Environ. Sci. Technol. 2000, 34, 3261-3268
Sustained Degradation of Trichloroethylene in a Suspended Growth Gas Treatment Reactor by an Actinomycetes Enrichment SEUNG-BONG LEE,† STUART E. STRAND,‡ AND H . D A V I D S T E N S E L * ,† Department of Civil and Environmental Engineering and College of Forest Resources, University of Washington, Seattle, Washington 98195
A laboratory-scale, suspended-growth, gas treatment bioreactor was used to study the sustained degradation of trichloroethylene (TCE)-contaminated gas by a phenolfed actinomycetes enrichment under aerobic conditions for 95 days at reactor TCE loadings of 27 and 72 mg L-1 d-1 with a 6-day solids retention time. The presence of TCE and phenol provided a selective pressure favoring the enrichment of actinomycetes over a TCE-degrading filamentous enrichment that was used to seed the reactor. The TCE transformation capacity of the actinomycetes enrichment in a batch test with phenol addition was 1.0 mg of TCE/mg of volatile suspended solids (VSS). This relatively high resistance to TCE metabolite toxicity resulted in sustained TCE degradation at a reactor transformation quantity (Tq) of 0.59 mg of TCE/mg of VSS wasted with a relatively high transformation yield (Ty) at 0.30 mg of TCE/ mg of phenol. Michaelis-Menten kinetic parameters for the actinomycetes enrichment were kTCE ) 0.15 d-1 and KS,TCE ) 0.14 mg/L for TCE and kP ) 10.2 d-1 and KS,P ) 0.34 mg/L for phenol. Phenol concentrations above 15 mg/L inhibited phenol degradation. Competitive inhibition between the growth substrate, phenol, and TCE was observed. The actinomycetes culture was capable of cis-DCE degradation.
Introduction Trichloroethylene (TCE) is one of the most common contaminants found at hazardous waste sites in both saturated and vadose zones of the subsurface environment. Biological treatment of TCE has received much interest as a potential lower cost remediation alternative and has been investigated for in situ subsurface treatment, treatment of liquids from pump-and-treat operation, or treatment of contaminated gases from soil vapor extraction (SVE). Biological treatment of TCE can occur by either aerobic oxidation or anaerobic dechlorination. Anaerobic degradation of TCE to ethene has been demonstrated, but in some cases these transformations may be incomplete and result in the accumulation of vinyl chloride, a carcinogenic metabolite (1). Aerobic degradation of TCE is preferable for aboveground operations treating SVE * Corresponding author telephone: (206)543-9358; fax: (206)6859185; e-mail:
[email protected]. † Department of Civil and Environmental Engineering. ‡ College of Forest Resources. 10.1021/es9907515 CCC: $19.00 Published on Web 06/29/2000
2000 American Chemical Society
gases because of the oxygen content in the gas. Landa et al. (2) showed that TCE-contaminated gas could be treated in a sparged, suspended growth reactor by Burkholderia cepacia G4 fed with toluene. TCE degradation efficiency in subsurface or in reactors can be negatively impacted by competitive inhibition between TCE and the growth substrate for the oxygenase enzyme and by metabolite toxicity from TCE degradation. The effect of metabolite toxicity observed in batch TCE degradation tests for many cultures has been characterized by the transformation capacity (TC), which is the specific amount of TCE degraded (mg of TCE/mg of biomass) in batch tests before the biomass is inactivated. Reported TC values from one study ranged from 0.05 to 0.10 mg of TCE degraded/mg of biomass inactivated for methanotrophs and from 0.031 to 0.034 for phenol-oxidizing bacteria (3). The higher values are obtained for batch tests with electron donor addition. A filamentous phenol-degrading bacteria culture showed no effect of TCE metabolite toxicity at a transformation quantity (Tq) of 0.51 mg of TCE/mg of volatile suspended solids (VSS) in batch tests with no electron donor addition (4). Their reported value was defined as Tq and not TC since the biomass was viable and still degrading TCE. Thus, some organisms are more resistant to inactivation than others. Most investigations of aerobic TCE degradation have used Gram-negative bacteria. Representative bacteria include various methanotrophs, Nitrosomonas europaea, Pseudomonas putida F1, Burkholderia cepacia G4, Pseudomonas fluorescens, Pseudomonas mendocina, Alcaligenes denitrificans, and Alcaligenes eutrophus (5). These organisms are believed to contain mono- or dioxygenases that are capable of biotransforming chlorinated aliphatic compounds. TCE degradation by Gram-positive bacteria has been shown for Mycobacterium (6) and Rhodococcus (7), which belong to the order Actinomycetales. Actinomycetes are known to biotransform a broad variety of substrates, including aliphatic and aromatic hydrocarbons (8). Actinomycetes isolated from soil and related substrates have shown primary biodegradative activity, secreting a range of extracellular enzymes, and have exhibited the capacity to metabolize recalcitrant compounds, such as TNT (9) and azo dye (10). However, relatively little is known about the use of phenol-degrading actinomycetes for aerobic degradation of chloroethenes. In this work, the aerobic TCE degradation of a phenoldegrading Gram-positive actinomycetes enrichment culture was studied. In contrast to reported TCE-degrading actinomycetes, our actinomycetes culture had different morphological characteristics. While Mycobacterium sp. and Rhodococcus sp. are mostly unicellular and do not sporulate, the actinomycetes enrichment studied here had fully developed mycelia and sporulation characteristics. The primary objective of the work presented in this paper was to evaluate the TCE degradation ability of an actinomycetes enrichment during long-term exposure to TCE in a suspended growth reactor, fed TCE in the sparged gas stream. Of particular interest were the TCE degradation kinetics of the enrichment, the effect and degree of metabolite toxicity, the possibility of competitive inhibition with phenol, and the ability of the culture to sustain long-term TCE degradation.
