Coupling Permanganate Oxidation With Microbial Dechlorination of Tetrachloroethene
By Sahl, Jason W; Munakata-Marr, Junko; Crimi, Michelle L; Siegrist, Robert L
For sites contaminated with chloroethene non-aqueous-phase liquids, designing a remediation system that couples in situ chemical oxidation (ISCO) with potassium permanganate (KMnO^sub 4^) and microbial dechlorination may be complicated because of the potentially adverse effects of ISCO on anaerobic bioremediation processes. Therefore, one-dimensional column studies were conducted to understand the effect of permanganate oxidation on tetrachloroethene (PCE) dechlorination by the anaerobic mixed culture KB-1. Following the confirmation of PCE dechlorination, KMnO^sub 4^ was applied to all columns at a range of concentrations and application velocities to simulate varied distances from oxidant injection. Immediately following oxidation, reductive dechlorination was inhibited; however, after passing several pore volumes of sterile growth medium through the columns after oxidation, a rebound of PCE dechlorination activity was observed in every inoculated column without the need to reinoculate. The volume of medium required for a rebound of dechlorination activity differed from 1.1 to 8.1 pore volumes (at a groundwater velocity of 4 cm/d), depending on the specific condition of oxidant application. Water Environ. Res., 79, 5 (2007).
KEYWORDS: in situ chemical oxidation, dechlorination, coupling, bioremediation.
Chlorinated ethenes, such as tetrachloroethene (PCE) and trichloroethene (TCE), are common contaminants found in groundwater systems. These contaminants can exist in the subsurface primarily in the following three phases: (1) dissolved-contaminant phase, (2) sorbed phase, and (3) non-aqueous-liquid phase (NAPL). As chloroethene NAPLs are released into the subsurface, their dense nature causes them to sink in saturated subsurface zones, forming NAPL source zones at the bottom of aquifers or on lenses of low permeability media. Because many chloroethenes have low absolute solubilities, NAPL source zones can take tens to hundreds of years to fully dissolve at typical groundwater velocities, thereby creating a long-term source of groundwater contamination (Johnson and Pankow, 1992). Furthermore, most chloroethenes are known or suspected carcinogens with regulatory limits lower than their solubility limits. The relatively insoluble, yet toxic, nature of these contaminants has prompted efforts to develop remediation systems designed to reduce the longevity and contaminant flux from NAPL source zones (Kavanaugh et al., 2003).
In situ chemical oxidation (ISCO) using potassium permanganate (KMnO^sub 4^) has been shown to be effective at treating chloroethenecontaminated soil and groundwater (Schnarr et al., 1998; Siegrist et al., 1999; Yan and Schwartz, 1999). Permanganate destroys chloroethenes, such as PCE (C^sub 2^C^sub 14^), by direct electron transfer mechanisms (Wiberg and Saegebarth, 1957; Yan and Schwartz, 1999) in the aqueous phase, resulting in the production of manganese dioxide solids [MnO^sub 2^(S)], carbon dioxide, and free ions, as follows:
3C^sub 2^Cl^sub 4^ + 4KMnO^sub 4^ + 4H^sub 2^O [arrow right] 12Cl^sup -^ + 6CO^sub 2^ + 4MnO^sub 2^ (s) + 4K^sup +^+8H^sup +^ (1)
Factors limiting the effectiveness of permanganate oxidation of chloroethene-contaminated sites are bypassing of oxidant around NAPL source zones because of soil heterogeneities, reduced permeability in areas containing NAPL, and generation of permanganate oxidation byproducts manganese dioxide (Li and Schwartz, 2000; Lee et al., 2003) and carbon dioxide (Schnarr et al., 1998). Furthermore, the formation of manganese dioxide on the NAPL-water interface can potentially decrease the mass transfer of contaminant out of the NAPL phase (MacKinnon and Thomson, 2002), limiting the contact and reaction with the oxidant. As a result of these limitations, residual NAPL may persist following the cessation of active ISCO, posing a continued threat to groundwater quality. The implementation of a secondary technology, such as bioremediation, following oxidation may be a cost-effective method to further remediate the site and reduce groundwater contaminant concentrations.
Anaerobic bioremediation of chloroethene contaminants has been well-documented in the laboratory (Duhamel et al., 2002; Fennell and Gossett, 1998; Maym-Gatell, Chien, Gossett, and Zinder, 1997; Maym- Gatell, Tandoi, Gossett, and Zinder, 1995) and in field-scale conditions (Hohnstock-Ashe et al., 2001; Major et al., 2002; Yager et al., 1997). The PCE can be completely dechlorinated to ethene by organisms capable of dehalorespiration (MaymoGatell, Chien, Gossett, and Zinder, 1997; Maymo-Gatell, Tandoi, Gossett, and Zinder, 1995), potentially increasing the rate of NAPL dissolution (Carr et al., 2000; Cope and Hughes, 2001; Gupta and Seagren, 2005; Seagren et al., 1994; Yang and McCarty, 2000, 2002), by reducing PCE to byproducts that more easily partition into the aqueous phase. However, accumulation of PCE degradation byproducts, such as TCE, cis-1,2-dichloroethene (cis-DCE), and vinyl chloride, has been observed; such accumulation has been attributed to the lack of appropriate organisms, reducing conditions, and electron donors (Maymo-Gatell, Chien, Gossett, and Zinder, 1997).
