Environ. Sci. Technol. 2003, 37, 4500-4506
Removal of Arsenic from Synthetic Acid Mine Drainage by Electrochemical pH Adjustment and Coprecipitation with Iron Hydroxide JENNY WEIJUN WANG, DORIN BEJAN, AND NIGEL J. BUNCE* Department of Chemistry and Biochemistry, University of Guelph, Guelph, Ontario, Canada, N1G 2W1
Acid mine drainage (AMD), which is caused by the biological oxidation of sulfidic materials, frequently contains arsenic in the form of arsenite, As(III), and/or arsenate, As(V), along with much higher concentrations of dissolved iron. The present work is directed toward the removal of arsenic from synthetic AMD by raising the pH of the solution by electrochemical reduction of H+ to elemental hydrogen and coprecipitation of arsenic with iron(III) hydroxide, following aeration of the catholyte. Electrolysis was carried out at constant current using two-compartment cells separated with a cation exchange membrane. Four different AMD model systems were studied: Fe(III)/As(V), Fe(III)/ As(III), Fe(II)/As(V), and Fe(II)/As(III) with the initial concentrations for Fe(III) 260 mg/L, Fe(II) 300 mg/L, As(V), and As(III) 8 mg/L. Essentially quantitative removal of arsenic and iron was achieved in all four systems, and the results were independent of whether the pH was adjusted electrochemically or by the addition of NaOH. Current efficiencies were ∼85% when the pH of the effluent was 4-7. Residual concentrations of arsenic were close to the drinking water standard proposed by the World Health Organization (10 µg/L), far below the mine waste effluent standard (500 µg/L).
Introduction Arsenic is both acutely and chronically toxic. Water in many areas of the world is naturally contaminated with high levels of arsenic, leading to serious toxicity problems among those who use such water supplies for drinking (1-3). A less publicized source of environmental arsenic contamination arises from acid mine drainage (AMD). For example, cattle grazing on abandoned mine lands and drinking from creeks contaminated by acid mine drainage have been found to concentrate heavy metals such as arsenic and zinc in their milk (4), liver, muscle, and blood (5). AMD is caused by biological oxidation of sulfide-bearing minerals, leading to the production of sulfuric acid, eq 1.
FeS2(s) + H2O(l) + 7/2O2(g) f Fe(aq)2+ + 2SO42-(aq) + 2H+(aq) (1) Mine tailings are especially prone to acidification, because of the high surface area of the crushed rock and access to atmospheric oxygen. Since pyrite (FeS2) is a common source * Corresponding author phone: (519)824-4120 ext 53962; fax: (519)766-1499; e-mail:
[email protected]. 4500
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of oxidizable sulfide, AMD is usually associated with high levels of dissolved iron, besides low pH. The pH of AMD is commonly between 2 and 5 (6), although pH < 0 has been observed at certain sites (7). Low pH causes solubilization of many metals, including lead, nickel, copper, zinc, and arsenic, depending on the geochemistry of the site. This metal-bearing solution may enter surface water during spring runoff or enter groundwater by leaching, posing a direct threat to aquatic life and drinking water supplies. Once started, AMD is virtually impossible to reverse (8); the ideal situation is the prevention of AMD formation by excluding air from the tailings, as implied by eq 1, but when this is not possible, remediation is necessary. Like sulfidic metal ores, coal is frequently associated with both pyrite and arsenic sulfides such as realgar (AsS) and orpiment (As2S3). Hence AMD from both coal and metal mines often contains high concentrations of dissolved arsenic (7, 9). The Canadian mining industry generates 1 million tons of waste rock and 950 000 tons of tailings per day, totaling 650 million tons of waste per year, most of which have the potential to cause AMD. Cleanup estimates for existing AMD sites in Canada are in the range $2-5 billion (10). AMD is conventionally treated with finely powdered limestone (CaCO3) or lime (Ca(OH)2) to raise its pH, whereupon the dissolved metals precipitate out as basic metal carbonates or oxyhydroxides. Although lime and limestone are cheap and readily available, the resulting sludge is difficult to dewater and costly to dispose of due to its metal content (11). Moreover, when the AMD contains arsenic, the calcium arsenate in the sludge is unstable with respect to atmospheric CO2, which converts it into calcium carbonate, releasing arsenic back to solution (12). Several solid materials adsorb dissolved arsenic. Some of these are exotic, such as basic yttrium carbonate (13), but much attention has been focused on zerovalent iron corrosion products (14-16) and iron(III) hydroxide (17-20), the latter of which is of interest in the context of AMD because, as already noted, AMD is usually rich in dissolved iron. Indeed, one of the environmental problems associated with AMD is the deposition of iron(III) hydroxide as an unsightly orange precipitate downstream of the AMD source, through oxidation of iron(II) and subsequent