Materials and Methods Gas Treatment Reactor Operation. TCE-contaminated air was continuously fed to a shallow sparged, suspended-growth enrichment reactor for 95 days with phenol as the sole source VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 1. Summary of Reactor Operating Conditionsa operation condition
I
II
reactor type period of operation, days phenol loading rate, mg L-1 d-1 TCE loading rate, mg L-1 d-1 TCE influent concen, mg/L applied phenol/TCE ratio, mg/mg TCE feeding period, h/day solids retention time, days hydraulic retention time, days
A B 1-50 167 27 0.175 6.2 24 6 3 6
A B 51-95 167 72 0.462 2.3 22 6 3 6
a Gas flow rate ) 215 mL/min, temperature ) 25 ( 1 °C, and pH ) 6.8 ( 0.2.
of carbon and energy. Phenol was fed to the reactors every 2 h at 14 mg/L in a mineral salts medium solution and was depleted in about 5 min. Two 2-L gas treatment reactors were operated in parallel with the only difference being the hydraulic retention time (HRT). Both reactors were operated at two different TCE loading rates and with a constant phenol loading rate (Table 1). These conditions were selected to evaluate operating strategies and the effect of TCE loading on the ability to sustain long-term TCE degradation. The reactors used for TCE degradation were glass cylinders with a 2-L liquid volume and a 0.4-m depth and with a stone diffuser and a stir bar located at the bottom to disperse the inlet gas and mix the reactor liquid. Temperature was maintained at 25 ( 1 °C. Feed air was supplied at a rate of a 215 mL/min from a compressed air cylinder, first passing through a glass tee that also received pure-phase TCE from a syringe pump (Sage model 355). The TCE-contaminated air then passed through a 2-L cylinder containing deionized water, which attenuated fluctuations in the gas TCE concentrations, before being sparged into the treatment reactor. Effluent gas from the treatment reactor was dried by passage through a 125-mL Erlenmeyer flask packed with CaCl2. Influent and effluent gas TCE concentrations were monitored through on-line sampling to a gas chromatograph with a flame ionization detector. The reactor was originally seeded with a filamentous, phenol-degrading bacterial enrichment that showed good TCE degradation ability in batch tests (4). The nutrient solution was batch fed every 2 h, producing an initial reactor phenol concentration of 14 mg/L (167 mg L-1 d-1). Peristaltic feed and reactor effluent withdrawal pumps were controlled by a timer (Chrontrol) to maintain a 6-day solids retention time (SRT). For reactor A, in which the HRT was less than the SRT, additional liquid was wasted by filtering 333 mL/ day of reactor contents through a paper filter with solids returned to the reactor. The reactor pH was maintained at 6.8 ( 0.2. The reactor feed solution contained 1000 mg/L phenol and the following mineral concentrations: 700 mg/L KH2PO4, 1000 mg/L K2HPO4, 200 mg/L NH4Cl, 30 mg/L MgSO4, 66.5 mg/L CaCl2‚2H2O, 300 mg/L NaHCO3, 55 µg/L CuCl2‚2H2O, 150 µg/L ZnCl2, 20 µg/L NiCl2‚6H2O, 880 µg/L FeSO4‚7H2O, 135 µg/L Al2(SO4)3‚18H2O, 280 µg/L MnCl2‚4H2O, 55 µg/L CoCl2‚6H2O, 30 µg/L NaMoO4‚2H2O, and 50 µg/L H3BO3. Measurement of Phenol Degradation Kinetic Parameters. The Michaelis-Menten parameters (kP and KS,P) for phenol were estimated using an indirect, dissolved oxygen (DO) respirometer method (11) in which the substrate removal rate was directly proportional to the oxygen depletion rate. A 1:10 diluted culture was placed in a 250-mL flask fitted with a side port valve and the contents mixed by a magnetic stir bar. A DO probe (Yellow Springs Instruments model 54) was inserted, without any headspace, and the DO concentration recorded continuously. After a stable endog3262
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enous oxygen uptake rate was obtained, phenol was injected into the test flask through a side port septum valve to give an initial phenol concentration of 4.3 mg/L. The DO concentration dropped at a zero-order rate, indicating a maximum phenol utilization rate, followed by a decline in the rate of DO depletion, and finally returning to the endogenous DO uptake rate. The conditions were adjusted to keep the DO concentration above 4 mg/L, so that the test was not oxygen limited. The oxygen uptake was recorded, corrected for the endogenous oxygen uptake, and then converted to an equivalent phenol concentration uptake. Specific phenol