Trichloroethylene Transformation in Aerobic Pyrite Suspension

Jul 30, 2009 - The pathways and kinetics of trichloroethylene (TCE) degradation in aerobic pyrite suspension were investigated. The detection of hydro...
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Environ. Sci. Technol. 2009, 43, 6744–6749

Trichloroethylene Transformation in Aerobic Pyrite Suspension: Pathways and Kinetic Modeling HOA T. PHAM,* KOICHI SUTO, AND CHIHIRO INOUE Graduate School of Environmental Studies, Tohoku University, 6-6-20 Aoba, Aramaki-Aza, Aoba-ku, Sendai, Miyagi, 980-8579, Japan

Received March 3, 2009. Revised manuscript received June 29, 2009. Accepted July 10, 2009.

researchers have worked on the mechanism involved in the pyrite oxidation reaction (15-24). In a review article, Ramstidt and Vaughan (22) proposed the mechanism of pyrite surface oxidation reactions in which the hydroxyl radical is produced while the electron transfers from the iron site of the surface to oxygen absorbed at the site. Therefore, the concentration of oxygen in the system may directly decide the quantity of radical produced. However, this correlation has not yet been verified. This paper aims to identify the mechanism of TCE degradation reactions by measuring the amount of radical produced in aerobic pyrite suspension and to propose the TCE degradation pathways and kinetic by investigating the pathways and rates of each elementary step involved in the degradation reactions of TCE.

Materials and Methods The pathways and kinetics of trichloroethylene (TCE) degradation in aerobic pyrite suspension were investigated. The detection of hydroxyl radical in aqueous pyrite suspension suggested that TCE was degraded by this strong oxidant. The reaction pathways of TCE degradation were proposed in which the degradation of TCE to formic acid and finally to CO2 was the main route. Degradation of TCE to oxalic acid and to dichloroacetic acid were found as minor pathways. Degradation rates of TCE to formic acid, glyoxylic acid, and dichloroacetic acid were obtained using kinetic model at 1.2 × 10-2, 9.8 × 10-4 and 4.6 × 10-4 h-1, respectively.

Introduction Trichloroethylene (TCE) is an important pollutant found in soil and groundwater. The release of this widely used compound into subsurface environments has contaminated many groundwater aquifers, prompting concerns due to the possible effects on human health and persistence in aquatic environments. Degradation of TCE, as well as other aliphatic chlorinated compounds, has received much attention and investigation. Cost-effective options for treating this pollutant are required. There has been extensive use of zerovalent iron (1), pure chemical oxidation (2), and photochemical oxidation (3, 4). Recently, the natural mineral iron disulfide (pyrite) was found to be active in the degradation of chlorinated pollutants (5-9). In our previous paper (5), degradation of TCE in pyrite suspension was found as a potential treatment option. The rate of this degradation depended on the concentration of oxygen; TCE degradation occurred under aerobic condition. Under anaerobic condition, a very limited amount of TCE can be transformed (5, 7). In aerobic pyrite solution, TCE transformed to dichloroacetic acid (DcA), glyoxylic acid (GA), oxalic acid (OA), and formic acid (FA), and finally transformed to CO2 and chloride ion. While these degradation reactions might be important for the application of pyrite in remediating chlorinated pollutants, the degradation pathways and kinetic of these reactions are still unclear. The oxidant involved is hydroxyl radical ( · OH), as this radical forms during the pyrite dissolution reaction (10-14); however, the role of oxygen in generation of this radical is still unclear. The presence of oxygen in the aqueous pyrite system causes a well-known phenomena referred to as acid mine drainage, which is the oxidation reaction of pyrite. Several * Corresponding author phone/fax: +84-(0)22-795-7404; e-mail: [email protected]. 6744

