Environ. Sci. Technol. 1989, 23, 1422-1425
as EPA method 10 as part of an Ascarite COPremoval system. Registry No. COz, 124-38-9. Literature Cited (1) Grob, R. Modern Practice of Gas Chromatography, 2nd ed.; John Wiley & Sons: New York, 1985; pp 124-125. (2) Burden, R.; Faires, J.; Reynolds, A. Numerical Analysis;
PWS Publishers: Boston, MA, 1981.
(3) VOST Method 0030. Test Methods for Evaluating Solid
Wuste,Field Manual; 3rd ed.; U.S. EPA Washington, DC,
1986; Vol. 11. (4) Mellor, J. W. A Comprehensioe Treatise ofhorganic and
Theoretical Chemistry; Longmans, Green & Co.: Harlow, 1946; Vol. 111.
Essex, U.K.,
Received for review March 27, 1989. Accepted June 13,1989.
Degradation of Trichloroethylene and trans4 ,2-Dichloroethylene by a Methanotrophic Consortium in a Fixed-Film, Packed-Bed Bioreactor Gerald W. Strandberg, Terrence L. Donaldson,' and Linda L. Farr
Chemical Technology Division, Oak Ridge National Laboratory,+Oak Ridge, Tennessee 37831
A fixed-film, packed-bed bioreactor containing a consortium of microorganisms utilizing methane as the primary carbon source was used to treat a synthetic groundwater containing trichloroethylene (TCE) and trans-1,2-dichloroethylene(DCE). With TCE and DCE influent concentrations of 1 mg/L each and a residence time of -50 min, >50% of the TCE and >90% of the DCE were degraded in a single pass through the bioreactor. Further degradation of TCE was obtained with liquid recycle. The TCE degradation rate appeared to be first order in TCE concentration. The apparent fmt-order rate constant for TCE degradation was -0.02 min-'. Introduction
The ability of methaneutilizing bacteria to cometabolize short-chain chlorinated hydrocarbons such as trichloroethylene (TCE) and tram-1,2-dichloroethylene(DCE) has been reported by several groups (1-6). Little et al. recently reported the mineralization of TCE by a pure culture of a methane-oxidizing organism isolated from TCE-contaminated groundwater (3). It is generally believed that the enzyme methane monooxygenase oxidizes these chloroalkenes to epoxides, which spontaneously degrade to intermediates that can be further metabolized. We have constructed and operated a fixed-film packed-bed bioreactor to evaluate the technical feasibility of bioremediation of TCEcontaminated groundwater using methanotrophic microorganisms. The performance of our bench-scale bioreactor system and observations on the kinetics of degradation of TCE and DCE in the bioreactor are described in this report. Materials and Methods
Bioreactor Design and Operation. The bioreactor (Figure 1) consisted of a 5-cm4.d. X 110-cm-long glass column packed with 0.6-cm ceramic berl saddles (Scientific Products, McGaw Park, IL) as a support matrix for the biofilm. A concentrated feed solution containing mineral salts (3) and TCE, DCE, or both was continuously bled into a stream of process water (nonchlorinated tap water meeting potable water standards). The mixture was distributed over the top of the packing at 10 mL/min unless noted otherwise. The influent concentrations of TCE and DCE (typically 1mg/L each) were controlled by adjusting their 'Operated by Martin Marietta Energy Systems, Inc., for the US. Department of Energy under Contract DE-AC05-840R21400, 1422
Environ. Sci. Technol., Vol. 23, No. 11, 1989