degradation rates with time were determined from the slope of the phenol concentration versus time, normalized by the initial VSS concentration. A statistical nonlinear regression of the specific degradation rates versus phenol concentration curve was performed using CurveExpert v1.34 software to fit the data to the Michaelis-Menten kinetic model. A value for kP was also determined directly by measuring phenol concentration with time in a batch test in which 10 mg/L of phenol was spiked into a 160-mL serum bottle with 80 mL of enrichment. To test for competitive inhibition, various TCE concentrations of up to 7.2 mg/L were added into 160-mL serum bottles together with 10 mg/L phenol. Phenol concentrations were measured over time, and initial phenol degradation rates were calculated at each time point. Measurement of TCE Degradation Kinetic Parameters. To determine the transformation capacity (TC) and the maximum specific TCE degradation rate (kTCE), pre-aerated VSS from the phenol-fed TCE gas reactor (42 mg/L) were added to four 160-mL batch serum bottles with 80 mL of nutrient solution, and the bottles were stoppered with septum valves (Mininert). The first tests were used to determine whether the culture demonstrated TCE metabolite toxicity (4). Water saturated with TCE (1.6 mL, 1100 ( 200 mg/L TCE) was added to duplicate bottles, while the other duplicate bottles were incubated without TCE. After 1 day of batch incubation in a shaker at 25 °C, preexposed bottles were sparged with nitrogen gas for 2 min, followed by 1 min of air sparging. After the sparging, various amounts of TCE and 5 mg/L of phenol were added to both sets of bottles. Additional phenol was injected after 5.3 (with TCE addition) and 9 days (without TCE). The TC values were calculated from the change in the measured aqueous TCE concentration after TCE degradation had stopped (∆CL), the gas and liquid volumes (VG and VL), the dimensionless Henry’s constant (H) (0.42), and biomass concentration (X), using eq 1, which assumed that the aqueous and vapor phase TCE were in continuous equilibrium:
TC )
(VL + HVG)∆CL VLX
(1)
Similarly the maximum specific TCE degradation rate (kTCE) was calculated using eq 1 in which kTCE was substituted for TC and the initial linear change in the liquid TCE concentration (∆CL/∆t) was substituted for ∆CL. Process Model. A mechanistic process model was used to evaluate the biokinetics of the gas treatment reactor. A model presented previously by Bielefeldt and Stensel (12) was modified to account for transient competitive inhibition kinetics between TCE and batch feeding of phenol. It describes TCE mass transfer between gas and liquid phase and biodegradation in the liquid phase. The following equations describe the effluent gas TCE concentration and the changes in the liquid TCE and phenol concentrations. The half-saturation coefficients for TCE and phenol are used for respective competitive inhibition coefficients since both substrates compete for the same enzyme (13). The bubble
residence time was observed to be less than 2 s, which was less than the time interval used to solve for the liquid TCE concentration. Thus the solution was simplified by calculating the effluent gas concentration, Tg,e, using analytical eq 2 for each interval for which eqs 3 and 4 were solved:
(
Tg,e ) HTL + (Tg,o - HTL) exp dTL Qg(Tg,o - Tg,e) ) dt V
dP ) dt
)
-KLaV Qg H
kTCETLX
(
TL + KS,TCE 1 +
kPPX
(
P + KS,P 1 +
TL KS,TCE
)
P KS,P
(2)
)
(3)
(4)
where H ) Henry’s constant for TCE, Lwater/Lair; V ) reactor volume, L; KLa ) mass transfer coefficient for TCE, d-1; Tg,e ) TCE concentration in effluent gas, mg/L; Tg,o ) TCE concentration in influent gas, mg/L; TL ) TCE concentration in liquid, mg/L; Qg ) gas flow rate, L/d; P ) phenol concentration in liquid, mg/L; X ) biomass concentration, mg/L; kTCE ) maximum specific TCE degradation rate, mg of TCE (mg of VSS)-1 d-1; kP ) maximum specific phenol degradation rate, mg of phenol (mg of VSS)-1 d-1; KS,TCE ) half-saturation coefficient for TCE, mg/L; and KS,P ) halfsaturation coefficient for phenol, mg/L. The KLa value used for TCE was set equal to the ratio of the TCE and oxygen diffusivity coefficient values, multiplied by the oxygen KLa value (0.21 min-1) as determined from clean water tests (12). For non-steady-state conditions, a numerical solution was used as a function of time to solve the equations to determine the variation of the effluent