility of NZVI in the presence and absence of As(III) with respect to pH (pH 5-9). Reaction conditions: 1.0 and 0.1 mg L-1 As(III) NZVI adsorbed on 1.0 g L-1 in 0.01M NaCl, 25 °C, reaction time 12 h. Ambient pH ) 6.5 and was adjusted with 1 M HCl or NaOH.
FIGURE 7. Adsorption of As(III) on NZVI as a function of pH (envelope plot). Reaction conditions: 1.0 mg L-1 As(III) adsorbed on 1.0 g L-1 NZVI in 0.01 M NaCl, reaction time 12 h. Ambient pH ) 6.5 and was adjusted with 1 M HCl or NaOH. decreased sharply at pH below 4 and above 10. The pHdependent behavior can be explained by ionization of both the adsorbate and the adsorbent causing repulsion at the surface and decreasing the net As(III) adsorption. Below pH 9.2, H3AsO3 is the predominant species and presumably the major species being adsorbed. When the pH is above 9.2, H2AsO3- is the predominant As(III) species while the NZVI corrosion product surfaces are also negative (Fe(III)-O-) causing electrostatic repulsion. A similar pH dependence trend in As(III) adsorption amorphous iron oxide, synthetic goethite, and magnetite has been observed (11, 43). The electrophoretic mobility of NZVI at different As(III) concentration and pH was measured to determine the point of zero charge (pHpzc) at which the net surface charge is zero. Untreated NZVI surfaces exhibit a net positive charge at pH lower than the pHpzc of 7.8 (Figure 8). The effect of increased As(III) adsorption causes a negative shift in the pHpzc of NZVI to 7.6 and 7.0 for 0.1 and 1.0 mg L-1 As(III), respectively. During the As(III) adsorption reaction, NZVI oxidizes to aqueous Fe2+ (2.4-6.7 mg/L) and to various corrosion products such as amorphous iron(II)/(III) oxide/hydroxide and discrete mineral phases such as magnetite and lepidocrocite. The As(III) adsorption reaction forms inner-sphere surface complexes on ZVI corrosion products (9, 21) ,which is consistent with the electrophoretic mobility results in Figure 8. The shifts in pHpzc to lower values are characteristics 1296
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FIGURE 9. Desorption by of As(III) and As(V) after 12 h sorption step on NZVI. Reaction conditions: 1.0 mg L-1 As(III) adsorbed on 1.0 g L-1 NZVI in 0.01 M NaCl, 25 °C, As(III) reaction time 12 h. For the desorption step, NaH2PO4 ) 0.1 M, pH 7, at 25 °C, reaction time 18 h. of inner-sphere complexation where H3AsO3 and H2AsO4species form complexes directly with the coordination environment of octahedral Fe(III) at surfaces and block Fe(III)-OH proton exchange with the solution. Desorption and Speciation of Arsenic(III/V). Experiments were performed to determine the speciation of As(III)/(V) after adsorption of As(III) on NZVI. The desorption study showed oxidation of As(III) to As(V) on NZVI corrosion product (Figure 9). The highest As(V) recovery was 20% (whereas As(III) desorption was 11%) in the 10 mg L-1 As(III) treatment. These data confirm that oxidation of As(III) to As(V) is involved in the NZVI removal mechanism. Recent research suggests that the formation of Fe2+ and H2O2 on the corroding Fe0 surface in turn forms OH• radical (44, 45):
Fe0 + O2 + 2H+ f Fe2+ + H2O2
(14)
Fe2+ + H2O2 f FeIII OH2+ + OH•
(15)
The As(III) oxidation reaction then proceeds as:
2OH• + H3AsO3 f H2AsO4- + H2O + H+
(16)
TABLE 4. Percentage of As(III) Removal in the Presence of Competitive Anionsa concn of % uptake of As(III) in presence of anionb anions c (mM/µM) NO3 SO42- c HCO3- c H4SiO40 c PO43- c AsO43- d 0.00 1.00 10.00 100.00
99.90 99.34 99.29 98.46
99.90 99.75 98.59 98.1
99.90 99.80 99.66 82.43
99.90 70.65 44.96 0.00
99.90 96.60 66.30 0.00
99.90 99.90 99.90 99.90
a Initial As(III): 1 mg/L; NZVI: 1 g/L in 0.01 M NaCl, pH 7. at 25 °C. Note: % uptake of As(III) in the presence of anions: n ) 3, average of triplicate results where the standard deviation is less than 5%. c Anions in mM. d Anion in µM. b