precipitation, as the stream pH rises upon admixture with uncontaminated water (4, 5, 21). Arsenic removal by adsorption to iron hydroxide is widely used in the treatment of drinking water and wastewater, and the adsorption envelopes and kinetics for arsenite with iron(III) hydroxide have been well studied (17-19). The sorption of arsenic is complicated, however, because inorganic arsenic is found in the environment both as As(III) (arsenite) and As(V) (arsenate), each of which can exist in both neutral and anionic forms, depending on the pH (22); hence the precipitation of arsenic is pH-dependent. The electrochemical remediation of AMD (but not arseniccontaining AMD) has been pursued in our own (21) and other (23-26) laboratories. Our approach was to raise the pH of AMD by cathodic reduction of H+ to H2 and subsequently to remove iron by air oxidation of Fe2+ to Fe3+, which then precipitated as Fe(OH)3, before release of the treated AMD to the environment. A rather different electrochemical proposal for the amelioration of AMD has been made by Shelp et al. (27, 28) who suggested establishing a galvanic electrochemical cell in situ using pyrite as cathode and iron or other metals (Al, Zn) as sacrificial anode. The voltage thus generated was high enough to maintain the pH of acid leachate at 5.6 and to alter the redox potential toward the inhibition of sulfide oxidation. Although this system 10.1021/es030359y CCC: $25.00
2003 American Chemical Society Published on Web 08/14/2003
claimed the removal of toxic elements such as Al, Cd, Co, Cu, and Ni from solution, no mention was made about As removal. The removal of arsenic by direct electrolysis seems impractical, based on our recent work (29), due to low current efficiencies and the formation of the toxic gas arsine. However, knowing that inorganic arsenic can sorb to iron(III) hydroxide (12, 17-19), we undertook the present study in order to assess whether arsenic remediation could be combined with our previously developed electrochemical treatment for the removal of iron.
Experimental Section A. Materials. Iron(II) sulfate heptahydrate FeSO4‚7H2O was purchased from VWR (Mississauga, ON). Iron(III) sulfate pentahydrate Fe2(SO4)3‚5H2O was purchased from Aldrich. Sodium hydroxide, anhydrous sodium sulfate, and sulfuric acid (98%) were purchased from Fisher Scientific Company (Toronto, ON). Sodium arsenate heptahydrate NaH2AsO4‚ 7H2O and sodium m-arsenite NaAsO2 were purchased from Sigma (Mississauga, ON). Iron and arsenic standard solutions were purchased from Alfa Aesar (Ward Hill, MA). Argon gas was obtained from BOC gases (Mississauga, ON). All solutions were prepared with deionized water. The material used for the cathode was stainless steel plate (composition Fe:Cr:Ni ) 70:19:11), supplied by CFF Specialties (Hamilton, ON). The DuPont Nafion-424 cation exchange membrane (CEM) used to divide the electrochemical reactor was purchased from Electrosynthesis Company (Lancaster, NY). B. Apparatus. Two kinds of electrochemical reactors were used: a plug-flow reactor and a batch reactor. The Plexiglas plug flow reactor, which was built in our laboratory, consisted of two compartments each having dimensions 58 mm × 15 mm × 4.5 mm and separated by a DuPont Nafion-424 cation exchange membrane. The cathode was a stainless steel (SS) plate with dimensions 40 mm × 12.5 mm. The dimensionally stable anode was a grid of IrO2-coated Ti (∼7 cm2). The electrodes were mounted vertically, with electrical connections of stainless steel wire. The glass batch reactor, which was also built in our laboratory, consisted of two cylindrical compartments disposed horizontally and separated by a Nafion 424 CEM. The two glass cylinders were connected with a butt joint SVL connection. The cathodic compartment had a volume of 38 mL, and the anodic compartment had a volume of 32 mL. The SS cathode and IrO2/Ti anode faced the edges of the two cylinders. Each SS electrical feeder passed through a hole in the cylindrical body, which also assured the contact with atmosphere and possibilities of solution supply and withdraw. Cells were controlled with an EG & G model 363 Potentiostat/Galvanostat operated in the galvanostatic mode. C. Experimental Procedures. Chemical Coprecipitation. As(V) stock solutions (40 mg/L and 120 mg/L) were prepared from sodium arsenate heptahydrate NaH2AsO4‚7H2O. As(III) stock solutions (40 mg/L and 120 mg/L) were prepared daily from sodium m-arsenite NaAsO2. Iron(III) sulfate pentahydrate Fe2(SO4)3‚5H2O (Aldrich) was used for the preparation of 3000 mg/L Fe(III) stock solution, with pH adjusted to 1.4 by addition of 2 mL of concentrated sulfuric acid per liter. pH was monitored using an IONcheck 10 standard pH meter. Volumes of Fe(III) and arsenic stock solutions were mixed in a 50 mL volumetric flask, to give defined ratios Fe:As; the pH was then adjusted to a desired value by addition of increments of 0.1 M NaOH solution or 0.1% (v/v) sulfuric acid. Water was then added to a final volume of 50 mL. Final concentrations of arsenic were in the range 8-60 mg/L. In the case of solutions containing Fe(III) and As(V), the solutions were left overnight and then filtered with 0.2 µm Nylon syringe filter for elimination of suspended materials,