gas TCE concentration and the liquid TCE and phenol concentrations by calculating changes in liquid TCE and phenol concentration for small time intervals (0.1 min). Analytical Methods. Gas-phase TCE concentrations were measured using a GC-FID (SRI 8610B) with a 30 m × 0.53 mm RTX-1 column (Restek). The gas chromatograph conditions were as follows: temperature isothermal at 131 °C, helium carrier gas flow 45 mL/min, air flow 21 mL/min, and hydrogen flow 8 mL/min. Aqueous TCE was extracted in hexane prior to quantification using gas chromatography. The hexane used for extraction contained 0.1 ppmv ethylene dibromide (EDB) as an internal standard. Gas chromatography of hexane extracts was performed on a Perkin-Elmer Autosystem equipped with a 30 m × 0.32 mm RTX-5 column (Restek) and a 63Ni electron capture detector. The chromatographic conditions were as follows: oven isothermal at 100 °C, carrier gas (helium) pressure 8.2 psi, injector temperature 250 °C, detector temperature 380 °C, injection volume 0.5 µL, and split valve on with split flow set at 40 mL/min. The phenol concentration in the gas treatment reactor was routinely monitored using a colorimetric method (Chemet phenol kit). For quantitative detection of phenol in the range of 0.1-10 mg/L, an acetic anhydride derivitization method was used (14). The GC-FID conditions were as follows: a 30 m × 0.52 mm DB5 column (Megabore), oven temperature 100 °C for 3 min, ramp at 20 °C/min to 285 °C, helium carrier gas flow rate 10 mL/min, injector temperature 200 °C, detector temperature 250 °C, and injection volume 1.0 µL. This method was used both for determination of low levels of phenol in the gas treatment reactor and for measurement of phenol in the competitive batch tests. Spectrophotometric analyses were performed on the reactor liquid. Filtered samples of the supernatants were
FIGURE 1. Changes in TCE concentration, TCE removal efficiency, and VSS concentration during reactor start-up. analyzed on a Perkin-Elmer Lambda 18 scanning spectrophotometer at wavelengths from 250 to 390 nm. Biomass was measured by VSS (15).
Results Selection of Actinomycetes Enrichment. A filamentous phenol-degrading enrichment (4), maintained in a benchscale CSTR, was used to inoculate the gas treatment reactor at a 300 mg/L VSS concentration. The reactor was continuously fed with TCE and operated at a 6-day SRT under condition I (Table 1). For the first 3 days, the reactor was fed TCE for 3 h/d at a gas influent concentration of 0.21 mg/L. TCE was removed from the gas stream with 64% efficiency. After 3 days, the TCE feed was continuous at 0.175 mg/L, and the TCE removal efficiency declined but then improved after about 10 days (Figure 1). The biomass concentration also increased with the improved TCE degradation efficiency from day 10 to day 15. Microscopic examinations revealed that the open culture had spontaneously changed from one dominated by a Gram-negative, nonflocculating, nonbranched, filamentous bacteria to one dominated by a Grampositive, branched, mycelial actinomycetes. Other cell types such as rods and cocci were observed but not at significant levels (Figure 2). The cell culture developed a yellow color in both the reactor broth and the biomass itself. During culture tests at nonfavorable conditions, such as depletion of nutrients or a cold environment, the culture developed spores at the ends of the cell filaments. Fully developed mycelia and formation of spores are distinct characteristics not shared by other TCE-degrading actinomycetes, such as Mycobacterium sp. and Rhodococcus sp. (16). Phenol Degradation Kinetics. Kinetic tests were performed four times (days 48, 65, 70, and 83 in Figure 5) during the study using the gas reactor culture. Figure 3 shows an example of the effect of phenol concentration on the specific phenol degradation rate. For all tests, the average maximum specific phenol degradation rate (kP) was 10.2 ( 0.4 mg of phenol (mg of VSS)-1 d-1, and the average half saturation coefficient (KS,P) was 0.34 ( 0.08 mg/L. The maximum specific phenol degradation rate determined by direct measurement VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 2. Phenol-degrading actinomycetes enrichment, 1000× magnification.
FIGURE 3. Phenol degradation kinetics with reactor culture at day 65. Nonlinear regression analysis determined a kP of 10.7 mg mg-1 d-1 and a KS,P of 0.23 mg/L at 51.7 mg/L of VSS and 23.2 °C.