concentrations in the feed concentrate and by varying the dilution with process water. A concurrent gas stream containing methane (4% v/v unless noted otherwise) and air was introduced at the top of the bioreactor at 20 mL/min. This type of bioreactor and mode of operation were chosen to promote transfer of oxygen and methane to the biofilm and to minimize stripping of TCE and DCE into the gas phase. The system was operated at ambient temperature (22-24 "C). Bioreactor performance in terms of TCE and DCE degradation was measured at liquid flow rates of 5,10,20, 35, and 50 mL/min. The mean liquid residence time at each flow rate was estimated by monitoring the effluent conductivity following pulses of NaCl (7). Influent and effluent lines were constructed primarily of glass, Viton tubing, or stainless steel. However, to minimize any errors due to adsorption of TCE and DCE, the feed and effluent streams were sampled at the points where they entered and left the reactor, respectively. Bioreactor Startup. The microbial consortium was obtained from C. D. Little (3). It originated from a TCE-contaminated groundwater monitoring well on the Oak Ridge Reservation. The culture was maintained in a mineral-salts medium (3) under an atmosphere of 20% CHI and 80% air. The microbial population in the bioreactor was established by adding approximately 50 mL of an actively growing culture and 150 mL of fresh medium and then operating the system at -99% liquid recycle at 10 mL/min. Fresh medium containing 1mg/L TCE was introduced at 0.1 mL/min. The gas stream (10 mL/min) contained 20% CHI and 80% air. After 3-4 weeks, a substantial growth of salmon-pink-colored biomass was visible throughout the column. The system was then switched to and routinely operated in a single-pass mode and the methane concentration was reduced to typically 4 % methane in air. Analytical Procedures. Trichloroethylene and DCE were analyzed by gas chromatography using a Varian 3700 gas chromatograph (Palo Alto, CA) equipped with an electron capture detector. Separation was achieved with a DB+1 megabore column (J&W Scientific, Folsom, CA) operated isothermally at 40 "C, with N2as the carrier gas (3-4 mL/min). Samples (20 mL) of liquid influent or effluent were placed in 65-mL amber bottles sealed with Teflon-lined septum closures. The bottles were placed on a rotator (20 rpm, Cole-Palmer, Chicago, IL), and the contents were allowed to equilibrate for 1 h. The headspace gas was then assayed. Samples (5 pL) of the headspace gas or the bioreactor off-gas were injected directly onto the column. Quanti-
0013-936X/89/0923-1422$01.50/0
@ 1989 American Chemical Society
Table 1. Degradation of T C E and D C E in a Trickle Bioreactor
flow rate, L/min 0.005 0.01w 0.010' 0.020 0.035 0.050
t,O min
75 50 50 30 16 11
influent, mg/L
effluent, mg/L
0.9 1.0 1.1 1.0
0.2 0.3 0.5 0.5 0.8 0.6
1.1 1.0
TCE degrad rate,b mg/day 5 7 6 12 14 29
influent, mg/L
k,E min-' 0.020 0.024 0.016 0.023 0.020 0.046
1.2
effluent, mg/L NDd NDd
1.2 0.8 1.0
0.03 0.03 0.02
1.1
DCE degrad rate? mg/day
k,' m i d
28
117
20.06 20.10
33 40 69
0.12 0.20 0.35
-
OMean residence time (see text). bCorrectedfor small losses of TCE and DCE in the off-gas (data not shown). CCemwt/Cmwt = exp(-kt). dNot detected. The detection limit was -0.01 mg/L. 'Also see Figure 2. The kt band from 0.5 to 1.2 corresponds to k values from 0.01 to OiO24 mi&. PROCESS WATER
t
5
I
m
' A
1
FIRST-ORDER KINETICS
METHANE AND AIR
t
2
BERLSADDLES
z = -kC
dC
-
In COUT = In CIN -kt
r S T A H L E S S STEEL WRE SCREEN
RESERVOIR
I
I
(STRAIGHT LINE WITH SLOPE=l)
1
-
0.5
-
0
A
-+
i
-
3
0
0 W
?
LUS TCE/DCE)
-
0.2
0 = TCE ONLY
OFF-GAS
0.05
0 =TCE
0
0.1
+ DCE
I
1
I
I
1
I
0.1
0.2
0.5
1
2
5
-
10
TCE IN (mg/L) LOUK)e-' SAMPLNG PORT
EFFLUENT
Figure 1. Schematic diagram of the fixed-film, packed-bed methanotrophic bioreactor system.
tation was based on integrated detector responses to headspace gas when known quantities of TCE and DCE were diluted in 20 mL of the mineral salts and process water feed. The detector response was linear over the concentration range from 0 to 10 mg/L TCE or DCE (initial liquid-phase concentrations). The detection limit was 10 pg/L.