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Materials. Pyrite mineral was obtained from the Yanahara mine, Japan. Pyrite samples were treated following the procedure described in our previous paper (5). Pyrite used in this study had a particle size of 20-38 µm and specific surface area of 0.2 m2/g (determined using the BET method). The pyrite contained minor amounts of impurities including Si, Zn, and Cu. The molar percentages for S, Fe, Si, Zn, Cu, which were all the elements present, were 61.7, 33.3, 4.0, 0.7, and 0.2%, respectively. Standard TCE solution was obtained at more than 97% purity from GL Sciences Inc., Japan, and stored at 4 °C. Four carboxylic acidssdichloroacetic acid 98%, glyoxylic acid 95%, oxalic acid 98%, and formic acid 99%swere obtained from Wako Co., Japan. 3′-p-aminophenyl fluorescein (APF) (5 mM) was purchased from Daiichi Co., Japan. Horseradish peroxidase (HRP), reagent-grade H2O2 solution (30%), potassium permanganate, and sodium oxalate were purchased from Wako Co., Japan. Kinetic Experiments. TCE degradation in aerobic pyrite suspension and the generation of carboxylic acids, CO2, and chloride with time were reported in our previous paper (5) and used to determine TCE degradation pathways and kinetic in this study. For investigation of the TCE degradation pathways, laboratory batch experiments were conducted for each of the identified intermediates. The experimental procedure was that used in our previous study (5) and it is only briefly described here. The kinetic study was conducted in 26-mL glass vials with silicone-septum-lined screw-top caps. Each vial was filled with 10 mL (1 mM) of carboxylic acid solution and 1 g of pyrite, and then tightly capped. The oxygen concentration in the closed system was initially equivalent to that in the atmosphere. The prepared vials were placed on a vibratory shaker and shaken at 400 rpm in an incubator at 25 °C in the dark until sampled. During each period of sampling, two vials were taken. The vial headspaces were tested for CO2 gas generation. The vials were then opened and the aqueous solutions were filtered by a micromembrane filter (0.45 µm). After filtration, the carboxylic acid and chloride concentrations of the solutions were determined by high-performance liquid chromatography (HPLC). Hydroxyl Radical Concentration Measurement. The procedure for measuring the · OH concentration was a modified version of the method described by Cohn et al. (11). APF was used as the fluorescence response compound. APF is not fluorescent until it reacts with · OH or H2O2 in the presence of HRP, resulting in the cleavage of the aminophenyl ring from the fluorescein ring system, which is highly fluorescent. In this study, the concentration of the radical 10.1021/es900623u CCC: $40.75

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FIGURE 1. Concentration profiles of organic acids (initial amount of 10 µmol) and their intermediates and products generated in aerobic pyrite suspension (20 m2/L) (without addition of TCE) and the calculated total carbon mass with time: (a) OA, (b) FA, (c) GA, and (d) DcA. produced in the pyrite system under aerobic condition was measured by comparing with the calibration curve of the concentration of radical produced from H2O2 solution in the presence of HRP. Radical Concentration Measurement for Building the Calibration Curve. A known amount of H2O2 dissolved in 50 mM potassium phosphate buffer at pH 7.4, along with 0.02 mg/mL HRP (equivalent to 4.5 units/mL), was added to the vial. APF was then added (resulting in 10 µM APF in the solution). The reaction between H2O2 and HRP oxidized APF, resulting in an increase in fluorescence intensity. The vials were shaken for 30 min at 400 rpm in an incubator at 25 °C, and then aqueous solution was taken for the measurement of fluorescence intensity using a fluorescence spectrophotometer. Absolute concentrations of H2O2 were determined by titration with KMnO4 solution. Sodium oxalate was used to calibrate the KMnO4 concentration via the titration method. Radical from Pyrite Suspension. This procedure was similar to that of the kinetic study, except that carboxylic acid solution was replaced by potassium phosphate buffer (50 mM, pH 7.4). The APF stock (20 µL) was injected into the prepared vials using a microsyringe (resulting in 10 µM APF in the solution). Afterward, the vial was placed on a vibratory

shaker and shaken at 400 rpm in an incubator at 25 °C until sampled. The fluorescence intensity of the filtered aqueous solution was measured using a fluorescence spectrophotometer. Chemical Analysis. The CO2 concentration was determined using a gas chromatography-thermal conductivity detector (GC-323, GL Sciences) and packed column (Porapak Q 50/80 column, GL Sciences), with an injection temperature of 120 °C, oven temperature of 60 °C, and detector temperature of 100 °C. Chloride ions in the supernatant were analyzed by HPLC (L-7300, Hitachi) using a GL-IC-A25 column (4.6 mm × 150 mm, Hitachi) (4 mM Na2CO3 eluent, 1.0 mL/min flow rate, 40 °C column temperature). Organic acids were analyzed by HPLC (L-7200, Hitachi) using a Gelpack GL-C610H-S column (7.8 mm × 300 mm, Hitachi) (0.1% H3PO4 eluent, 0.5 mL/ min flow rate, 50 °C oven temperature, UV detector set at 210 nm). Concentrations were quantified by comparisons of gas chromatography and HPLC peaks using five-point standard curves. Fluorescence Analysis. Fluorescence was measured using a fluorescence spectrophotometer (Hitachi F-4500). The excitation and emission wavelengths were 485 and 515 nm, respectively, and the scan speed was 240 nm/min. VOL. 43, NO. 17, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Pseudo-First-Order Degradation Rate Constants for TCE and the Organic Acids in Pyrite Suspension (20 m2/L)

reactant

rate constant (h-1)

rate constant ((h · m2)-1)

correlation coefficient (R2)