TABLE 2. Effect of TCE Concentration on Initial Phenol Degradation Rates in Batch Testsa TCE concen, mg/L
phenol degradation rate, mg mg-1 d-1
0 0.2 0.5 0.8 2.8 8.5
11.1 11.2 9.8 9.4 7.4 6.4
a VSS ) 42.3 mg/L, initial phenol concentration ) 10 mg/L, and temperature ) 24 °C.
of phenol depletion (Table 2) in batch tests with no TCE present was in agreement with the kP determined by the indirect respirometry method. The ability of higher phenol concentrations to inhibit phenol degradation was also investigated by observing the specific oxygen uptake rate at different initial phenol concentrations. At phenol concentrations above about 15 mg/L, substrate inhibition by phenol occurred. For example, at a phenol concentration of 25 mg/L the phenol degradation 3264
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FIGURE 4. Time course of measured average aqueous TCE concentrations in duplicate batch bottle tests at 25 °C with the actinomycetes enrichment. The average percent variation between points was 3.3 ( 2.1% for TCE-preexposed culture and 6.1 ( 6.2% for non-TCE-preexposed culture. The non-TCE-preexposed culture had the same growth condition as the preexposed culture except for 1 day without TCE at the beginning. The TC value is 1.0 mg of TCE/mg of VSS, based on the 42.3 mg/L initial VSS concentration. Phenol was added at 5 mg/L at days 1, 5, and 9. rate was 85% of the maximum degradation rate. Both reactor cultures showed similar substrate inhibition when the phenol concentration was higher than about 15 mg/L. Similar substrate inhibition by phenol has been reported by others (17). Effect of TCE Exposure: Transformation Capacity and Degradation Kinetics. Figure 4 shows the results of a batch TCE degradation test with the actinomycetes culture conducted over a 16-day period in an attempt to observe the effects of TCE transformation and to measure the TC value. Duplicate bottles were initially exposed to 16.5 mg/L of liquid concentration of TCE after spiking and mixing, while a second set of duplicate bottles was shaken under endogenous conditions without TCE. After 24 h, both TCE preexposed and non-preexposed duplicate bottles were respiked with initial liquid concentrations of 16.2 mg/L TCE and 5.0 mg/L phenol. Additional phenol and TCE were added at 5 days,
FIGURE 5. VSS, TCE removal efficiency (% Rem.), and TCE influent (Tg,o) and effluent (Tg,e) gas concentration for operating conditions I and II for the two gas treatment reactors, A and B. Each point represents the average value of three consecutive dates. Both reactors are at 6-day SRT and 25 °C. Reactors A and B operated with a 3- and 6-day hydraulic retention time, respectively. and phenol was added again at 9 days. On the basis of the phenol degradation kinetics, the estimated time for phenol depletion was about 8.3 h. The decrease in the TCE removal rate by day 15 may have been caused by a depletion of the electron donor supply or cell inactivation due to TCE metabolite toxicity. Because abiotic loss of TCE in uninoculated batch bottles was negligible (not shown), all TCE removal shown in Figure 4 was due to biological activity. The transformation quantity (Tq) of the preexposed culture for the first 24-h was 0.15 mg of TCE/mg of VSS, which apparently had a small affect on the TCE removal rate following the 24-h TCE feeding. Over the next 4 days, the non-preexposed culture removed 22% more TCE. The total observed transformation capacity (TC) for the 24-h preexposed culture based on the amount of TCE removed in 16 days was 1.0 mg of TCE/mg of VSS versus 0.89 mg of TCE/mg of VSS for the non-preexposed culture. The actual TC for the test culture was higher than this since the actinomycetes culture used for the batch bottle test was from the treatment reactor with continuous TCE degradation. On the basis of the reactor solids wasting and the amount of TCE removed, the average Tq of the culture before the batch test was 0.31 mg of TCE/mg of VSS. The specific TCE degradation rates (kTCE) were calculated over the initial period of linear TCE degradation from Figure 4 and the initial VSS concentration. The kTCE values for the preexposed and non-preexposed cultures were 0.15 and 0.14 mg of TCE (mg of VSS)-1 d-1, respectively. By day 15, the specific TCE degradation rate was significantly reduced at 0.03 mg of TCE (mg of VSS)-1 d-1. The half-saturation coefficient (KS,TCE) for TCE was estimated from a steady-state solution of the TCE removal process model with 10 days of influent and effluent gas TCE concentration data when phenol was depleted for reactor operating condition I (18). Thus, this analysis was done without competitive inhibition between phenol and TCE and used a kTCE value of 0.15 mg of TCE (mg of VSS)-1 d-1. The resulting average estimate of KS,TCE was 0.14 mg/L with a standard deviation of 0.017. Long-Term TCE Removal. Figure 5 summarizes the performance of the gas reactor from day 30 to day 95, after
