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Environ. Sci. Technol. 2003, 37, 5823-5828

Trichloroethylene Degradation in a Coupled Anaerobic/Aerobic Reactor Oxygenated Using Hydrogen Peroxide B. TARTAKOVSKY, M.-F. MANUEL, AND S. R. GUIOT* Biotechnology Research Institute, NRC, 6100 Royalmount Avenue, Montreal, Quebec, Canada H4P 2A2

In this work, trichloroethylene (TCE) degradation under combined anaerobic-aerobic conditions was studied in an ethanol-fed biofilm reactor oxygenated using hydrogen peroxide. The reactor was inoculated with a biomass originating from an anaerobic digestor. Granulated peat was added to the reactor as a substratum for biofilm development. Extensive characterization of reactor populations using activity tests and PCR analysis revealed the development of a mutualistic consortium, particularly methanotrophic and methanogenic microorganisms. This consortium was shown to degrade TCE by a combination of reductive and oxidative pathways. A near complete degradation of TCE at a load of 18 mg LR-1 day-1 was evidenced by a stoichiometric release of inorganic chloride.

Introduction Biological degradation of complex organic molecules is a multistep process which proceeds via a sequence of biotransformations and often requires a mixed bacterial population. Moreover, a combination of anaerobic (reductive) and aerobic (oxidative) conditions is often required to achieve mineralization rather than partial transformation of complex organic molecules. Although combined anaerobic/aerobic treatment can be achieved in a sequential reactor system (1), recent advances in the development of coupled anaerobic/aerobic systems have demonstrated that a steep oxygen gradient across a biofilm can be exploited to create combined anaerobic/aerobic conditions in a single reactor volume (24). Successful examples of anaerobic/aerobic coupling include degradation of azo dyes (5) and chlorinated organic compounds (2,6-8). Anaerobic degradation of TCE proceeds via formation of dichloroethylenes (DCEs) and vinyl chloride (VC), and the rate of dechlorination decreases with decreasing extent of chlorination (9). Consequently, reductive dechlorination often results in the accumulation of dechlorination intermediates. At the same time, the dechlorination intermediates can be efficiently degraded by aerobic bacteria, methanotrophs in particular (10,11). Thus, a combination of reductive and oxidative degradation pathways was expected to maximize the net degradation rate due to the rapid elimination of toxic intermediates (3). Limited aeration of naturally formed anaerobic granules was shown to result in the formation of a methanotrophic/ * Corresponding author phone: (514)496-6181; fax: (514)496-6265; e-mail: [email protected]. 10.1021/es030340v CCC: $25.00 Published on Web 11/15/2003

Published 2003 by the Am. Chem. Soc.

methanogenic bacterial consortium suitable for degradation of chloroorganic compounds (7,12) Aeration of biomass granules, however, resulted in losses of chloroorganic compounds due to volatilization. Furthermore, proliferation of aerobic bacteria decreased density of biomass granules, thus decreasing the volumetric efficiency of the reactor. In this study the performance of the anaerobic/aerobic bacterial consortium was improved by using hydrogen peroxide as a source of oxygen and granulated peat as a substratum for biofilm formation. This reactor system was used to study TCE degradation under combined anaerobic/ aerobic conditions. Throughout the study, the development of a TCE degrading methanogenic/methanotrophic microbial consortium was monitored using biomolecular analysis of microbial community DNAs (13,14) in combination with metabolic activity tests.