TCE dichloroacetic acid glyoxylic acid formic acid oxalic acid

1.30 × 10-2 1.50 × 10-3 1.15 × 10-2 2.29 × 10-2 2.00 × 10-4

6.50 × 10-2 7.50 × 10-3 5.75 × 10-2 1.15 × 10-1 1.00 × 10-3

1.00 0.99 0.99 0.98 0.55

Reaction Rates. The degradation rates of TCE and the carboxylic acids were evaluated using a kinetic modeling program developed using Digital Visual Fortran. Reaction pathways were hypothesized and evaluated on the basis of the quality of the fitting to the experimental data. Mass transfer resistance at the vapor/liquid interface was not considered as these phases were assumed to be in equilibrium.

Results and Discussion Degradation of TCE Intermediates in Aerobic Pyrite Suspension. To assess the degradation pathway of TCE, each identified intermediate of TCE degradation (OA, GA, DcA, and FA) was used as a starting material to conduct degradation experiments under the same conditions as for TCE experiments. These batch experiments had oxygen concentrations initially equivalent to that of the atmosphere. Under this condition, oxygen was sufficient for the degradation of TCE into its final products (5). Figure 1 shows the degradation profiles, generated intermediates, and products of these carboxylic acids with time. All compounds degraded in aerobic pyrite suspension at different rates. Ninety percent of FA had degraded after 100 h reaction time, 60% of GA had degraded after 90 h, and 30% of DcA had degraded after 240 h. OA degraded very slowly during the experimental course. Carbon balances were obtained for all batches. Therefore, the adsorption of organic acids on the mineral surface2 was minor and not taken into account in kinetic calculations. Apparent Reaction Rate Constants. The degradation rates of all intermediates (OA, FA, DcA, and GA) in aerobic pyrite suspension were calculated by assuming pseudo-first-order reactions for all degradations. Table 1 shows the apparent rate constants calculated from the experimental data (including the degradation rate of TCE reported by Pham et al. (5). The rate of TCE degradation was lower than the rate of FA degradation and similar to the rate of GA degradation, but 8 times the rate of DcA degradation and 65 times the rate of OA degradation. This means that FA was more reactive in the aerobic pyrite suspension than TCE and GA were. DcA was less reactive than TCE under the same conditions. OA was the most stable compound in the aerobic pyrite suspension. The difference in the degradation rates of these organic compounds may be due to the difference in the chemical structure. Several studies have reported the structure depending of reactions between · OH and organic compounds (25-28). Hydroxyl Radical Generated in Pyrite Suspension under Aerobic Conditions. The fluorescence intensities obtained with different H2O2 concentrations are presented in the Supporting Information. The hydroxyl radical generated in pyrite aqueous suspension was quantified using the obtained calibration curve. The fluorescence intensities from aerobic pyrite suspension were monitored for 170 h. Fluorescence intensity increased significantly with time. It was rapidly generated within 48 h, reaching 5000 (a.u.), then increased slowly, reaching 5600 after 168 h. The quantification of · OH is shown in Figure 2. The quantity of · OH generated after 6746

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FIGURE 2. Calculated quantity of · OH generated from aerobic pyrite suspension (pyrite 20 m2/L suspended in phosphate buffer (50 mM and pH 7.4) containing APF 10 µM; initial oxygen concentration was equivalent to that of atmospheric air). Quantity of radical was calculated as an equivalent amount of H2O2, which reacted with HRP and oxidized the APF. 100 h was equivalent to the amount of radical generated from 10 µM H2O2 in the presence of HRP. Therefore, this radical can be the important oxidant for the degradation of TCE and its intermediates in aerobic pyrite suspension. Degradation Pathways of Organic Acids in Aerobic Pyrite Solution. When OA was used as the starting material, there was very little degradation of the acid within 350 h. CO2 was detected as the only degradation product (Figure 1a). The degradation of OA to CO2 by hydroxyl radical generated in other processes has been reported; for example, in electrochemical oxidation reactions (29, 30) and oxidation reactions in ozone treatment and UV irradiation (31). Because · OH was found in the aerobic pyrite suspension, the reaction scheme for OA degradation in pyrite suspension is proposed in eq 1. C2H2O4 + 2 · OH f 2CO2 + 2H2O