the establishment of the actinomycetes culture. Both reactors A and B had the same SRT (6 days), but the liquid HRT was 3 days for reactor A and 6 days for reactor B. In our earlier work (19), metabolite accumulation and inhibition were suspected at higher TCE loadings, so the HRT of one reactor was decreased so that it would have a lower concentration of degradation products. In phase I, the percent TCE removal was similar for both reactors at about 75%, and the VSS concentration increased to about 500 mg/L. For operating condition II, the influent TCE loading was tripled (72 mg L-1 d-1) at day 50. In addition, the daily TCE feeding period for both reactors was changed to 22 h instead of 24 h to provide a 2-h recovery period each day for possible repair of cell damage due to the presence of inhibitory metabolites. Previous work had shown improved TCE degradation capacity with intermittent TCE feeding (20). At the higher TCE loading (operating condition II shown in Figure 5), the TCE removal rate increased and the treatment efficiency decreased slightly with time. The average TCE removal efficiency declined slightly from 77 ( 2.1% to 74 ( 1.9% in 5 days after the loading change. After 5 days, the treatment efficiency decreased dramatically for the next 10 days, reaching a minimum value of 63% for reactor A and 55% for reactor B. The decline in performance was greater in reactor B, which had the longer HRT, suggesting that the accumulation of a metabolite of phenol or TCE degradation was affecting performance. After day 70, the TCE removal efficiency improved to near 70%, suggesting a biological response to the earlier change. During this period, microscopic observations indicated that the actinomycetes culture still dominated the microbial community. During condition II, the biomass concentration varied between 500 and 550 mg/L, and the average observed yield in condition II was 0.51 mg of VSS/mg of phenol. A reported yield for B. G4 was 0.49 mg/mg of toluene (2). At the higher loading a yellow color was observed in the reactors. Analyses of filtered liquid samples (0.45 µm filter) on a Perkin-Elmer Lambda 18 scanning spectrophotometer at wavelength and pH conditions used by Kojima et al. (21) indicated that the yellow color was due to 2-hydroxymuconic semialdehyde (2-HMS). 2-HMS is produced from the meta cleavage of catechol by catechol-2,3-dioxygenase. In alkaline solutions, the ring-fission product produces an intense yellow color, which is abolished upon acidification (22). To confirm the involvement of catechol 2,3-dioxygenase and the production of 2-HMS, a brief batch experiment was done in which catechol was mixed with a toluene-treated reactor culture. A yellow color appeared in less than 1 min, indicating meta cleavage activity. Absorbance at 375 nm increased following this reaction. This result further supports the conclusion that phenol degradation occurred via catechol by the meta cleavage pathway and that the yellow compound observed to accumulate in the gas reactors was 2-HMS. The TCE degradation rates did not decline in batch tests with 2-HMS addition (18). Effect of Phenol Addition on TCE Removal. Following the normal injection of 14 mg/L phenol into the gas treatment reactor every 2 h, the effluent gas TCE concentration typically increased, reached a peak value after the phenol was depleted, and declined slightly for the next 20 min. At the highest TCE concentration the TCE removal efficiency decreased by only about 6%. A possible cause of the higher effluent gas TCE concentration following phenol addition is competitive inhibition between phenol and TCE. Batch phenol degradation test with different TCE concentrations showed that TCE inhibited the phenol degradation rates (Table 2), thus confirming competitive inhibition. Figure 6 compares the process model (eqs 2-4) predictions, which include competitive inhibition, to the measured liquid phenol and effluent gas TCE concentrations, after a VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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cis-DCE exposure test showed that the enrichment could degrade cis-DCE. The removal efficiency was 86% when the reactor was fed 0.304 mg/L cis-DCE. The higher cis-DCE removal efficiency may have been due to higher degradation kinetics for cis-DCE or to a greater mass transfer rate. The cis-DCE mass transfer rate should be greater than TCE, based on its lower Henry’s coefficient (Henry’s constant 0.17 versus 0.39 at 25 °C; 23).
Discussion
FIGURE 6. Comparison of model prediction to observed phenol and effluent gas TCE concentration following phenol spike concentration of 28 mg/L at the normal 2-h feeding interval. Liquid TCE concentration from model predictions are shown (Tg,o ) influent gas TCE concentration, Tg,o m ) influent gas TCE concentration model prediction, Tl m ) liquid TCE concentration model prediction, Tg,e ) effluent gas TCE concentration, Tg,e m ) effluent gas TCE concentration model prediction, P ) phenol concentration, and Pm ) phenol concentrated model prediction).