FIGURE 10. As(III) removal from groundwater using NZVI and FeCl3 from groundwater of Nepal and Bangladesh. As (III): 1 mg/L in 0.01 M NaCl; NZVI: 1 g/L, pH 7, 25 °C. Our results suggest that a mixture of As(III) and As(V) was present in our experiments. Oxidation of As(III) by Fe(II) and oxygen present in a batch reactor has also been shown (46). Effect of Competing Anions. Table 4 shows the effect of individual anions (HCO3-, SO42-, NO3-, H2AsO4-, H4SiO40, and H2PO42-) on the adsorption of As(III) on NZVI. The HCO3-, SO42-, and NO3- ions at concentrations up to 10 mM have no effect on As(III) uptake. The presence of 10 mM H4SiO40 and H2PO42- reduced the uptake of As(III) from 99.9% to 44.94 and 66.3%, respectively. Competing anion concentrations up to 100 mM cause decreases in As(III) adsorption on NZVI from 99.9% to 98.46, 98.1, and 82.43% for NO3-, SO42-, and HCO3- ions, respectively, whereas 100 mM H4SiO4° and H2PO42- reduce the adsorption of As(III) to zero due to the competitive reaction with As(III) (33, 47). Similar results were reported by Su and Puls (20); however, others have shown that H2PO42- adsorbs more strongly than H4SiO40 on synthetic iron oxides and oxyhydroxides (48). These results are informative for in-situ remediation using ZVI and NZVI due to the coexistence of competing anions in natural groundwater. Removal of As(III) from Groundwater. Batch studies on the removal of As(III) were carried out on the groundwater obtained from Bangladesh and Nepal (Figure 10). The As(III) was added to Bangladesh and Nepal groundwater, and no As(III) removal without NZVI addition was observed. Arsenic(III) was removed with 100% efficiency with 1.0, 2.5, and 4.5 g/L NZVI for As spiked in DI water, in the Bangladesh groundwater, and in the Nepal groundwater, respectively. The Bangladesh groundwater was also treated with FeCl3 of different concentrations in the presence of 10 mg L-1 NZVI as a further test treatment as shown in Figure 10, and it is found that a very low concentration of NZVI is required when
As is removed by FeCl3 in the presence of NZVI. A greater amount of NZVI is required for complete removal of As(III) from field-collected groundwater due to the presence of anions such as HCO3- and SO42- and possibly trace amounts of silica and H2PO4-. Mechanism of As(III) Removal. Evidence has been presented showing that As(III) can be removed by adsorption on NZVI in a very short time (minute scale) and is strongly adsorbed on NZVI over a wide range of pH and anion environments. The identity of the actual reactive surface site on NZVI is likely either a stable or metastable iron(II), mixed iron(II)/(III), or iron(III) oxide, hydroxide, or oxyhydroxide corrosion product. The results of time-resolved XRD (Figure 4) suggest that initially (0-24 h) amorphous Fe(II)/(III) and magnetite (or maghemite) are sites for adsorption. As NZVI corrodes over longer periods, more crystalline magnetite and lepidocrocite products are present for ongoing adsorption of As(III). Therefore the suite of available reactive sites for As(III) adsorption changes with time. In addition, As(III) near or in contact with the corroding NZVI surface is oxidized to As(V), which in turn is adsorbed by an inner-sphere mechanism similar to As(III) on iron(III) oxides (9, 49, 50). Layers of As(III) and As(V) adsorbed on NZVI corrosion product films are buried by successive layers and become occluded from the surrounding solution. The results of this study show that NZVI is both an efficient material for the treatment of As(III) and may be used as a new material for permeable reactive barrier walls as well as a material for ex-situ treatment. Implication to In Situ Remediation. NZVI has applicability in ex-situ as well as in-situ remediation of pollutants including arsenic. NZVI can be used to remediate pollutants already present in soil and groundwater. Zhang (27) reported an in-situ application of NZVI and found that 11.2 kg of NZVI could remediate about a 100 m2 area using a single injection well to treat chlorinated organic pollutants (TCE, PCE, and DCE) in about 4-6 weeks. In addition, NZVI promotes anaerobic microbial growth in the subsurface by increasing pH, decreasing redox potential, producing hydrogen gas, and releasing ferrous iron ion (27). We have presented evidence that As(III) can be removed by adsorption on NZVI on a minute time scale. As(III) strongly sorbs on NZVI in a wide range of pH, and various As(III) and As(V) coprecipitates on iron(III) oxide/hydroxide corrosion products are involved. Engineering studies to develop this NZVI technology are currently going on in our laboratory. The results of this study show that NZVI is an efficient material for the treatment of As(III) and may be used in a permeable reactive barrier as well as for ex-situ groundwater treatment.