(1)

In experiments with 10 µmol of FA, more than 90% of FA degraded by 120 h, and CO2 was the only product found (Figure 1b). FA is oxidized to CO2 by hydroxyl radical in ozone treatment and UV irradiation (31) and electrochemical oxidation (32). The reaction scheme proposed for the degradation of this acid by radical generated from aerobic pyrite suspension is as eq 2. CH2O2 + 2 · OH f CO2 + 2H2O

(2)

In the experiments with 10 µmol GA as starting material, GA degrades to produce OA and CO2 (Figure 1c). Because of the low degradation rate from OA to CO2, there must be another pathway in which GA transforms to CO2 through other intermediates besides OA. However, the intermediate of this pathway was not detected in our study. The degradation of GA to OA agreed with the results obtained by Matsumoto and Kozai (31). The reaction schemes proposed for this acid are as eqs 3 and 4. C2H2O3 + 2 · OH f C2H2O4 + H2O

(3)

C2H2O3 + 2 · OH f 2CO2 + 2H2O

(4)

Similar experiments were conducted for 10 µmol DcA. DcA degraded with time to produce OA and CO2, as shown

FIGURE 3. Proposed degradation pathways of TCE in aerobic pyrite suspension. k1 to k9 are the reaction rate constants of each reaction. All compounds are detected and measured intermediates or products. in Figure 1d. The degradation of DcA to CO2 agreed with the results obtained in Fenton process (25). Besides DcA transforming to CO2 through OA, there must be another pathway by which DcA transforms to CO2 through intermediates besides OA. The reaction schemes for DcA are proposed as eqs 5 and 6. C2H2O2Cl2 + 4 · OH f C2H2O4 + 2Cl- + H2O

(5)

C2H2O2Cl2 + 4 · OH f 2CO2 + 2Cl- + H2O

(6)

Proposed TCE Degradation Pathway. The TCE degradation profile and the production of organic acids, CO2 and chloride were reported in our previous paper (5). That information is used here to analyze the kinetic of TCE degradation. In addition, hydroxyl radical was detected in the aerobic pyrite suspension. Eliminating radicals from TCE/ pyrite suspensions by adding APF stabilized TCE (data not shown), because APF was very reactive with · OH. From the degradation profiles of TCE (5) and its intermediates in the aerobic pyrite system (Figure 1), it can be seen that TCE transformed to organic acids and these organic acids

transformed to CO2. These pathways are in agreement with pathways of TCE degradation by Fenton process reported by Qiang et al. (25), in which TCE degraded to dichloroacetic acid and finally to CO2. In their study, they detected the intermediates by extraction with MTBE, then derivation by diazomethane, and analyzing by GC/MS. It was hypothesized that some simple byproducts could be coeluted with MTBE after diazomethane derivatization and thus were not detected by GC/MS. Therefore only dichloroacetic acid was detected as the main intermediate of TCE degradation by Fenton process. In addition, these pathways also agreed with pathways of TCE degradation by permanganate (2) or by hydroxyl radical generated from UV/H2O2 oxidation (3), radiation (32), and UV photolysis (4). However, for radiation or UV photolysis, there are several intermediates besides the above-mentioned organic acids during the chain reactions initiated by hydroxyl radical. In addition, Cai and Guengerich (34) reported the degradation pathways of TCE oxide, in which TCE oxide transformed to GA, FA, DcA, and finally CO2. TCE oxide is a reactive electrophile formed during TCE oxidation and rearranges as acylating intermediates (34). Based on our obtained results, TCE degradation pathways in aerobic pyrite suspension are those proposed in Figure 3. TCE degrades following three main pathways, which are the degradations to DcA, GA, and FA, and then these organic acids transform to CO2 as the final products. Pseudo-First-Order Kinetic Model. A pseudo-first-order kinetic model was developed to calculate the rates of all three pathways of TCE degradation. The calculated results were evaluated on the basis of the quality of fitting to the experimental data. Three assumptions were hypothesized for the kinetic calculation: (1) oxygen is always available in the reactors, (2) all degradations are first-order reactions that only depend on the concentrations of organic compounds, and (3) pyrite dissolution reactions do not affect the degradation reactions of organic compounds. The time derivatives of the concentrations of the compounds are formulated and solved using the Runge-Kutta method (35) (Supporting Information). Data fitness between computed and observed values was evaluated using the root-meansquare values (rms). Among the nine parameters, k4 and k9 can be estimated using the experimental data for OA and FA degradations (Table 1). Values of k7 and k8 can be obtained from DcA experiments. The DcA degradation profiles are shown in Figure 4a. Values of rms for DcA, OA, and CO2 were