FIGURE 1. Principle of operation for electrochemical experiments. and the final pH was recorded. In the case of Fe(III) and As(III), the solutions were first bubbled with argon for 5 min to avoid possible oxidation of As(III). Electrochemical Experiments. The principle of operation for both flow reactor and batch cell is presented in Figure 1. The Plexiglas reactor was operated in plug-flow mode, with separate solutions passed through the cathodic and anodic compartments at equal flow rates of 1 mL/min, using a Masterflex Compact/Low Flow pump. The catholytes were aqueous solutions of arsenic and iron compounds, and the anolyte was 0.1 M Na2SO4. Four different AMD model systems were studied with this plug flow reactor, Fe(III)/As(V), Fe(III)/As(III), Fe(II)/As(V), and Fe(II)/As(III), with concentrations of 260 mg/L for Fe(III), 300 mg/L for Fe(II), and 8 mg/L arsenic, with the initial pH adjusted to 2.0-2.1. The pH of the catholyte effluent increased gradually, and samples for analysis were not collected until the pH of the effluent stabilized (∼1 h). Samples were stored for 24 h for complete coprecipitation (19) and then filtered with a 0.2 µm Nylon syringe filter to eliminate suspended materials. The pH was measured and recorded after filtration. After each experiment, the cell was taken apart and washed with deionized water, and the electrodes were checked for fouling and electrodeposition. The glass reactor was operated in batch mode, only one AMD model being studied with this kind of reactor. The catholyte was 300 mg/L of Fe(II) and 8 mg/L of As(V) solution, and the anolyte was 0.1 M Na2SO4. A constant current of 50 mA was applied to the reactor for a determined period of time, and then the catholyte was withdrawn from the reactor and aerated for 30 min. Samples were stored for 24 h and then filtered with a 0.2 µm Nylon syringe filter before being analyzed for arsenic and iron. After each experiment, the cell was taken apart and washed with deionized water, and the electrodes were checked for fouling and electrodeposition. D. Analysis. The final filtrate was analyzed for iron by Atomic Absorption Spectroscopy (AAS, Perkin-Elmer, model: 2380); the detection limit was 50 µg/L. Arsenic was analyzed by hydride generation ICP-AES (Thermo Jarrell Ash, model: Atomscan 16) with direct flow injection and no prereduction technique. The method was modified from other HG-ICPAES methods published in the literature which required prereduction of As(V) to As(III) (30, 31). Those authors found that if the tubing was long enough and the concentration of NaBH4 was sufficient, As(V) was reduced to arsine, and thus there was no need to prereduce As(V) to As(III). The advantages of the method include no sample preparation and avoidance of the delay of 1.5 h for prereduction of As(V). As(V) solution prepared from NaH2AsO4‚7H2O was first analyzed by direct ICP-AES, and then the solution was diluted to lower concentration and analyzed by HG-ICP-AES with prereduction of As(V) to As(III) (30, 31) and also analyzed by HG-ICP-AES without prereduction of As(V). These three methods agreed well with each other when the prereduction VOL. 37, NO. 19, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 2. Influence of solution final pH on As(V) removal from Fe(III)/As(V) solutions (Fe(III) concentration fixed at 300 mg/L while As(V) concentrations varied): [ As(V) ) 8 mg/L, 0 As(V) ) 20 mg/L, 2 As(V) ) 30 mg/L, 9 As(V) ) 60 mg/L.
FIGURE 3. Influence of solution final pH on As(III) removal from Fe(III)/As(III) solutions (Fe(III) concentration fixed at 300 mg/L while As(III) concentrations varied): [ As(III) ) 8 mg/L, 9 As(III) ) 20 mg/L, 4 As(III) ) 30 mg/L, × As(III) ) 60 mg/L. time for As(V) was sufficient (5 h). To obtain consistent results, one has to wait for 5 h, but even this is less than the 24 h delay advocated by Edwards et al. (31). The optimized analytical conditions were reaction tube 5 m, reducing agent 50% HCl with 1% NaBH4 in 1% NaOH. NaBH4, HCl, and the sample solutions were simultaneously delivered to the nebulizer by a peristaltic pump through Tygon tubing. The detection limit for arsenic with this method was 5 µg/L. The possible formation of arsine during electrolysis was checked by sweeping product gases out of the catholyte with argon through a U-tube containing silver diethyldithiocarbamate and morpholine in chloroform. The solution that was originally yellowish would give a red coloration if arsine were present (32).
Results and Discussion Chemical Adsorption. For both As(V)/Fe(III) and As(III)/ Fe(III), the concentration of arsenic remaining in solution after the ultimate filtration was studied as a function of final pH for different initial ratios of Fe:As (Figures 2 and 3). The residual concentration of As(V) was minimized at ∼4 < pH < 6, that of As(III) at ∼6 < pH < 9. In each case, the residual concentration of As in the optimal pH range was