Acknowledgments This work was supported by a grant (B10) from Suitainable Water Resources Research Center (SWRRC) of 21st Century Frontier R&D Program through the Water Reuse Technology Center (WRTC) at Gwangju Institute of Science and Technology (GIST).
Literature Cited (1) Agency for Toxic Substances and Disease Registry. Top 20 Hazardous Substances: ATSDR/EPA Priority List for 1997. http://atsdr1.atsdr.cdc.gov: 8080/cxcx3.html. (2) Ferguson, J. F.; Garvis, J. Review of the arsenic cycle in natural waters. Water Res. 1972, 6, 1259-1274. (3) Smedley, P. L.; Kinniburgh, D. G. A review of the source, behaviour and distribution of arsenic in natural waters. Appl. Geochem. 2002, 17, 517-568. (4) Anderson, L. C. D.; Bruland, K. W. Biogeochemistry of arsenic in natural waters: the importance of methylated species. Environ. Sci. Technol. 1991, 25, 420-427. VOL. 39, NO. 5, 2005 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
1297
(5) Tseng, W. P.; Chu, H. M.; How, S. W.; Fong, J. M.; Lin, C. S.; Yeh, S. J. Prevalence of skin cancer in an endemic area of chronic arsenism in Taiwan. J. Natl. Cancer Inst. 1968, 40, 453-463. (6) Smith, A. H.; Hopenhayn-Rich, C.; Bates, M. N.; Goeden, H. M.; Hertz-Picciotto, I.; Duggan, H. M.; Wood, R.; Kosnett, M. J.; Smith, M. T. Cancer risks from arsenic in drinking water. Environ. Health Perspect. 1992, 97, 259-267. (7) Bajpai, S.; Chaudhary, M. Removal of arsenic from groundwater by manganese dioxide-coated sand. J. Environ. Eng. 1999, 125, 782-784. (8) WHO. Guidelines for Drinking Water Quality, Vol 1: Recommendations, 2nd ed.; WHO: Geneva, 1993. (9) Manning, B. A.; Hunt, M.; Amrhein, C.; Yarmoff, J. A. Arsenic(III) and arsenic(V) reactions with zerovalent iron corrosion products. Environ. Sci. Technol. 2002, 36, 5455-5461. (10) Cullen, W. R.; Reimer, K. J. Arsenic speciation in the environment. Chem. Rev. 1989, 89, 713-764. (11) Dixit, S.; Hering, J. G. Comparison of arsenic(V) and arsenic(III) sorption onto iron oxide minerals: implications for arsenic mobility. Environ. Sci. Technol. 2003, 37, 4182-4189. (12) Manning, B. A.; Martens, D. A. Speciation of arsenic(III) and arsenic(V) in sediment extracts by high-performance liquid chromatography-hydride generation atomic absorption spectrophotometry. Environ. Sci. Technol. 1997, 31, 171-177. (13) Gulens, J.; Champ, D. R.; Jackson, R. E. In Chemistry of Water Supply Treatment and Distribution; Rubia, A. J., Ed.; Ann Arbor Science Publishers: Ann Arbor, MI, 1973. (14) Clifford, D. A.; Zang, Z. Arsenic chemistry and speciation. Proceedings of the 1993 Water Quality Technology Conference, AWWA, Denver, CO, April 1994. (15) Chiu, V. Q.; Hering, J. G. Arsenic adsorption and oxidation at manganite surfaces. 1. Method for simultaneous determination of adsorbed and dissolved arsenic species. Environ. Sci. Technol. 2000, 34, 2029-2034. (16) Research Plan for Arsenic in Drinking Water; U.S. Environmental Protection Agency, Office of Research and Development: Washington, DC, 1988; EPA/68/R-98/042. (17) Bissen, M., Frimmel, F. Arsenicsa review. Part II: Oxidation of arsenic and its removal in water treatment. Acta Hydrochim. Hydrobiol. 