Materials and Methods Chemicals and Analytical Methods. TCE, DCEs, and VC in the effluent were analyzed using the GC headspace method (Tekmar 14-4401-000 headspace autosampler; Sigma 2000 gas chromatograph with a FID detector, Perkin-Elmer, Norwalk, CT). The GC was equipped with 100 m × 0.25 mm × 0.25 µm DB-petro 100 column (Supelco, Bellefonte, PA). The concentrations of TCE, DCEs, and VC in the off-gas were determined using a Sigma 2000 gas chromatograph with a FID detector (Perkin-Elmer, Norwalk, CT). A 300 µL gas sample was injected on a 1.8 m Carbopack B/1% SP-1000 column (Supelco) using nitrogen as the carrier gas. The off-gas composition (CH4 and CO2) was determined by gas chromatography (Sigma 2000, Perkin-Elmer). Details of the method are given elsewhere (15). The inorganic chloride concentration was determined using the mercuric thiocyanate method (16). The chemical oxygen demand (COD) of the effluent and volatile suspended solids (VSS) content of the anaerobic sludge were determined using standard methods (16). Dissolved oxygen concentration was measured periodically in the recirculation loop of the reactor using a Cole-Parmer dissolved oxygen meter model 01972-00 equipped with a polarographic electrode 05643-00 (ColeParmer, IL). Reactor Setup and Operation. The experimental setup consisted of two 1.0 L reactors made of glass. The reactors were operated at a hydraulic retention time (HRT) of 17-20 h, a liquid upflow velocity of 4-6 m h-1, and a temperature of 25 °C. Both reactors were loaded with 300 mL of 1-2 mm soaked granulated peat (Berger, Saint-Modeste, Que, Canada) and inoculated with 200 mL of anaerobic sludge obtained from an upflow anaerobic sludge bed (UASB) reactor treating wastewater from a food plant industry (Sensient Flavors Canada Inc, Cornwall, Ontario, Canada). Mixed bacterial community of the inoculum mostly consisted of anaerobic methanogenic microorganisms, although the presence of methanotrophic microorganisms was expected (12). Inoculum characterization showed a fermentative activity of 2-3 g of glucose g-1 of VSS day-1, an acetogenic activity of 30-40 mg of propionate g-1 of VSS day-1, and a methanogenic activity of 0.3-0.5 g of acetate g-1 of VSS day-1. Prior to inoculation, the sludge was homogenized using a homogenizer (Kinematica CH-6010, Brinkmann Instruments, Westbury, NY). The reactors were fed at a rate of 35 mL day-1 with a solution of nutrients containing (in g L-1): KH2PO4, 0.8088; K2HPO4, 1.0324; NH4HCO3, 12.459. The chloride-free trace metal solution was fed separately at the same rate and contained (in mg L-1): FeSO4‚7H2O, 195; H3BO3, 3.5; ZnSO4‚ VOL. 37, NO. 24, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Operational Conditions of Test and Control Reactors phase parameter -1

day-1

COD load, g of COD LR H2O2 load,a g of O2 LR-1 day-1 TCE load, mg LR-1 day-1 duration, day

1

2

3

1.0 0.35b 0 30

1.0 0.7b 0 34

1.0 0.7b 18 31

a Three percent aqueous solution of H O was used. The H O load 2 2 2 2 is shown in oxygen equivalents. b No H2O2 fed to control (anaerobic) reactor.