In coupling ISCO with bioremediation, ISCO using KMnO^sub 4^ may decrease microbial activity by lowering the pH (Kastner et al., 2000), changing subsurface oxidation-reduction (redox) conditions (Klens et al., 2001), generating moderate oxidants [i.e., MnO^sub 2^(s)], or directly oxidizing microbial DNA (Bui and Cotton, 2002). The effect of ISCO on biological processes has been explored in the laboratory and in field applications using catalyzed hydrogen peroxide (Alien and Reardon, 2000; Kastner et al., 2000; Miller et al., 1996) and permanganate (Droste et al., 2002; Gardner et al., 1996; Hazen et al., 2000; Hrapovic et al., 2005; Klens et al., 2001; Macbeth et al., 2005), to determine if the two processes can be coupled in the remediation of a variety of contaminants. A rebound of biological activity and/or biomass was observed in all of these studies, following the application of various oxidants at different concentrations. However, only two of these studies focused specifically on the effect of potassium permanganate on reductive dechlorination processes in the remediation of chloroethene NAPLs. One of the studies focused on bioaugmentation following permanganate application (Hrapovic et al., 2005), while the other study was conducted at a field site and did not quantify the extent of TCE dechlorination (Droste et al., 2002).
This study was performed to determine the effects of ISCO with potassium permanganate at a range of oxidant masses, concentrations, and application velocities on the reductive dechlorination of PCE in controlled laboratory experiments without reinoculation. The columns were not reinoculated following oxidation to determine if the original culture was sterilized by the oxidant application or could resume dechlorination activity once preoxidation conditions were reestablished. Using glass columns containing porous media and a microbial culture capable of PCE dechlorination, studies were performed to determine the effect of permanganate oxidation on PCE dechlorination activity. Pre- and post-oxidation PCE biodegradation activity was compared to determine if (1) reductive dechlorination activity could resume following oxidation without reinoculation, and (2) specific experimental parameters, such as oxidant concentration, mass, and application velocity, affected subsequent microbial reduction of PCE. A range of oxidation parameters were used to determine how organisms exposed to different oxidation conditions experienced during a pressurized oxidant flush are affected based on their proximity to an oxidant injection well.
Materials and Methods
Chemicals and Reagents. The following chemicals and reagents were used: hexane (Mallinckrodt-HPLC) (Mallinckrodt Inc., St Louis, Missouri), methanol (Aldrich-HPLC) (SigmaAldrich, St. Louis, Missouri), PCE (Mallinckrodt-99%), TCE (Mallinckrodt-99%), cis-DCE (Aldrich-97%), KMnO^sub 4^ (Carustechnical grade) (Carus Corporation, Peru, Illinois), vinyl chloride (Aldrich-99.5+ %), and ethene (Aldrich-99.5+ %). All PCR reagents (magnesium chloride, deoxynucleotide triphosphates, 10 buffer, and Taq polymerase) were obtained from Promega Corporation (Madison, Wisconsin).
Culture Growth and Maintenance. The KB-I culture, an anaerobic culture capable of dehalorespiration, was anaerobically enriched from soil and groundwater obtained from a Southern Ontario TCE- contaminated site and has been used in bioaugmentation applications (Major et al., 2002). Two batches of KB-I were grown in 1-L Pyrex bottles (Pyrex, Charleroi, Pennsylvania) in an anaerobic glove box using a nutrient medium amended with methanol (2.0 mM) and aqueous- phase PCE (0.15 mM). The nutrient medium is a buffered solution consisting of the following constituents added per liter of MiIH-Q water (Millipore, Billerica, Massachusetts) (Edwards and Grbic- Galic, 1994): 10 mL phosphate buffer (27.2 g potassium phosphate monobasic and 34.8 g potassium phosphate dibasic per liter), 2 mL trace mineral solution (0.3 g ortho\boric acid, 0.1 g zinc chloride, 0.75 g nickel chloride hexahydrate, 1.0 g manganese chloride tetrahydrate, 0.1 g copper chloride dihydrate, 1.5 g cobalt chloride hexahydrate, 0.02 g sodium selenite, 0.1 g aluminum sulfate, 1 mL sulfuric acid, 10 mL salt solution (53.5 g ammonium chloride, 7.0 g calcium chloride hexahydrate, and 2.0 g iron chloride tetrahydrate per liter), 2 mL magnesium sulfate solution (62.5 g/L), 1 mL resazurin (1 g/L), 10 mL saturated bicarbonate solution (200 g sodium bicarbonate per liter), 10 mL filter-sterilized vitamin stock solution (0.02 g biotin, 0.02 g folie acid, 0.1 pyridoxine hydrochloride, 0.05 g nicotinic acid, 0.05 pan to then ic acid, 0.05 g p-amino-benzoic acid, 0.05 cyanocobalamin, 0.05 g thioctic acid, and 1 g mercaptoethanesulfonic acid per liter), and 10 mL amorphous ferrous sulfide solution (39.2 g ferrous ammonium sulfate and 24.0 g sodium sulfide). The ingredients were combined, autoclaved, and stored in an anaerobic chamber.
The cultures differed in rates of chloroethene dehalorespiration, ranging from 1 to 3 weeks to completely dechlorinate the PCE feed amendment, although the variation between cultures occurred primarily in the conversion of vinyl chloride to ethene. In each run of the column study, the columns were inoculated with culture from one of the two batches. A similar optical density (0.10 to 0.12 AU/ cm) was recorded for each batch before column inoculation, but these measurements were confounded by the presence of sulfide solids in the medium. Each culture was inoculated in the columns after the batch PCE had been completely dechlorinated to ethene.
Experimental Design. One-dimensional, flow-through experiments were conducted in six glass columns (25 cm long, 3 cm internal diameter). Teflon fittings and Viton tubing were used in the experiments for chemical compatibility with all reagents. A Swagelok (Swagelok Corporation, Solon, Ohio) three-way valve was placed at the effluent (top) end of each column, where flow could be directed either into a 1-mL glass sample syringe or into a waste container. The columns were wet-packed with the nutrient medium in an anaerobic glove box. Sand was added in 1-cm lifts with a constant head of 2 cm and packed tight after each addition of sand. This sand addition resulted in a 3.5-cm coarse silica sand (#16) source zone overlaid by a fine silica sand (#70) matrix; porous media properties are summarized in Table 1. The coarse sand zone surrounded by a fine matrix was designed to create a capillary barrier preventing the migration of the NAPL from the source zone into the upper sections of the column.