2003, 31, 97-107. (18) Nriagu, J. O. Arsenic in the Environment, Part I: Cycling and Characterization; John Wiley and Sons Inc.: New York 1994. (19) Su, C.; Puls, R. W. Arsenate and arsenite removal by zerovalent iron: kinetics, redox transformation, and implications for in situ groundwater remediation. Environ. Sci. Technol. 2001, 35, 1487-1492. (20) Su, C.; Puls, R. W. Arsenate and arsenite removal by zerovalent iron: effects of phosphate, silicate, carbonate, borate, sulfate, chromate, molybdate, and nitrate, relative to chloride. Environ. Sci. Technol. 2001, 35, 4562-4568. (21) Farrell, J.; Wang, J.; O’Day, P.; Coklin, M. Electrochemical and spectroscopic study of arsenate removal from water using zerovalent iron media. Environ. Sci. Technol. 2001, 35, 2026-2032. (22) Lackovic, J. A., Nikolaids, N. P., Dobbs, G. M. Inorganic arsenic removal by zero-valent iron. Environ. Eng. Sci. 2000, 17, 29-39. (23) Charlet L.; Manceau, A. Structure, formation and reactivity of hydrous oxide particles; insights from X-ray absorption spectroscopy. In Environmental Particles II; Buffle, J., Van Leeuwen, H. P., Eds.; IUPAC Environmental Analytical and Physical Chemistry Series; Lewis Publishers: Chelsea, MI, 1993. (24) Ponder, S. M.; Darab, J. C.; Mallouk, T. E. Remediation of Cr(VI) and Pb(II) aqueous solutions using supported, nanoscale zerovalent iron. Environ. Sci. Technol. 2000, 34, 2564-2569. (25) Wang, C. B.; Zhang, W. Synthesizing nanoscale iron particles for rapid and complete dechlorination of TCE and PCBs. Environ. Sci. Technol. 1997, 31, 2154-2156. (26) Lien, H.-L.; Zhang, W. Transformation of chlorinated methanes by nanoscale iron particles. J. Environ. Eng. 1999, 125, 10421047. (27) Zhang, W. X. Nano scale iron particles for environmental remediation: an overview. J. Nanopart. Res. 2003, 5, 323-332. (28) Cantrell, K. J.; Kaplan, D. I. Zero-valent iron colloid emplacement in sand columns. J. Environ. Eng. 1997, 123, 499-505. (29) Mukherjee, A. B.; Bhattacharya, P. Arsenic in groundwater in the Bengal Delta Plain: Slow poisoning in Bangladesh. Environ. Rev. 2001, 9, 189-220. (30) Meharg, A. A.; Rahman, M. M. Arsenic contamination of Bangladesh paddy field soils: implications for rice contribution
1298
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 39, NO. 5, 2005
(31)
(32)
(33)
(34)
(35)
(36)
(37)
(38)
(39)
(40)
(41) (42) (43)
(44)
(45)
(46)
(47)
(48)
(49)
(50)
to arsenic consumption. Environ. Sci. Technol. 2003, 37, 229234. Kanel, S. R.; Choi, H.; Kim, K. Analysis and Monitoring of Arsenic Contamination in Groundwater in Rautahat District, Nepal; UNU/ISTS Report; K.-JIST: Korea 2003; pp 1-32. Berg, M.; Tran. H. C.; Nguyen, T. C.; Pham, H. V.; Schertenleib, R.; Giger, W. Arsenic contamination of groundwater and drinking water in Vietnam: a human health threat. Environ. Sci. Technol. 