7H2O, 12.5; CuSO4, 4.5; MnSO4‚H2O, 23.8; Co(NO3)2‚6H2O, 15.5; NiSO4‚6H2O, 7; (NH4)6Mo7O24‚4H2O, 12.5; AlK(SO4)2‚ 12H2O, 1.5; Na2-EDTA, 44.75; MgSO4‚7H2O, 307.5; Na2SeO4, 1.25. The dilution stream contained a bicarbonate buffer (NaHCO3 0.68 g L-1, KHCO3 0.87 g L-1). The carbon source (ethanol) was fed separately using a syringe pump at a rate of 1 g of COD LR-1 day-1. During the TCE-feeding phase of the experiment, TCE was added to the ethanol feed to obtain a TCE load of 18 mg LR-1 day-1, which corresponded to 10 mg L-1 (76 µM) of TCE in the influent. A 3% aqueous solution of hydrogen peroxide was fed to the test (coupled) reactor, while the second reactor was operated under anaerobic conditions. The two reactors were operated through the experimental phases outlined in Table 1. During each phase, constant operational parameters (H2O2, COD, and TCE feeding rates) were maintained for a period of 1 month (35 retention times). Reactor monitoring included measurements of chemical oxygen demand (COD), chloride, TCE, DCEs, and VC concentrations in the effluent and methane, carbon dioxide, TCE, DCEs, and VC in the off-gas. At the end of each experimental phase, the metabolic activity of the biomass was assessed in batch tests for methane consumption, production, and TCE mineralization, as described below. Methanotrophic and Methanogenic Activity Tests. The specific methanotrophic activity of the biomass was determined by measuring the rate of methane consumption under methane and oxygen nonlimiting conditions. The specific rate of methane production (methanogenic activity) was measured under anaerobic conditions using ethanol as a carbon source. Both tests were carried out in 120 mL serum bottles. The methanotrophic test was conducted using a mixture of air and methane (4:1) in the bottle headspace, while for the methanogenic test the bottles were flushed with N2/CO2 and sealed. The bottles were maintained at 30 °C in a rotary shaker at 100 rpm. For the methanotrophic test, each bottle contained 2 mL of peat-attached biomass in 10 mL of low nitrate mineral salt medium, LNMSM (17). For the methanogenic test, bottles were inoculated with 2 mL of peatattached biomass in 0.05 M phosphate buffer with a final volume of 10 mL. The test was initiated upon addition of 50 µL of anhydrous ethanol. The specific activity of the biomass was calculated by dividing the rate of methane consumption or production by the volume of peat-attached biomass in the bottle. Mineralization Tests. Mineralization of TCE under aerobic methanotrophic conditions was tested in 120 mL serum bottles equipped with a KOH trap. Each bottle contained 2 mL of peat-attached biomass in 10 mL of LNMSM. Bottle headspace contained a mixture of air and methane (4:1). The bottles were spiked with 100 000 dpm of 14C-labeled TCE, and the nonradioactive TCE was added to obtain a final concentration of 2 mg L-1. The bottles were incubated at 30 °C on a rotary shaker at 100 rpm. Radiolabeled TCE was mineralized to 14CO2, which was trapped by 1 M KOH solution. The KOH was recuperated periodically from the bottle, and 5824

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CO2 was quantified using a scintillation counter (model 2100 TR, Packard Instrument Company, Meriden, CT). Abiotic Degradation of TCE and DCE in the Presence of H2O2. Abiotic degradation of TCE and DCE in the presence of H2O2 was studied in 120 mL serum bottles filled with 60 mL of distilled water and supplemented with nutrients and microelements to obtain concentrations similar to those in the reactor. The experiments were initiated by adding either TCE or DCE to obtain a final concentration of 76 µM. The bottles were incubated on a rotary shaker at 100 rpm, and gas-phase samples were taken periodically. Liquid-phase concentrations were calculated using Henry’s law constants for TCE and DCE. DNA Extraction from Sludge. A 1 mL volume of sample was centrifuged for 5 min at 10 000xg. The pellet was resuspended in 700 µL TEN (10 mM Tris-HCl (pH 8.0), 1 mM EDTA, 100 mM NaCl, and 5 mg/mL lysozyme). Glass beads (250 mg 0.1 mm diameter, 250 mg 0.5 mm diameter, 1 large 6 mm glass bead) were added to the suspension, vortexed for 30 s, and incubated for 15 min at 30 °C. A 35-µL aliquot of 20% SDS was added, and the suspension was shaken twice for 10 s at 4 m‚s-1 in a FastPrep Instrument (Bio101, Mississauga, ON, Canada). The sample was centrifuged for 15 min at 10 000xg. The supernatant was decanted, and the pellet was resuspended in 145 µL of 10 N ammonium acetate, incubated on ice for 20 min, and centrifuged for 10 min at 10 000xg. Following centrifugation, the supernatant was removed, combined with 10 µL of 1 mg/mL RNAase, and incubated for 20 min at 37 °C. The DNA was purified by phenol-chloroform extraction as described in Tartakovsky et al. (18). Polymerase Chain Reaction (PCR) Amplification. A series of PCR amplifications were done on 2 µL of extracted DNA using primers specific for the functional genes encoding the sMMO (mmoX) and pMMO (pmoA) enzyme complexes in methanotrophs and the methanol dehydrogenase gene (mxaF) in methylotrophs. In addition, phylogenetic primers targeting the 16SrRNA gene of Archaea were used. Details of the primers are given in Table 2. The PCR amplification was performed using the protocol described by Levesque et al. (19). Cycling parameters used in the Perkin-Elmer 2400 thermal cycler (Perkin-Elmer, Mississauga, ON) were an initial denaturation for 5 min at 94 °C followed by 20 cycles at 94 °C for 15 s, 65 °C for 30 s, and 72 °C for 30 s, with a final extension at 72 °C for 7 min. A nested PCR was performed on 2 µL samples from the first PCR amplification using the sMMO specific primers moX536 and moX898 (Table 2) under the same PCR conditions described above.