Column pore volume was determined gravimetrically by subtracting the porous media mass from the total column weight. Dividing the pore volume mass by the density of water yielded the column pore water volume. The pore volume varied between columns, with an average calculated pore-space volume of 64.6 mL (standard deviation 2.1 mL). A bromide tracer test was performed on three of the columns to verify the accuracy of the pore volume calculation. One-tenth of one pore volume of potassium bromide solution was pumped through each column, with an initial bromide ion concentration of 12.5 mM, at a rate of 1 mL/min. Effluent samples were taken and analyzed on a Dionex DX-300 ion chromatograph with an ASIl-HC Ion Pac Column (Dionex Corporation, Sunnyvale, California) for bromide concentrations. The bromide concentrations were then analyzed by the method of moments to calculate a pore volume. The pore volume obtained by the tracer test corresponded with the gravimetrically determined pore volume (95% confidence), indicating that no bypassing of flow was occurring.
Five columns were inoculated, while one uninoculated column acted as a control. After the columns were packed, all inoculated columns were amended with 2 pore volumes of active KB-1 in the nutrient medium amended with methanol (0.13 mM) and resazurin, a visual indicator of low redox potential conditions. The culture was pumped into each column upflow for 6 days, using a syringe pump at a slow delivery velocity (4 cm/d). Columns LL-1, LL-2, and MM were inoculated using one batch of KB-1, while columns LH and HH were inoculated with a different batch; this second batch completely dechlorinated PCE more rapidly than the batch used to inoculate LL- 1, LL-2, and MM.
After inoculation, pure-phase PCE (1 mL) dyed with Sudan IV (0.05 M), a visual indicator, was then slowly injected, with a 1 mL glass syringe, to the bottom of the column to fill 15% of the coarselens pore space. After emplacing PCE, nutrient medium with methanol and resazurin amendment was then pumped upflow into each column with a syringe pump at a Darcy velocity representative of ambient groundwater flow (4 cm/d or approximately 0.4 pore volumes/d). Column effluent was monitored for PCE dechlorination before oxidant application. Chloroethene concentrations in column effluent were monitored following application of the first pore volume of nutrient medium, during which any residual dechlorination byproducts from the inoculum should have been flushed from the columns.
After establishing reductive dechlorination in inoculated columns, permanganate was applied to all of the columns. Based on their proximity to the oxidant injection well, microbes in the subsurface may be exposed to a range of oxidant concentrations and application velocities. The columns therefore were subjected to oxidation at different oxidant concentrations and Darcy velocities of oxidant application (each qualitatively labeled low [L], medium [M], or high [H]) (Table 2). The mass of oxidant injected was chosen on a stoichiometric basis to destroy between 6% (condition LL) and 25% (condition HH) of the residual NAPL in the source zone. A percentage of PCE mass destruction was chosen to leave significant NAPL to generate aqueous PCE contamination following oxidation. The uninoculated control column (MM-C) was oxidized with a medium concentration of permanganate at a medium application velocity, but was not inoculated with KB-1. Permanganate was pumped through columns with a syringe pump programmed to the designated flowrate to deliver a predetermined mass of oxidant. Following oxidation with permanganate, feeding with nutrient medium immediately resumed at the initial application velocity (4 cm/d) in all columns. Samples were taken from the column effluent and analyzed for postoxidation PCE dechlorination activity.
Analytical Methods. Column effluent samples were collected in a 1- mL glass syringe. Then, 10-L aqueous samples were extracted in 1 mL hexane and analyzed on a HP 6890 gas chromatograph equipped with an electron capture detector and a HP624 special analysis column (Hewlett-Packard Inc., Palo Alto, California) (30 m 530 m 3.0 m) for PCE, TCE, and cis-DCE. Standards were made volumetrically in hexane, and a calibration curve was generated with five points that encompassed all of the anticipated concentrations.
For the analysis of vinyl chloride and ethene, a 1-mL aqueous sample was collected and directly injected to a 2-mL glass vial topped with a Mininert valve. The vial contents were allowed to equilibrate for 2 hours, and then a 100-L headspace sample was manually injected to the gas chromatograph equipped with a flame ionization detector. Gas-phase concentrations were converted to aqueous concentrations using dimensionless Henry’s constant values of 1.14 for vinyl chloride (Gossett, 1987) and 7.75 for ethene (Carr and Hughes, 1998), at 25C. Standards were created by injecting gas- phase vinyl chloride and ethene to 60-mL sealed serum bottles. A calibration curve was subsequently generated using a minimum of four calibration points.
Concentrations of dissolved manganese in the column effluent were analyzed by inductively coupled plasma (ICP) optical emission spectroscopy. Manganese dioxide concentrations in the columns were also determined by ICP analysis. Following the column study, four columns (LL-1, LL-2, MM, and LH) were dissected into five 5-cm sections (labeled A to E, from bottom to top). Approximately 1 g of soil from each section was then rinsed with deionized water and barium chloride (0.1 M) to extract dissolved manganese, then the remaining manganese was extracted with hydroxylamine hydrochloride (0.1 M) containing nitric acid (0.01 M) (Struse et al., 2002).