2001, 35, 2621-2626. Glavee, G. N.; Klabunde, K. J.; Sorensen, C. M.; Hadjipanayis, G. C. Chemistry of borohydride reduction of iron(II) and iron(III) ions in aqueous and nonaqueous media. formation of nanoscale Fe, FeB, and Fe2B powders. Inorg. Chem. 1995, 34, 28-35. Ko, I.; Kim, J.; Kim, K. W. Arsenic speciation and sorption kinetics in the As-hematite-humic acid system. Colloids Surf. A 2004, 234, 43-50. Namasivayam, C.; Senthilkumar, S. Removal of arsenic(V) from aqueous solution using industrial solid waste: adsorption rates and equilibrium studies. Ind. Eng. Chem. Res. 1998, 37, 48164822. Richmond, W. R.; Loan, M.; Morton, J.; Parkinson, G. M. Arsenic removal from aqueous solution via ferrihydrite crystallization control. Environ. Sci. Technol. 2004, 38, 2368-2372. Melitas, N.; Wang, J.; Conklin, M.; O’Day, P.; Farrell, J. Understanding soluble arsenate removal kinetics by zerovalent iron media. Environ. Sci. Technol. 2002, 36, 2074-2081. Ponder, S. M.; Darab, J. G.; Bucher, J.; Caulder, D.; Craig, I.; Davis, L.; Edelstein, N.; Lukens, W.; Nitsche, H.; Rao, L. Shuh, D. K.; Mallouk, T. E. Surface chemistry and electrochemistry of supported zerovalent iron nanoparticles in the remediation of aqueous metal contaminants. Chem. Mater. 2001, 13, 479-486. Reddad, Z.; Gerente, C.; Andres, Y.; Cloirec, P. L. Adsorption of several metal ions onto a low-cost biosorbent: kinetic and equilibrium studies. Environ. Sci. Technol. 2002, 36, 2067-2073. Ajmal, M.; Khan, A. H.; Ahmad, S.; Ahmad, A. Role of sawdust in the removal of copper(II) from industrial wastes. Water Res. 1998, 32, 3085-3091. McKay, G.; Blair, S. H.; Garden, J. R. Adsorption of dyes on chitin. J. Appl. Polym. Sci. 1982, 27, 3043-3057. Atkins, P. W. Physical Chemistry, 5th ed.; W. H. Freeman and Company: New York, 1994. Manning, B. A.; Fendorf, S. E.; Goldberg, S. Surface structures and stability of arsenic(III) on goethite: spectroscopic evidence for inner-sphere complexes. Environ. Sci. Technol. 1998, 32, 2383-2388. Voegelin, A.; Hug, S. J. Catalyzed oxidation of arsenic(III) by hydrogen peroxide on the surface of ferrihydrite: an in situ ATR-FTIR study. Environ. Sci. Technol. 2003, 37, 972-978. Joo, S. H.; Feitz, A. J.; Waite, T. D. Oxidative degradation of the carbothioate herbicide, molinate, using nanoscale zero-valent iron. Environ. Sci. Technol. 2004, 38, 2242-2247. Hug, S. J.; Leupin, O. Iron-catalyzed oxidation of arsenic(III) by oxygen and by hydrogen peroxide: pH-dependent formation of oxidants in the fenton reaction. Environ. Sci. Technol. 2003, 37, 2734-2742. Appelo, C. A. J.; Weiden, M. J. J. V. D.; Tournasat, C.; Charlet, L. Surface Complexation of ferrous iron and carbonate on ferrihydrite and the mobilization of arsenic. Environ. Sci. Technol. 2002, 36, 3096-3103. Meng, X. G.; Korfiatis, G. P.; Bang, S. B.; Bang, K. W. Combined effects of anions on arsenic removal by iron hydroxides. Toxicol. Lett. 2002, 133, 103-111. Waychunas, G. A.; Rea, B. A.; Fuller, C. C.; Davis, J. A. Surface chemistry of ferrihydrite: Part 1. EXAFS studies of the geometry of coprecipitated and adsorbed arsenate. Geochim. Cosmochim. Acta 1993, 57, 2251-2269. Fendorf, S. E.; Eick, M. J.; Grossl, P.; Sparks, D. L. Arsenate and chromate retention mechanisms on goethite. 1. surface structure. Environ. Sci. Technol. 1997, 31, 315-320.
Received for review July 1, 2004. Revised manuscript received October 2, 2004. Accepted October 20, 2004. ES048991U