TABLE 3. Model Parameters Used for Model Prediction symbols
parameters
values
kTCE
maximum specific TCE degradation rate, 0.15 mg of TCE (mg of VSS)-1 d-1 KP maximum specific phenol degradation 11.7 rate, mg of phenol (mg of TCE)-1 d-1 KS,TCE half saturation coefficient for TCE, mg/L 0.14 KS,P half saturation coefficient for phenol, mg/L 0.34 X, mg/L biomass concentration, mg/L 520 KLa mass transfer coefficient for TCE, d-1 138 Qg gas flow rate, L/d 310 H dimensionless Henry’s constant for 0.42 TCE at 25 °C V reactor volume, L 2
test spike of 28 mg/L phenol. The Michaelis-Menten kinetic coefficient values determined for phenol and TCE were used in the model simulations (Table 3). Though the experimental effluent gas TCE concentration measurements lagged the model predictions, the model provided a reasonable representation of the observed changes in liquid phenol and effluent gas TCE concentrations after the phenol spike. Because of competitive inhibition from phenol, the TCE degradation rate was lower, causing an increase in the liquid TCE concentration. The effluent gas TCE concentration also increased and then decreased as the TCE accumulated in the reactor liquid was degraded. The phenol degradation rate was not significantly affected by the TCE, since the TCE concentration was relatively low as compared to the phenol concentration. The predicted time for phenol depletion with and without competitive inhibition was 5.6 and 4.6 min, respectively. Degradation of DCE. Since TCE and cis-DCE often occur together in contaminated aquifer and SVE gas streams, the ability of the actinomycetes culture to attack cis-DCE was studied. TCE feeding was stopped for 3 h and then replaced with cis-DCE feeding for 1 day. The operating conditions were the same as those for the TCE tests. The short-term 3266
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In this research, TCE degradation was sustained for 95 days in a laboratory-scale, suspended-growth, gas treatment bioreactor by a phenol-fed actinomycetes enrichment at a reactor TCE loading as high as 72 mg L-1 d-1. When both phenol and TCE were present, the actinomycetes culture out-competed a filamentous culture that was maintained for many months in a reactor with only phenol feeding. In batch TCE degradation tests, the filamentous culture produced the highest reported Tq value for resting cells (4). However, these results suggest that continuous exposure to TCE for only a few days resulted in an inhibitory effect on the filamentous bacteria growth, even though short-term batch TCE degradation tests indicated no negative effect of TCE exposure (4). These results also show that the presence of TCE can provide a selective pressure on cometabolic microbial populations and that the population dominating with TCE present are likely to be different than those grown with only the growth substrate. The bacteria least affected by long-term TCE exposure and metabolite toxicity will probably dominate the culture. Michaelis-Menten kinetic parameters for the actinomycetes culture were kP ) 10.2 d-1 and KS,P ) 0.34 mg/L for phenol and kTCE ) 0.15 d-1 and KS,TCE ) 0.14 mg/L for TCE. The kinetic coefficients for phenol are in the range of those reported for other phenol-degrading cultures investigated for TCE removal. Folsom et al. (17) reported a kP of 32 mg of phenol (mg of VSS)-1 d-1 and a KS,P of 0.8 mg/L for B. cepacia G4 (2 mg of VSS assumed equal to 1 mg of protein) at room temperature, but this was for a very young culture with a cell age of less than 1 day. Shurtliff et al. (24) reported a kP of 9.3 mg of phenol (mg of VSS)-1 d-1 and a KS,P of less than 3.3 mg/L for a phenol-fed mixed culture at 25 °C. The kP for the phenol-fed filamentous bacterial enrichment initially used in our reactor was 17 mg of phenol (mg of VSS)-1 d-1 (4) at 20 °C. Comparison of kTCE values to other cultures can only be qualitative, since the various cultures were grown under different conditions of cell age, temperature, and TCE exposure before TCE degradation tests. The kTCE value for actinomycetes culture (0.15 d-1) is lower than those reported by others but the same order of magnitude. Folsom et al. (17) reported a kTCE of 0.75 mg of TCE (mg of VSS)-1 d-1 and a KS,TCE of 0.39 mg/L at 26 °C for B. cepacia G4 at l0-20-h cell age. Shurtliff et al. (24) reported a kTCE of 0.33 mg of TCE (mg of VSS)-1 d-1 and a KS,TCE of 11 mg/L at 25 °C for a phenol-fed mixed culture (5-day cell age). Chang and Alvarez-Cohen (3) reported a kTCE of 0.21 mg of TCE (mg of cells)-1 d-1 and a KS,TCE of 2.04 mg/L at 20 °C for a phenol-oxidizing mixed culture (3-day cell age). Bielefeldt et al. (4) reported a kTCE of 0.18 mg of TCE (mg of VSS)-1 d-1 and a KS,TCE of 0.25 mg/L at 20 °C for the filamentous bacterial enrichment (6-day cell age). Batch TCE degradation tests have been performed by numerous investigators with and without electron donor addition to observe the effects of metabolite toxicity. In batch tests without electron donor addition, if the effect of metabolite toxicity is great, TCE degradation may cease due to cell inactivation before the depletion of stored electron donor supply limits TCE degradation. In these cases, a low TC value is observed. However, for cultures with high tolerance