DNA Extraction and Amplification. DNA was extracted from a soil sample (0.22 g) in each dissected section of columns LH and HH using the Mobio Powersoil DNA extraction kit. Additionally, a 50-mL sample of active KB-1 was spun down, and the DNA was extracted to be used as a positive control for the PCR reaction; DNA extracted from Escherichia coli was used as a negative control. Following extraction, two primers, Fp DHC 1 (5′-GATGAACGCTAGCGGCG-3′) and 1386r (5-CCTCCTTGCGGTTGGCACATC3′) (Duhamel et al., 2004; Hendrickson et al. 2002), designed to target Dehalococcoides (DHC)-specific 16S rRNA small subunit gene fragments were used in a polymerase chain reaction (PCR) designed to confirm the presence of DHC within each dissected section (A to E) of sand. The PCR thermocycling program and reaction matrix was identical to reported protocols (Duhamel et al., 2004) designed to target DHC gene sequences in KB-1, with products verified by agarose gel electrophoresis. This procedure was used to determine the presence, but not the quantity or viability, of DHC within each section of the columns.
Chloroethene Concentrations. During the preoxidation phase of the experiment, in which all columns were operated under the same conditions, the total chloroethene concentrations in the effluent gradually increased in all columns. For example, in column LH, the total chloroethene concentration rose from 0.45 to 1.6 mM during the preoxidation phase, in which approximately 4 pore volumes of growth media were applied (Figure 1). The PCE d\echlorination activity was verified by the accumulation of the dechlorination daughter products TCE and cis-DCE; vinyl chloride and ethene were not detected during the preoxidation phase of the experiment. Statistically significant differences (95% confidence) in total chloroethene concentrations were observed between some columns in preoxidation conditions (Figure 2, labeled “pre”), indicating initial variability in culture activity or differential contact of the growth medium with the NAPL.
During the oxidation phase of the experiment, the dechlorination byproducts TCE and cis-DCE were not observed. Aqueous-phase PCE concentrations did not significantly drop in any column, indicating that the permanganate-PCE reaction may have been limited by permanganate reaction with other targets (i.e., iron sulfide in the nutrient medium or microbial biomass). Other indications of permanganate reaction with chloroethenes were observed, such as pH depression in all columns during oxidant application, consistent with the generation of free protons in the permanganate reaction with PCE (eq 1). Furthermore, the evolution of gas bubbles (likely carbon dioxide from the permanganate-PCE reaction) and the deposition of MnO^sub 2^(S) were observed in all columns following oxidation, indicating permanganate reaction.
Following oxidation, the pH in all columns returned to preoxidation levels after 1 pore volume of the phosphate-buffered nutrient medium was applied. Immediately following the permanganate flush, PCE dechlorination daughter products remained absent in the column effluent of all inoculated columns. However, daughter products were eventually detected in all inoculated columns following a period of inactivity. The period of inactivity differed between columns run under different operational conditions. The volume of medium applied before an observed rebound of dechlorination byproduct formation in each column, pH change, and hours of oxidant application are listed in Table 3.
Average steady-state total chloroethene concentrations for each column during each phase of the experiment are summarized in Figure 2. A 95% confidence interval range was calculated from all time points during each phase of the experiment once a steady-state concentration was observed. Furthermore, the concentration ratios of cis-DCE to PCE during the pre- and post-oxidation phases of the experiment are also presented in Figure 2 to provide a sense of the extent of dechlorination activity in each column. Oxidation data are not presented for column HH, because too few samples were analyzed to accurately calculate a steady-state chloroethene concentration.
Manganese Dioxide Deposition. Manganese dioxide deposition was clearly visible in all columns, the extent of which depended directly on the concentration and velocity of oxidant application. Columns with a high oxidant application velocity (LH and HH) showed MnO^sub 2^(s) deposition throughout the entire length of the column, with visual permanganate breakthrough in the column effluent. In columns with a high oxidant application velocity, the shorter column retention time limited permanganate reaction in the NAPL source zone, allowing the permanganate solution to travel further in the column. Conversely, MnO^sub 2^(s) deposition was limited to the bottom one-half of the column, with a lower oxidation application velocity (LL-1, LL-2, MM, and MM-C), with no permanganate ever visually observed in column effluent.
Manganese dioxide concentrations for each 5-cm section (A to E) in the dissected columns are shown in Figure 3. The measured concentrations quantify the visual observations described above. The average concentration of MnO^sub 2^(s) was highest in the two columns that received a greater mass of KMnO^sub 4^ (MM and LH) and lowest in the columns that received a lower KMnO^sub 4^ mass (LL-1 and LL-2). In column HH, complete reduction of manganese dioxide solids was visually observed during the postoxidation phase; therefore, MnO^sub 2^(S) concentrations were not measured. Furthermore, MnO^sub 2^(s) concentrations were not quantified in the control column (MM-C), because investigating the relationship between MnO^sub 2^(S) concentrations and bioactivity was the primary goal of MnO^sub 2^(S) quantification.
Concentrations of dissolved manganese were only elevated over the nutrient medium concentration (2.2 M) in column HH following oxidation. Once MnO^sub 2^(s) was visually absent from column HH, concentrations of dissolved manganese decreased to levels observed in the uninoculated control. The highest manganese concentration observed in the column effluent of column HH was 6.0 M.
Identification of Dehalococcoides. The PCR products were observed in sections A to E of columns HH and LH (data not shown), confirming the presence of DHC. The positive control also amplified with the same product size as column extracts, while the negative control showed no amplification. The PCR product in each DHC-positive sample was between 1000 and 1500 kb, consistent with the product size selected for by primers Fp DHCl and 1386r. The positive confirmation of DHC in both columns was anticipated because of postoxidation dechlorination activity. While this method is not quantitative and does not distinguish between active and inactive cells, the detection of DHC throughout both columns indicates that the KB-I culture was distributed throughout the columns.