-
Results and Discussion Degradation of TCE is illustrated by the data shown in Table I and Figure 2. Typically, about half of the TCE was transformed when the influent TCE concentration was -1 mg/L, whereas DCE was degraded more rapidly and to a greater extent. The TCE degradation rate in the bioreactor appeared to be first order with respect to TCE concentration over the range of 0.15-5 mg/L and was not significantly affected by the presence of DCE (Figure 2). In addition, an influent feed containing -20 mg/L TCE was treated for several days with no apparent negative effect on the system (data not shown). Approximately 40% of the TCE was transformed in a single pass at this higher TCE concentration, which is consistent with the first-order kinetics observed at lower concentrations. A t flow rates of 11,32, and 66 mL/min, the mean residence times were estimated to be 47, 17, and 10 min, respectively, based on analyses of the effluent conductivity following pulses of NaCl(7). The liquid holdup was thus about 500-650 mL, depending upon the flow rate. The conductivity responses indicated a large degree of back-
Figure 2. Performance of the fixed-film bloreactor at steady state (2 is the mean liquid residence time: see text). Liquid flow rate, 10 mL/min; gas flow rate, 20 mL/min, 4-20% methane in air.
mixing (meaning relatively stagnant or inactive regions) in the bioreactor. This suggests that equivalent or improved performance may be achievable with improvements in bioreactor design. Mean residence times (t)were interpolated for the flow rates in Table I. A fnst-order rate constant of 0.016-0.024 min-' was found throughout this range of flow rates, except for a single high value of 0.046 min-' at 50 mL/min. The DCE degradation rate may also be first order in DCE concentration; however, our limited data to date are inconclusive. Although the TCE and DCE concentrations in the effluent rose with increasing flow rate as expected, the total degradation increased, presumably because of higher reaction rates at higher average concentrations of TCE and DCE in the bioreactor. The data in Table I and Figure 2 show that -50% of the TCE was removed during a single pass through the bioreactor. However, substantially more TCE could b-e degraded when the liquid effluent was totally recycled to extend the residence time and simulate a batch experiment (Figure 3). After -1.5 h the TCE concentration decreased to 50-100 pg/L, but did not decrease further. Although we are aware through personal contacts that others have observed this lower limit, there appears to be no satisfactory explanation yet for it. If there is substrate competition between methane and TCE/DCE for the methane monooxygenase enzyme, then restriction of the methane supply might improve the TCE degradation rate. However, the lower limit seen in Figure 3 still prevailed when methane was removed either at the Environ. Sci. Technol., Vol. 23, No. 11, 1989
1423
O W V c
0.2
1% ;9,.
0
SEPARATE EXPERIMENTS A
+&A
-*-*
*
I I
shake flask experiments using, among others, a culture initiated from our bioreactor (8). No priority pollutants other than TCE and DCE could be detected in the liquid effluent from the bioreactor. Although vinyl chloride has been shown to be a product of chlorinated alkene degradation by anaerobic organisms, it is not produced by aerobic, methanotrophic organisms (3) and was not observed in our system. In a separate batch experiment, vinyl chloride at 0.3 mg/L was removed by the methanotrophic population from an actual groundwater sample containing TCE, DCE, and vinyl chloride. One additional peak was often observed during the course of bioreactor operation with DCE and batch-type DCE degradation experiments. It decreased with time during recycle and batch experiments. Mass spectrometric analysis revealed that the compound had a mass of 112 or greater and likely contained two carbon atoms, two chlorine atoms, and possibly an oxygen atom. Although the compound was not identified, its characteristics are consistent with the DCE epoxide that Janssen et al. ( 4 ) recently found when methanotrophs were exposed to DCE. On occasion, particularly during recycle, other chromatographic peaks whose elution times were between those of TCE and DCE were observed. These peaks also appeared to arise as a result of DCE metabolism; they were not noted when only TCE was fed to the system. During recycle, these peaks would disappear with time. Finally, three independent observations indicate that TCE disappearance was in fact due to microbial action. Greater than 90% of influent TCE was accounted for in the effluent liquid and off-gas both with a blank column (without packing) and when the biological activity was virtually eliminated by shutting off the gas flow. Shake flask experiments with the mixed culture from the bioreactor using 14C-labeledTCE showed that in excess of 60% of the transformed TCE was mineralized to C02 (8). About 25% of the label from transformed TCE appeared in the cell mass and 5-10% appeared in water-soluble products. Conclusions Bioremediation of TCE-contaminated groundwater appears to be technically feasible based on the performance of the bench-scale bioreactor. Further development and demonstration of this new technology is needed at the pilot scale, and plans have been made to construct and operate a l-gpm bioreactor at a field site using actual contaminated groundwater. Alternative packing materials for support of the biofilm in the bioreactor will be tested. Acknowledgments