FIGURE 4. Concentration profiles of DcA and GA and their intermediates and products generated in aerobic pyrite suspension. Symbols denote experimental data. Lines indicate data calculated using the kinetic model; (a) DcA experiments, k7 and k8 were obtained as 1.8 × 10-4 and 1.3 × 10-3 h-1, respectively; b) GA experiments, k5 and k6 were obtained as 1.5 × 10-3 h-1 and 1.0 × 10-2 h-1, respectively. VOL. 43, NO. 17, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 2. Reaction Rate Constants of TCE Reaction Pathways (Pyrite 20 m2/L) degradation

rate constant

k (h-1)

k ((h · m2)-1)

TCEf formic acid TCEf glyoxylic acid TCE f dichloroacetic acid formic acid f CO2 glyoxylic acidf CO2 glyoxylic acid f oxalic acid dichloroacetic acid f oxalic acid dichloroacetic acid f CO2 oxalic acid f CO2

k1 k2 k3 k4 k5 k6 k7 k8 k9

1.2 × 10-2 9.8 × 10-4 4.6 × 10-4 2.3 × 10-2 1.5 × 10-3 1.0 × 10-2 1.8 × 10-4 1.3 × 10-3 2.0 × 10-4

5.8 × 10-2 4.9 × 10-3 2.3 × 10-3 1.1 × 10-1 7.5 × 10-3 5.0 × 10-2 9.0 × 10-4 6.6 × 10-3 1.0 × 10-3

0.015, 0.006, and 0.025 respectively, which represented a good fit between calculation and experimental data. Values of k5 and k6 can be obtained from GA experiments. Values of rms for GA, OA, and CO2 were 0.050, 0.011, and 0.016 respectively, which represented the fitness of the kinetic model results to experimental data in GA experiment (Figure 4b). Finally, the three unknown parameters (k1, k2, and k3) can be obtained using kTCE and the time derivative equations of all compounds. The calculated kinetic data were fitted to the experimental data to obtain best-fit values for k1, k2, and k3. The obtained degradation rates of all pathways are presented in Table 2. Figure 5 shows the time course of the contents of TCE, intermediates, and products with time, presenting data for both the experiment and kinetic model calculation. The calculated profiles of DcA, GA, and OA in TCE transformation pathways were close to experimental data (rms values for DcA, GA, and OA were calculated at 0.01, 0.07, and 0.02, respectively). However, the calculated FA profile lay below the experimental profile, showing that the calculated degradation rate of FA was higher than the real rate (rms of FA was 1.36). Therefore, the degradation rate obtained in the experiments using FA as the starting compounds was higher than the degradation rate of this intermediate during TCE experiments. The difference can be explained by the difference in the aqueous conditions

FIGURE 5. Concentration profiles of TCE and its intermediates and products in aerobic pyrite suspension. Symbols denote experimental data. Lines indicate data calculated using the kinetic model. 6748

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among the experiments. In TCE experiment, there were more compounds with different solubility and degradability generated compared to the experiments with FA. Their simultaneous effects could lead to the difference in degradation rates between TCE experiments and the FA experiments. The calculated profiles of CO2, Cl, C mass, and Cl mass lay above the experimental profiles (rms values of 4.03, 3.12, 3.35, and 2.20, respectively). Experimental obtained C mass and Cl mass decreased with time. There was probably the loss of TCE out of the vials or the absorption of TCE inside the vials (5). In addition, other intermediates of chain reactions initiated by hydroxyl radical may have been produced, which were not detected in this study. However, the sum of the loss and the amount of undetected intermediates accounted for less than 20% of the total mass. To determine the main pathway of TCE oxidation, we look at k1 to k3 as a percentage of kTCE, which are 89.0, 7.5, and 3.5%, respectively, and k4, k5, k8, and k9 as a percentage of kCO2, which are 88.5, 5.8, 5.0, and 0.7%, respectively. These data indicated that the degradation of TCE to FA and FA to CO2 is the main pathway of TCE degradation. The chlorinated intermediate, which is DcA, only accounts for a very small fraction of the intermediates.

Acknowledgments This work was financially supported by the Japan Society for the Promotion of Science (A) 17206089. We thank Dr. Kayamori for his help in measuring fluorescence using the spectrofluorometer.

Supporting Information Available Calibration curve for hydroxyl radical concentration (Figure S-1) and pseudo-first-order kinetic model formulation and calculation. This material is available free of charge via the Internet at http://pubs.acs.org.

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