Preoxidation average total chloroethene concentrations and cis- DCE:PCE ratios varied between some column runs, likely because of the use of different inocula or differences in nutrient medium contact with the NAPL. The rate of PCE dechlorination in each batch was not determined at the point of inoculation, but differences in the rate of dechlorination, especially in the dechlorination of vinyl chloride, existed between the column inocula. However, total chloroethene concentrations and cis-DCE:PCE ratios varied between duplicate columns (LL-I and LL-2) inoculated with the same batch, indicating some variability in NAPL-nutrient medium contact.
During oxidation, no dechlorination daughter products were detected, suggesting that reductive dechlorination ceased. Additionally, aqueous PCE concentrations in column effluent during the oxidation phase did not significantly differ between experimental conditions. Once PCE dissolved into the medium in the source zone, insufficient KMnO^sub 4^ reaction with the aqueous PCE resulted in PCE presence in the effluent of all columns. For example, permanganate reaction with PCE in columns run under low oxidant concentration was likely stoichiometrically limited by low permanganate concentrations, as evidenced by the lack of permanganate in the effluent; only 0.57 moles of KMnO^sub 4^ were provided per mole of aqueous (saturated) PCE, compared with 1.3 moles KMnO^sub 4^ required for complete oxidation of aqueous PCE (eq 1). In column HH, run under high application velocity with 57 moles of KMnO^sub 4^ per mole of aqueous PCE, PCE persisted in the effluent concurrently with KMnO^sub 4^; the limited reaction may have been a result of the short column detention time (1.8 hours versus 19 or 28 hours for low or medium velocities). Sufficient KMnO^sub 4^ (5.7 moles per mole of aqueous PCE) was added to column MM to oxidize the aqueous PCE, but it was consumed either by reaction in the NAPL source zone or by reaction with competing oxidation targets (i.e., iron disulfide in growth medium), as it was not detected in the column effluent.
Following the permanganate flush, a relatively rapid rebound of significant PCE dechlorination activity was observed in every inoculated column. These column studies used a nutrient medium to promote preoxidation PCE dechlorination activity and the rebound of bioactivity postoxidation. The columns were not reinoculated following oxidation, indicating that the original culture rebounded following oxidant application. Although other studies have reported the rebound of biological activity following permanganate application (Azadpour-Keeley et al., 2004; Hazen et al., 2000; Klens et al., 2001;), only one other study (Hrapovic et al., 2005) showed PCE dechlorination following oxidation. In the other study, only one column showed dechlorination activity after flushing for over 2 months with site groundwater containing dehalorespiring organisms and biostimulating with acetate and ethanol; dechlorination was observed after a significant lag time (222 days after oxidation), and concentrations of dechlorination byproducts were very low (0.24 M).
A rebound of microbial biomass and/or bioremediation activity to preoxidation levels following ISCO has been observed in other systems and has been attributed to bioavailable oxidation by- products (Azadpour-Keeley et al., 2004; Droste et al., 2002;), the decrease in toxicity associated with contaminant mass removal (Chapelle et al., 2005; Miller et al., 1996), or selection for micro- organisms capable of bioremediation achieved by reduced substrate competition following oxidant application (Miller et al., 1996). Some research has shown an increase in biomass and decrease in diversity following permanganate application (Macbeth et al., 2005; Miller et al., 1996), which could result in reduced substrate competition for microorganisms capable of bioremediation.
In four of the five inoculated columns, the postoxidation concentration of total chloroethenes was statistically greater (Figure 2) than the saturated concentration of aqueous-phase PCE (approximately 1.10 mM). This indicates that enhanced dissolution of PCE occurred concomitantly with biological dechlorination near the PCE source zone, consistent with other findings (Carr et al, 2000; Cope and Hughes, 2001; Yang and McCarty, 2000). The enhanced mass transfer of chloroethenes postoxidation could potentially decrease the longevity of mass-transfer-limited chloroethene \NAPLs. However, downgradient aqueous concentrations could be elevated because of enhanced mass transfer, which would need to be considered when designing remediation systems.
The PCE and ci’s-DCE were the primary chloroethenes detected in column effluent throughout the experiment. The TCE appeared to be a transient dechlorination byproduct, with concentrations never exceeding 0.038 mM. Vinyl chloride and ethene were detected in the effluent samples of three columns (LL-I, LL-2 and HH), but only after applying 5 pore volumes of media postoxidation and then in low concentrations (
The spatial deposition of manganese dioxides in these studies was highly dependent on the oxidant application velocity and concentration (Figure 3). Column LH showed the most rapid rebound of PCE dechlorination activity and had the highest average concentration of MnO^sub 2^(s) distributed throughout the column. The MnO^sub 2^(S) is a moderate oxidant (Xie and Barcelona, 2003), and the rebound of anaerobic activity in the presence of these solids is surprising. However, the presence of bioavailable manganese [Mn(IV)] has been shown to induce the mineralization of cis-DCE in manganese-reducing conditions (Bradley et al., 1998), demonstrating that anaerobic processes may proceed in the presence of an oxidant.
Dissolution of MnO^sub 2^(s) was visibly evident, and elevated effluent concentrations of dissolved manganese were observed in column HH immediately following the rebound of postoxidation reductive dechlorination; MnO^sub 2^(S) dissolution appeared complete following 3 pore volumes of media applied. It is unclear why MnO^sub 2^(S) dissolution was only visible in column HH, but it may have been a result of a lower redox potential than was present in other columns. Manganese dioxide reduction typically results from a drop in redox potential because of bacterial metabolism (Gounot, 1994). The redox potential was not monitored in these studies, so the relationship between reductive dechlorination activity and manganese reduction could not be determined. The reduction of manganese dioxides was reported in another laboratory study using dechlorinating organisms (Hrapovic et al., 2005), but a direct link between reductive dechlorination activity and manganese reduction has not been established. However, the reduction of Mn(IV) has been associated with increased DCE mineralization (Bradley et al., 1998), suggesting a potential relationship between anaerobic remediation processes and manganese reduction.