Results and Discussion Effect of Hydrogen Peroxide on Reactor Consortium. Initial attempts to establish coupled anaerobic/aerobic bacterial consortia were carried out using natural granular anaerobic biomass as a precursor. These experiments, which were carried out at a hydrogen peroxide load of 0.35 g of O2 LR-1 day-1, demonstrated a negative impact of H2O2 on the structural integrity of the biomass granules resulting in granule disintegration and washout within 1 week of reactor operation (results not shown). Consequently, a decision was made to use peat as a biofilm support (20). In the subsequent reactor study, two reactors were loaded with 1-2 mm peat granules and seeded with anaerobic sludge. One of the reactors was oxygenated using H2O2 (test reactor). The second reactor was maintained under anaerobic conditions throughout the experiment (control reactor). To establish adequate anaerobic colonization in the coupled reactor, a low H2O2 load of 0.35 g of O2 LR-1 day-1 was maintained throughout the first month of the experiment (phase 1). The load was then doubled to 0.7 g of O2 LR-1 day-1 and thus maintained throughout phases 2 and 3 (Table 1).

TABLE 2. Primers Used for PCR Amplification of Target Genetic Fragments

FIGURE 1. Methane yields in the presence (test reactor) and absence (control reactor) of hydrogen peroxide at different experimental phases. While the presence of hydrogen peroxide in the reactor containing natural anaerobic biomass granules affected the structural integrity of the granules, no biomass washout was observed when peat granules were used as a support. High porosity of peat provided a large surface area for bacterial attachment. It can be hypothesized that aerobic bacteria colonized the outer layer of peat granules, which was exposed to both carbon source (ethanol) and hydrogen peroxide, and the anaerobes colonized the inner, anaerobic parts of peat granules. Reactor monitoring showed the presence of methane in the off-gas of both test (coupled) and control (anaerobic) reactors throughout all experimental phases. The yield of methane throughout phase 1 in the test reactor was found to be 0.13-17 L of CH4 (g of COD)-1 on average, which was significantly lower than that of the control reactor (0.320.35 L of CH4 (g of COD)-1). Increased H2O2 loading in phases 2 and 3 resulted in an even lower methane yield (Figure 1). Nevertheless, the presence of methane in the off-gas of the coupled reactor over the entire course of the experiment indicated survival of methanogenic microorganisms. An additional confirmation of methanogenic survival in the coupled system was obtained through batch anaerobic tests of methane production under ethanol nonlimiting conditions. These tests were carried out at the end of each experimental phase. In agreement with the results of reactor monitoring, the tests showed methane production in test bottles inoculated both with the samples of test and control reactors (Figure 2a). Understandably, methane production rates were higher for the samples originating from the control reactor. Batch activity tests under aerobic methanotrophic conditions, i.e., using methane as a sole carbon source, were used to evaluate the presence and proliferation of methanotrophic bacteria. These tests showed significant methanotrophic activity in the coupled reactor by the end of phase 1, while methane consumption was undetectable in the control reactor at all times (Figure 2b). An increase in the hydrogen