Assistance from Dave Brown, DOE Kansas City Plant, and Sid Garland, Nic Korte, Tony Palumbo, and William Eng, Oak Ridge National Laboratory, is greatly appreciated. Registry No. TCE, 79-01-6; DCE, 156-60-5; CHI, 74-82-8. Literature Cited (1) Wilson, J. T.; Wilson, B. H. Appl. Enuiron. Microbiol. 1985, 49,242-3. (2) Wilson, J. T. et al. Biological treatment of trichloroethylene in situ. Proceedings, Symposium on Groundwater Contaminations, ASCE National Convention, Atlantic City, NJ, 27-30 April 1987. (3) Little, C. D.et al. Appl. Enuiron. Microbiol. 1988,54,951-6. (4) Janssen, D.B. et al. Toxicity of chlorinated aliphatic hydrocarbons and degradation by methanotrophic consortia. Proceedings, 4th European Congress on Biotechnology;
Environ. Sci. Technol. 1989, 23, 1425-1428
Neijassel, 0. M. et al., Eds.; Elsevier Science: Amsterdam, 1987; Vol. 3, pp 515-519. (5) Henson, J. M. et al. FEMS Microbiol. Ecol. 1988, 53, 193-201. ( 6 ) Fliermans, C. B. et al. Appl. Environ. Microbiol. 1988,54,
TTichloroetheneat the Department of Energy Kansas City Plant; ORNL/TM-llOM Oak Ridge National Laboratory:
1709-14. (7) Levenspiel, 0. Chemical Reaction Engineering, 2nd ed.; Wiley: New York, 1972; Chapter 9. (8) Garland, S. B. et al. The Use of Methanotropic Bacteria for the Treatment of Groundwater Contaminated with
Received for review January 17,1989. Revised manuscript received June 12,1989. Accepted July 24,1989. This work was supported by the Kansas City Plant, Office of Defense Programs, U.S. Department of Energy, under Contract DE-ACO5840R21400.
Oak Ridge, TN, in press.
NOTES Hydrogen Peroxide Concentration in a Northern Lake: Photochemical Formation and Diel Variability William J. Cooper*,+and David R. S. Lead Drinking Water Research Center, Florida International University, Miami, Florida 33 199, and National Water Research Institute, P.O. Box 5050, Burlington, Ontario, Canada L7R 4A6
Diel changes in H202concentration in Jacks Lake, Ontario, suggested that photochemical processes were responsible for its formation. The concentrations of H202 reached 200-400 nM by late afternoon on a sunny day and declined to below 10 nM during the night. The depletion of H202observed in near-shore lake sampling sites was faster than the dark decay rate of H20z. The dark decay rate of H20z obeyed first-order kinetics and was much faster than those previously observed in marine environments. Rain was shown to have H202concentrations up to 34 pM and may contribute to the surface water H20z concentration. Introduction Sunlight-induced photochemical processes in natural waters have implications in redox cycling, pollutant transport, and possibly biological activity (1,2 ) . Of the possible reactive photochemical intermediates that can be formed, H202is one of the more stable species (3). While numerous studies have been published regarding the spatial and temporal variability of H202in oceanic (4-10) and estuarine ( 1 1 , 12) environments, relatively little is known about freshwater systems (13-20). The in situ photochemical formation of H202is thought to result from the disproportionation of superoxide, Of (17, 18, 21-26). As such, 02-may also be important in aqueous processes and H202may be used as a way of estimating the formation rates of 02-in water (23). Several reports of diel and seasonal changes in metal speciation in freshwater have appeared (27-30). It is possible that these changes result from reactions that are initiated photochemically and may involve Hz02,as observed in oceanic environments (7,10,31). Because of the influence of pH and chloride ion on iron and copper speciation it is difficult to extend marine systems studies to freshwater. However, it is possible that reactions of this 'Florida International University. 8 National Water Research Institute. 0013-936X/89/0923-1425$01.50/0
nature are at least in part responsible for the observation of reduced forms of metals in oxygenated water. Although H202is thermodynamically a good oxidant, it is usually kinetically limited, at natural water pH and in the absence of catalysts, when reacting with organic compounds (32,33). Superoxide, on the other hand, may react directly with some pollutants and may be important in determining their fate in natural waters (3). Two such examples are benzidine and benzo[a]pyrene, which have second-order rate constants with 02-in water of >2.5 X lo7 and