Dechlorinating microbes clearly survived permanganate exposure, given the dechlorination byproducts observed. Dehalococcoides was detected throughout columns LH and HH, indicating that these microbes were distributed throughout the columns. Whether their activity was localized or distributed is unknown, because PCR does not discriminate between live and dead organisms. Permanganate is known to have disinfectant properties; coliform bacteria in water were completely inactivated within 2 hours at doses up to 6 mg/L KMnO^sub 4^, though some pathogenic bacteria remained after 24 hours exposure to 20 mg/L KMnO^sub 4^ (U.S. EPA, 1999). However, KMnO^sub 4^ has been shown to oxidize cellular DNA (Bui and Cotton, 2002), which would likely not generate subsequent PCR product. Other studies have also reported the persistence of DHC following direct exposure to KMnO^sub 4^ (Hrapovic et al, 2005; Macbeth et al., 2005), suggesting that DHC can be resilient to permanganate oxidation.
Results from these anaerobic column studies demonstrate that biological remediation activity may diminish during potassium permanganate application, but can recover without reinoculation. As permanganate is applied in field situations, an ideal condition of permanganate application may exist for coupling with biological remediation processes. Based on the results of this work, the most rapid rebound of reductive dechlorination activity occurred using a low concentration of oxidant and a large volume of oxidant applied at a high application velocity. In field applications of permanganate, increased volumes of permanganate injection would likely be associated with an increased remediation expense. However, higher remediation costs associated with a larger injection volume may potentially be offset by improved efficiency of posttreatment bioprocesses. Furthermore, other work has determined that a low oxidant concentration coupled with an increased oxidant application velocity may be more efficient in treating NAPLs in soils with a high natural oxidant demand (NOD) (Crimi and Siegrist, 2004). Therefore, designing remediation systems that use a lower permanganate concentration at a high velocity may limit adverse effects of ISCO on native microbes and minimize unproductive oxidant consumption by NOD.
The results from this study suggest the following:
(1) Though PCE reductive dechlorination activity may stop during permanganate application, it can resume following oxidant application without reinoculation;
(2) PCE mass transfer can be increased by the activity of dechlorinating organisms living near NAPL source zones; and
(3) Dechlorinating organisms can actively dechlorinate PCE in the presence of manganese dioxide.
The authors thank Elizabeth Edwards (University of Toronto, Canada) and Melanie Duhamel (GeoSyntec Consultants, Canada) for providing the KB-I culture, Ann Kaplan (Geomega Inc., Boulder, Colorado) for propagating the culture for this study, Lauren Nagy (Colorado School of Mines, Golden) for guidance with PCR, and Jimmy Kopp (Colorado School of Mines) for ICP analysis. This work was supported by the Strategic Environmental Research and Development Program (Arlington, Virginia) (CU-1290).
Submitted for publication February 16, 2006; revised manuscript submitted May 25, 2006; accepted for publication July 25, 2006.
The deadline to submit Discussions of this paper is April 15, 2007.
Adamson, D. T.; McDade, J. M.; Hughes, J. B. (2003) Inoculation of a DNAPL Source Zone to Initiate Reductive Dechlorination of PCE. Environ. Sci. Technol., 37, 2525-2533.
Allen, S. A.; Reardon, K. F. (2000) Remediation of Contaminated Soils by Combined Chemical and Biological Treatments. In Physical and Thermal Technologies: Remediation of Chlorinated and Recalcitrant Compounds, Wickamanayake, G. B., Gavaskar, A. R. (Eds.); Battelle Press: Columbus, Ohio.
Azadpour-Keeley, A.; Wood, L. A.; Ue, T. R.; Mravik, S. C. (2004) Microbial Responses to In Situ Chemical Oxidation, Six-Phase Heating, and Steam Injection Remediation Technologies in Groundwater. Remediation, 14 (4), 5-17.
Bradley, P. M.; Landmeyer, J. E.; Dinicola, R. S. (1998) Anaerobic Oxidation of [1,2-14^sup C^] Dichloroethene under Mn (IV)- Reducing Conditions. Appl. Envrion. Micmbiol., 64 (4), 1560-1562.
Bui, C. T.; Cotton, R. G. H. (2002) Comparative Study of Permanganate Oxidation Reactions of Nucleotide Bases by Spectroscopy. Bioorg. Chem., 30 (2), 133-137.
Carr, C. S.; Garg, S.; Hughes, J. B. (2000) Effect of Dechlorinating Bacteria on the Longevity and Composition of PCE- Containing Nonaqueous Phase Liquids under Equilibrium Dissolution Conditions. Environ. Sd. Technol., 34, 1088-1094.
Carr, C. S.; Hughes, J. B. (1998) Enrichment of High-Rate PCE Dechlorination and Comparative Study of Lactate, Methanol, and Hydrogen as Electron Donors to Sustain Activity. Environ. Sd. Technol., 32,1817-1824.
Chapelle, F. H.; Bradley, P. M.; casey, C. C. (2005) Behavior of a Chlorinated Ethene Plume Following Source-Area Treatment with Fenton’s Reagent. Ground Water Monit. Remed., 25 (2), 131-141.
Cope, N.; Hughes, J. B. (2001) Biologically-Enhanced Removal of PCE from NAPL Source Zones. Environ. Sd. Technol., 35, 2014-2021.
Crimi, M. L.; Siegrist, R. L. (2004) Technical Report: Experimental Evaluation of In Situ Chemical Oxidation Activities at the Naval Training Center (NTC) Site, Orlando, Florida. Colorado School of Mines: Golden, Colorado.