FIGURE 2. Results of batch activity tests under methanogenic (a) and methanotrophic (b) conditions. peroxide feeding rate at phase 2 had no effect on the methanotrophic activity, which remained similar to that at phase 1. A comparison of the methanogenic and methanotrophic activities showed that despite reactor oxygenation, the methanogens were capable of producing methane in excess of the amount required for the methanotrophs. In the tests, the ratio of the methane production rate to methane consumption rate was 2:1 (Figure 2). Molecular analysis of microbial community DNAs provided additional confirmation of the presence and development of a methanogenic-methanotrophic microbial consortium in the coupled reactor. PCR amplification using 16S rRNA Archaeal probes showed the presence of methanogens in both reactors (Figure 3). Thus, the ability of methanogens to successfully colonize the peat substratum and exist in the presence of oxygen in the bulk liquid was confirmed. The presence of methanotrophic species was confirmed by PCR amplification of key enzymes of methanotrophs, namely, pmoA gene coding for a 27kDA polypeptide active site of particulate methane monoxygenase (pMMO) and the mxaF gene, which codes for the large subunit and active site of the methanol dehydrogenase (MDH) enzyme complex. Positive amplification of both pmOA and mxaF genes was observed in the samples originated from the test (coupled) VOL. 37, NO. 24, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 4. Concentrations of TCE and degradation products in the effluent and off-gas of the control and test reactors during the TCE-feeding phase (phase 3).

FIGURE 3. PCR amplification (a) and a densitometric diagram (b) showing PCR product intensities for methanotrophs, methylotrophs, and Archaea during different operational phases in control and test reactors. Notation: 1, 2, 3scontrol reactor, phases 1, 2, 3, respectively; 4, 5, 6stest reactor, phases 1, 2, 3, respectively; 7s inoculum (anaerobic sludge); 8speat. reactor, while no amplification was observed in the control reactor samples (Figure 3a). Since both pmoA and mxaF genes are found in all methanotrophs, they are reliable and useful indicators of methanotrophic presence. Although the pmoA gene is unique to methanotrophs, the mxaF gene is not since it is found in nonmethanotrophic methylotrophs and is the signature target gene for methylotrophic representation. Nevertheless, positive results obtained for both primers confirmed the presence of methanotrophs in the test reactor. The intensity of the methanotrophic and methylotrophic bands increased toward the end of the experiment (Figure 3b). The intensity of PCR bands cannot be used as a quantitative measure of population density. Even so, the increase in band intensity with time observed for both primers suggested an increase in density of these populations with a sustained presence during TCE feeding. A primer set targeting the mmoX gene of the sMMO enzyme complex yielded no positive signal, thus suggesting insignificant quantities of sMMO-positive methanotrophs in both reactors. Although all methanotrophs possess the particulate form of the MMO, only some may possess the soluble form. Expression of either form is known to be dependent on the amount of copper present in the medium. Considering the low levels of copper in the reactor effluent, it can be suggested that other factors, such as levels of oxygen and methane and the presence of significant quantities of heterotrophic bacteria, apparently resulted in the selection of pMMO-bearing methanotrophic populations. Traditionally, sMMO bearing methanotrophs have been associated with a high TCE degrading capacity and rapid oxidation rates for TCE and cis-DCE (21). Nevertheless, studies have shown that pMMO can achieve fairly high transformation yields for t-DCE, c-DCE, and vinyl chloride (22). Studies have also demonstrated that pMMO can mineralize TCE to CO2 by a degradative pathway, which differs from that of the sMMO, notably, production of the very toxic chloral hydrate is bypassed (23), thus improving TCE degradation. As mentioned above, specificity of the pmoA primer set used 5826