to metabolite toxicity, the observed TC may be due to electron donor depletion and not to cell inactivation from metabolite toxicity. If electron donor is added, the TC value in these tests may be higher than that from only metabolite toxicity since the carbon addition may provide energy and carbon for cell repair to overcome the effects of metabolite toxicity (25). Such tests are at best a qualitative indication of tolerance to metabolite toxicity, especially at high TC. In our tests, a modest amount of phenol was added, and the TC value (1.0 mg of TCE/mg of VSS) was much higher than reported for other batch tests with electron donor addition. For methane oxidizers. the highest TC value was reported by Chang and Alvarez-Cohen (26) and was 0.54 mg of TCE/mg of cells for mixed-chemostat culture with 20 mM formate addition (energy supply only). For phenol degraders, the highest value was reported by Hopkins et al. (27), 0.38 mg of TCE/mg of cell for a mixed culture with 10 mM formate addition. Their formate addition was 0.12 mmol/mg of cells, which was much higher than the amount of electron donor addition used in our test of 0.0038 mmol phenol/mg of VSS or equivalent of 0.051 mmol of formate/mg of VSS. In other batch tests with phenol addition, the TC for the phenol-degrading culture was only 0.034 mg of TCE/mg of cells (3). Although these TC values give a qualitative indication of the effect of metabolite toxicity, they do indicate that actinomycetes culture has much a high resistance to metabolite toxicity. Again, it should be noted that the filamentous culture (4) produced a Tq value of 0.51 without electron donor addition; higher than the TC values shown for other phenol degraders (28), except the actinomycetes culture. In our reactor operation, the actinomycetes out-competed the filamentous, also suggesting that it had greater resistance to metabolite toxicity. TCE degradation was sustained with the actinomycetes culture for 13 weeks at loadings up to 72 mg L-1 d-1 in the shallow suspended-growth reactor for the treatment of TCEcontaminated gas. This work represents a significant step toward the application of a mixed culture for treating TCE in an open reactor with minimal growth substrate requirements. Previous works by others had shown higher growth substrate use, used pure cultures (which may be difficult to maintain in field applications), or only showed treatment over relatively short time intervals. Landa et al. (2) operated a gas treatment reactor similar to ours, but containing a pure culture of B. cepacia G4 fed with toluene for 6 weeks at various TCE loading rates, without apparent toxic effects of TCE or TCE conversion products. They used a much larger quantity of growth substrate per unit of TCE removal as compared to our study, but their study was not intended to optimize substrate use. For 70% TCE removal, their applied toluene to TCE ratio was 47 g/g as compared to 2.3 g/g for this study. The transformation yield value at this condition was 0.015 mg of TCE degraded/mg of toluene used, which compares to value of 0.30 mg of TCE/mg of phenol used at the same treatment level for our reactor. Over a range of TCE loadings, Ty values in the work by Landa et al. (2) ranged from 0.015 to 0.071 mg/mg. They operated the reactor with a very young cell age (0.5 days), so that cells were wasted without achieving their full TCE degradation capacity. A higher Ty would have been possible if they had used a higher cell age, but the Ty level possibly is unknown and depends on the cultures longterm resistance to metabolite toxicity. When B. cepacia G4 was cultivated in a fed-batch system without cell wasting (29) the Ty value was 0.4 mg of TCE/mg of toluene, higher than the Ty value observed in our gas treatment reactor. In the fed-batch system (29), toluene was added as an electron donor and to maintain biomass levels. The fed-batch reactor was operated for only 3 weeks with about 65% TCE conversion. There may be little practical significance to their high Ty since their study showed a that a mutant population developed that was incapable of TCE
degradation. In a follow-up study (30) B. cepacia G4 could not be maintained when it was mixed with other toluene degraders in the same fed-batch system. In contrast, the actinomycetes culture in the present study dominated over other organisms in the seed culture and was maintained long over a period of several months. For another longer-term reactor application, a membraneattached methanotrophic biofilm reactor demonstrated a sustainable TCE removal efficiency of 80-90% for about 18 weeks for operated TCE loading from 100 to 320 µmol m-2 d-1 (31). However, a Ty value of only 0.015 mg of TCE/mg of CH4 was attained. For a phenol-fed chemostat operation with a mixed culture, Ty values from 0.05 to 0.22 mg of TCE degraded/mg of phenol used were reported by Shurtliff et al. (24), which are lower than the 0.30 Ty value from our reactor operation. The amount of phenol needed for a given amount of TCE removal for the gas treatment reactor varies as a function of the reactor design, the required treatment efficiency, and the operating condition, resulting in different Ty values. Both mass transfer and biodegradation considerations affect the performance of the gas treatment reactor. In this study the reactor depth was limited to 40 cm. At greater depth, the model predicts more TCE removal from the gas phase to increase the TCE removal efficiency and improve Ty results. At higher depths and KLa values, the TCE removal efficiency will be governed by the liquid concentration, which depends on the biomass concentration and biodegradation kinetics. A higher removal efficiency requires a greater biomass concentration and growth substrate, resulting in lower Ty values. On the other hand, if the cell age was increased by decreased cell wasting, less growth substrate would be required and Ty could increase. Although the optimal Ty value was not determined in this study, the Ty value for sustained treatment was shown to be greater than found for other TCEdegrading cultures and gas treatment systems.
Acknowledgments This research was funded by a grant from the National Institute of Environmental Health and Science (NIEHS), Superfund Basic Research Program (5P42 ESO4696-07), and through CRESP funded by the Department of Energy (DSFC-1-95 EW55084).
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Received for review July 5, 1999. Revised manuscript received April 3, 2000. Accepted April 17, 2000. ES9907515