Droste, E. X.; Marley, M. C.; Parikh, J. M.; Lee, A. M. (2002) Observed Enhanced Reductive Dechlorination After In Situ Chemical Oxidation Pilot Test. In Remediation of Chlorinated and Recalcitrant Compounds; Battelle Press: Monterey, California.
Duhamel, M.; Mo, K.; Edwards, E. A. (2004) Characterization of a Highly Enriched De/iafococcojWes-Containing Culture that Grows on Vinyl Chloride and Trichloroethene. Appl. Environ. Microbiol., 70, 55385545.
Duhamel, M.; Wehr, S. D.; Yu, L.; Rizvi, H.; seepersad, D.; Dworatzek, S.; Cox, E. E.; Edwards, E. A. (2002) Comparison of Anaerobic Dechlorinating Enrichment Cultures Maintained on Tetrachloroethene, Trichloroethene, c/s-dichloroethene and Vinyl Chloride. Water Res., 36, 4193-4202.
Edwards, E. A.; Grbic-Galic, D. (1994) Anaerobic Degradation of Toluene and o-xylene by a Methanogenic Consortium. Appl. Environ. Microbiol., 60 (1), 313-322.
Fennell, D. E.; Gossett, J. M. (1998) Modeling the Production of and Competition for Hydrogen in a Dechlorinating Culture. Environ. Sd. Technol., 32, 2450-2460.
Gardner, F. G.; Korte, N.; Strong-Gunderson, J.; Siegrist, R. L.; West, O. R.; Cline, S. R.; Baker, J. L. (1996) Implementation of Deep Soil Mixing at the Kansas City Plant, ORNL/TM-13552; Oak Ridge National Laboratory: Oak Ridge, Tennessee.
Gossett, J. M. (1987) Measurement of Henry’s Law Constants forCl and C2 Chlorinated Hydrocarbons. Environ. Sd. Technol., 69 (2), 996- 1003.
Gounot, A.-M. (1994) Microbial Oxidation and Reduction of Manganese: Consequences \in Groundwater and Applications. FEMS Micmbiol. Rev., 14 (4), 339-350.
Gupta, S.; Seagren, E. A. (2005) Comparison of Bioenhancement of Nonaqueous Phase Liquid Pool Dissolution with First- and Zero-Order Biokinetics. J. Environ. Eng., 131 (1), 165-169.
Hasten, Z. C.; McCarty, P. L. (1999) Chlorinated Ethene Half- Velocity Coefficients (K8) for Reductive Dehalogenation. Environ. Sd. Technol.. 33 (2), 223-226.
Hazen, T. C.; Sewell, G.; Gavaskar, A. R. (2000) The Effect of Source Remediation Methods on the Presence and Activity of Indigenous Subsurface Bacteria at Launch Complex 34, Cape Canaveral Air Station, Florida. Battelle: Columbus, Ohio.
Heiderscheidt, J. L. (2005) DNAPL Source Zone Depletion During In situ Chemical Oxidation (ISCO): Experimental and Modeling Studies, PhD Dissertation, Colorado School of Mines, Golden, Colorado.
Hendrickson, E. R.; Payne, J. A.; Young, R. M.; Starr, M. G.; Perry, M. P.; Fahnestock, S.; Ellis, D. E.; Ebersole, R. C. (2002) Molecular Analysis of Dehalococcoides 16S Ribosomal DNA from Chloroethene-Contaminated Sites Throughout North America and Europe. Appl. Environ. Micmbiol., 68 (2), 485-495.
Hohnstock-Ashe, A. M.; Plummer, S. M.; Yager, R. M.; Baveye, P.; Madsen, E. L. (2001) Further Biogeochemical Characterization of a Trichloroethene-Contaminated Fractured Dolomite Aquifer: Electron Source and Microbial Communities Involved in Reductive Dechlorination. Environ. Sd. Technol., 35,4449-4456.
Hrapovic, L.; Sleep, B. E.; Major, D. J.; Hood, E. (2005) Laboratory Study of Treatment of Trichloroethene by Chemical Oxidation Followed by Bioremediation. Environ. Sd. Technol., 39, 2888-2897.
Johnson, R. L.; Pankow, J. F. (1992) Dissolution of Dense Chlorinated Solvents into Groundwater. Environ. Sd. Technol., 26, 896-901.
Kastner, J. R.; Santo Domingo, J.; Denham, M.; Molina, M.; Brigmon, R. (2000) Effect of Chemical Oxidation on Subsurface Microbiology and Trichloroethene Biodegradation. Bioremediation, 4 (3), 219-236.
Kavanaugh, M. C.; Rao, P. S. C.; Abriola, L.; Cherry, J. A.; Destouni, G.; Falta, R.; Major, D. J.; Mercer, J.; Newell, C.; Sale, T.; Shoemaker, S.; Siegrist, R. L.; Teutsch, G.; Udell, K. (2003) The DNAPL Cleanup Challenge: Source Removal or Long Term Management. Report of an Expert Panel to the U.S. EPA National Risk Management Laboratory and Technology Innovation Office, EPA-600/R-03-143; U.S. Environmental Protection Agency: Washington, D.C.
Klens, J.; Scarborough, S.; Graves, D. (2001) The Effects of Permanganate Oxidation on Subsurface Microbial Populations. In Natural Attenuation of Environmental Contaminants, Leeson, A., Kelley, M. E., Rifai, H. S., Magar, V. S. (Eds.); Battelle Press: Columbus, Ohio, 253-259.
Lee, E. S.; Seol, Y.; Fang, Y. C.; Schwartz, F. W. (2003) Destruction Efficiencies and Dynamics of Reaction Fronts Associated with the Permanganate Oxidation of Trichloroethylene. Environ. Sd. Technol., 37(11), 2540-2546.