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in this study is not limited to methanotrophs since it also amplifies amoA gene sequences of the ammonia oxidizers (24). Due to the homology of these enzyme complexes, methanotrophs and ammonia oxidizing bacteria are similar in their ability to oxidize CH4 and NH4 as well as fortuitously co-metabolize various nongrowth substrates including chlorinated substrates such as TCE (25,26). Nitrifying bacteria may therefore also impact positively the success of the TCE degrading process as well as contribute to the overall dynamics of TCE, oxygen, and CH4 utilization within the coupled system. Furthermore, heterotrophic associations may contribute to the stability of the TCE degrading consortia by providing nutrients and removing byproducts which would otherwise be toxic. Overall, reactor conditions enabled the proliferation of a variety of autotrophic as well as heterotrophic microorganisms, which formed an efficient mutualistic consortium. Hydrogen peroxide was found to be an adequate source of oxygenation, offering improved oxygen delivery in comparison to the aeration approach. In the reactor, H2O2 was readily transformed to oxygen by catalases present in the heterotrophic community and oxygen was consumed by both heterotrophic and methanotrophic bacteria. Although aerobic heterotrophic bacteria consumed part of the ethanol, a sufficient amount was left for the metabolism of anaerobic microorganisms. Oxygen consumption by aerobic microorganisms created a steep oxygen gradient, which protected the anaerobes from oxygen toxicity. Furthermore, methane produced by methanogenic microorganisms was used as a source of carbon by methane consuming bacteria, namely, methanotrophs and possibly autotrophic nitrifying bacteria. TCE Degradation. The third phase of reactor operation was marked by the onset of TCE feeding at an influent concentration of 10 mg L-1 (76 µM). Off-gas analysis suggested no negative effect of TCE feeding on the methane yield of both reactors (Figure 1), although batch activity tests carried out at the end of this phase demonstrated a decline in the methanogenic activity of the control reactor (Figure 2a) and the methanotrophic activity of the test reactor (Figure 2b). Previous studies showed that chlorinated ethylenes had a potential toxic effect on both methanogenic and methanotrophic populations (27,28). Off-gas levels of TCE were found to be low in both reactors, although somewhat higher in the control reactor (Figure 4). No DCE was detected in the offgas of the control reactor, while trace amounts (0.2 µM) of 1,2-DCE were found in the off-gas of coupled reactor. Overall, volatilization losses of TCE and DCE were extremely low and remained below 0.1% of the amounts of TCE fed to the reactors. Measurements of TCE and DCE in the reactor effluent showed that upon TCE feeding, TCE in the effluent of the control reactor increased progressively throughout the

FIGURE 5. TCE mineralization rates obtained in batch tests. Abiotic, catalase, and heterotrophic tests were carried out in the presence of air. In addition, heterotrophic bottles contained ethanol. Phase 1-3 tests were carried out under methanotrophic conditions.

FIGURE 6. Concentrations of TCE and DCE in the reactor effluent during H2O2 shutdown experiment. experiment (Figure 4). In contrast, effluent concentrations of TCE in the coupled reactor remained at around 0.38 µM. Quantification of DCE peaks showed 3.1 µM 1,2-DCE in the effluent of the coupled reactor. No DCE was detected in the effluent of control reactor. The release of inorganic chloride, indicative of TCE biodegradation, was estimated by comparing effluent chloride levels at phases 2 (no TCE feeding) and 3 (with TCE feeding). The control reactor was found to have average chloride concentrations of 51 ( 14 and 73 ( 3 µM during phases 2 and 3, respectively. A significant release of inorganic chloride was observed in the coupled reactor with chloride levels increasing from 50 to 230 ( 40 µM between phases 2 and 3 (Figure 4). Using 50 µM as the background concentration, the amount of chloride produced due to TCE degradation in control and test reactors can be estimated to be 20 and 180 µM, respectively. To demonstrate the effect of H2O2 on the degradation sequence of TCE, H2O2 feeding to the coupled reactor was terminated 1 week prior to the end of phase 3. In response, an increase of 1,2-DCE from 3 to 34 µM in the reactor effluent was observed almost immediately (Figure 6). Mineralization tests carried out at the end of each experimental phase using biomass from both coupled and control reactors clearly demonstrated TCE mineralization under methanotrophic conditions, i.e., with adequate oxygen and methane present in the bottle headspace. TCE mineralization rates measured early in the experiment (at the end of phase 1) were found to be comparable in the control and coupled reactors (Figure 5). By the end of phase 2, TCE mineralization rate in the control reactor decreased and remained low at phase 3, while a 2-fold increase in the mineralization capacity of the coupled reactor by the end of phase 3 was noted (Figure 5).

Significant mineralization was only observed in test bottles inoculated with coupled reactor biomass and maintained under methanotrophic conditions. Mineralization tests in bottles supplemented with ethanol and in the absence of methane showed rates comparable to those of abiotic control bottles (Figure 5). Thus, mineralization activity due to aerobic heterotrophic metabolism was insignificant. In addition to the mineralization tests, batch tests of abiotic degradation of TCE and DCE were carried out in order to evaluate removal due to chemical transformations similar to those observed in the Fenton reaction (29), i.e., in the presence of hydrogen peroxide and metals. Only insignificant reduction in the levels of either component was observed (results not shown). It can be concluded that a combination of anaerobic and aerobic conditions in the coupled reactor not only allowed for the development of an efficient mutualistic system at the level of carbon source transformations, it also enhanced the capacity of the bacterial consortium to degrade TCE. Although both anaerobes and methanotrophs are capable of TCE degradation, the presence of 1,2-DCE in the coupled reactor suggests that the first step of TCE degradation was its reductive dechlorination to 1,2-DCE by anaerobic microorganisms (9,30). The dominance of this degradation pathway in the absence of oxygen was demonstrated in the hydrogen peroxide shutdown experiment (Figure 6). The absence of hydrogen peroxide almost immediately resulted in the appearance of elevated concentrations of 1,2-DCE, while the level of TCE in the effluent remained low. This observation agrees well with a number of publications, which reported accumulation of DCE in anaerobic degradation of TCE (31). Low levels of DCE occurred when H2O2 was fed to the reactor. This observation suggested that the dominant pathway of 1,2-DCE degradation was its oxidation by the methanotrophic bacteria to DCE epoxide followed by epoxide mineralization by heterotrophic microorganisms (32-34). Contrary to the amount of inorganic chloride released in the coupled reactor, only a slight increase in the chloride levels was observed in the control (anaerobic) reactor. While the input/output material balance of the anaerobic reactor showed a TCE removal of 81%, the inorganic chloride released in this reactor accounted for between 9% and 30% of TCE disappearance, depending on the assumed extent of dechlorination (i.e., transformation to DCE or complete dechlorination). This means that only a small part of TCE removal in the anaerobic reactor could be attributed to TCE dechlorination, whereas most of the TCE removal was likely due to TCE adsorption onto the solid support. Reductive dechlorination of TCE under anaerobic conditions is well documented (9,30). Electron donor availability is critical for the dechlorination process as it will dictate how and to what extent dechlorination will proceed. Some authors (32,35) showed dechlorination rates were low when using ethanol as compared to other sources of carbon since it is a good source of hydrogen. Therefore, under strict anaerobic conditions, dechlorinating microorganisms could have been outcompeted by hydrogen-consuming methanogenic and sulfidogenic populations since hydrogen is more readily available. In contrast, in the coupled reactor a significant part of the ethanol was metabolized by the aerobic heterotrophs, which limited the fermentation of ethanol and as a result the amount of hydrogen released. In addition, differences in the anaerobic populations evolving from the same inoculum can also be considered as an explanation. In conclusion, TCE degradation in a coupled anaerobic/ aerobic reactor oxygenated using hydrogen peroxide provided a number of advantages over existing biodegradation processes. First of all, the presence of methanogenic species in the reactor allowed for an in-situ production of methane, which was used by the methanotrophs, thus eliminating the VOL. 37, NO. 24, 2003 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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need for methane delivery and simplifying logistics of reactor operation. Moreover, reactor oxygenation with hydrogen peroxide eliminated volatilization losses of TCE. A combination of reductive and oxidative pathways of TCE degradation reduced the influence of toxic degradation intermediates, such as DCEs, on the overall degradation rate and provided an almost complete mineralization of TCE.

Acknowledgments The authors acknowledge the insightful scientific discussions with Dr. Darwin Lyew, which contributed to this study. This is NRC Paper No. 45928.

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Received for review January 28, 2003. Revised manuscript received August 12, 2003. Accepted October 6, 2003. ES030340V