Li, X. D.; Schwartz, F. W. (2000) Efficiency Problems Related to Permanganate Oxidation Schemes. In Chemical Oxidation and Reactive Barriers; Battelle: Monterey, California, 41-48.
Macbeth, T. W.; Peterson, L. N.; Starr, R. C.; Sorenson, K. S.; Goehlert, R.; Moor, K. S. (2005) ISCO Impacts on Indigenous Microbes in a PCEDNAPL Contaminated Aquifer. In In Situ and On-Site Bioremediation Symposium; Battelle Press: Baltimore, Maryland.
MacKinnon, L. K.; Thomson, N. R. (2002) Laboratory-Scale In Situ Chemical Oxidation of a Perchloroethylene Pool Using Permanganate. J. Contam. Hydrol, 56 (1-2), 49-74.
Major, D. W.; McMaster, M. L.; Edwards, E. A.; Dworatzek, S. M.; Hendrickson, E. R.; Starr, M. G.; Payne, J. A.; Buonamici, L. W. (2002) Field Demonstration of Successful Bioaugmentation to Achieve Dechlorination of Tetrachloroethene to Ethene. Environ. Sd. Technol., 36,5106-5116.
Maym-Gatell, X.; Anguish, T.; Zinder, S. H. (1999) Reductive Dechlorination of Chlorinated Ethenes and 1,2-Dichloroethane by “Dehaloccoides ethenogenes” 195. Appl. Environ. Microbiol., 65, 3108- 3113.
Maym-Gatell, X.; Chien, Y.; Gossett, J. M.; Zinder, S. H. (1997) Isolation of a Bacterium that Reductively Dechlorinates Tetrachloroethene to Ethene. Science, 276, 1568-1571.
Maym-Gatell, X.; Tandoi, V.; Gossett, J. M.; Zinder, S. H. (1995) Characterization of an H^sub 2^-Utilizing Culture that Reductive Dechlorinates Tetrachloroethene to Vinyl Chloride and Ethene in the Absence of Methanogenesis and Acetogenesis. Appl. Environ. Microbiol., 61 (11), 3928-3933.
Miller, C. M.; Valentine, R. L.; Roehl, M. E.; Alvarez, P. (1996) Chemical and Microbiological Assessment of Pendimethalin- Contaminated Soil after Treatment with Fenton’s Reagent. Water Res., 30 (11), 2579-2586.
Schnarr, M.; Truax, C.; Hood, E.; Gonully, T.; Stickney, B. (1998) Laboratory and Controlled Field Experimentation Using Potassium Permanganate to Remediate Trichloroethylene and Perchloroethylene DNAPLs in Porous Media. J. Contant. Hydrol., 29 (3), 205-224.
Seagren, E. A.; Rittmann, B. E.; Valocchi, A. J. (1994) Quantitative Evaluation of the Enhancement of NAPL-Pool Dissolution by Flushing and Biodegradation. Environ. Sci. Technol., 2S, 833- 839.
Seitz, S. J. (2004) Experimental Evaluation of Mass Transfer and Matrix Interactions During In situ Chemical Oxidation Relying on Diffusive Transport., Master’s Thesis, Colorado School of Mines, Golden, Colorado.
Siegrist, R. L.; Lowe, K. S.; Murdoch, L. C.; Case, T. L. (1999) In Situ Oxidation by Fracture Emplaced Reactive Solids. J. Environ. Eng., 125 (5), 429-440.
Struse, A. M.; Siegrist, R. L.; Dawson, H. E.; Urynowicz, M. A. (2002) Diffusive Transport of Permanganate During In Situ Oxidation. J. Environ. Eng., 128 (4), 327-334.
U.S. Environmental Protection Agency (1999) Alternative Disinfectants and Oxidants Guidance Manual, EPA-815/R-99-014; U.S. Environmental Protection Agency: Washington, D.C.
Wiberg, K. B.; Saegebarth, K. A. (1957) The Mechanisms of Permanganate Oxidation: IV. Hydroxylation of Olefins and Related Reactions. J. Am. Chem. Soc., 79, 2822-2824.
Xie, G.; Barcelona, M. C. (2003) Sequential Chemical Oxidation and Aerobic Biodegradation of Equivalent Carbon Number-Based Hydrocarbon Fractions in Jet Fuel. Environ. Sd. Technol., 37 (20), 4751-4760.
Yager, R. M.; Bilotta, S. E.; Mann, C. L.; Madsen, E. L. (1997) Metabolic Adaptation and In Situ Attenuation of Chlorinated Ethenes by Naturally Occurring Microorganisms in a Fractured Dolomite Aquifer near Niagara Falls, New York. Environ. Sd. Technol., 31, 3138-3147.
Yan, Y. E.; Schwartz, F. W. (1999) Oxidative Degradation and Kinetics of Chlorinated Ethylenes by Potassium Permanganate. J. Contant. Hydrol., 37 (3), 343-365.
Yang, Y.; McCarty, P. L. (2000) Biologically Enhanced Dissolution of Tetrachloroethene DNAPL. Environ. Sd. Technol., 34, 2979-2984.
Yang, Y.; McCarty, P. L. (2002) Comparison Between Donor Substrates for Biologically Enhanced Tetrachloroethene DNAPL Dissolution. Environ. Sci. Technol., 36, 3400-3404.
Jason W. Sahl1*, Junko Munakata-Marr2, Michelle L. Crimi3, Robert L. Siegrist4
1* Ph.D. student, Colorado School of Mines/ESE division, Golden, CO 80401; e-mail: firstname.lastname@example.org.
2 Associate Professor, Colorado School of Mines, Golden, Colorado.
3 Assistant Professor, Eastern Tennessee State University, Johnson City, Tennessee.
4 Professor and Division Director of Environmental Science and Engineering, Colorado School of Mines, Golden, Colorado.
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