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In Situ Bioremediation of Trichloroethylene-Contamirtated Water by d Resting-Cell Methanotrophic Microbial Filter

Robert T. Taylor* A. G. Duba UCRL-JC—112062

W. B. Durham M. L. Hanna DE93 005828 K. J. Jackson

M. C. Jovanovich R. B. Knapp

J. P. Knezovich N. N. Shah

D. R. Shonnard A. M. Wijesinghe

* Corresponding author Biomedical Sciences Division L-452 Building 361; Room 1241 Lawrence Livermore National Laboratory, University of California 7000 East Avenue Livermore, California 94550-9900

Key words: In situ microbial filter; trichloroethylene bioremediation; flow-through test bed; methanotroph; methane monooxygenase longevity; biotransformation capacity

R.T.Taylor, M. L. Hanna, N. N. Shah, D. R. Shonnard, Biomedical Sciences Division, Lawrence Livermore National Laboratory, University of California, Livermore, CA 94550.

A. G. Duba, W. B. Durham, K. J. Jackson, R. B. Knapp, A. M. Wijesinghe, Earth Sciences Department, Lawrence Livermore National Laboratory, University of California, Livermore, CA 94550.

J. P. Knezovich, M. C. Jovanovich, Environmental Sciences Division, Lawrence Livermore National Laboratory, University of California, Livermore, CA 94550.

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ABSTRACT

We are testing and developing an in situ microbial filter technology fo" remediating migrating subsurface plumes contaminated with low concentrations of trichloroethylene (TCE). Our current focus is the establishment of a replenishable bk.active zone (catalytic filter) along expanding plume boundaries by the injection of a representative methanotrophic bacterium, Methylosinus trichosporium OB3b. We have successfully demonstrated this microbial filter strategy using emplaced, attached resting cells (no methane additions) in a 1.1-m flow-through test bed loaded with water-saturated sand. Two separate 24 h pulses of TCE (109 ppb and 85 ppb), one week apart, were pumped through the system at a flow velocity of 1.5 cm/h; no TCE (<0.5 ppb) was detected on the downstream side of the microbial filter. Subsequent excavation of the wet sand confirmed the existence of a TCE-bioactive zone 19 days after it had been created. An enhanced longevity of the cellular, soluble-form methane monooxygenase produced by this methanotroph is a result of our laboratory bioreactor culturing conditions. Additional experiments with cells in sealed vials and emplaced in ttie 1.1-m test t>sd yielded a high resting-cell finite TCE biotransformation capacity of -0.25 mg per mg of bacteria; this is suitable for a planned sand-filled trench field demonstration at a Lawrence Livermore National Laboratory site.

INTRODUCTION

Groundwater aquifers at many ind'.istrial and defense-related sites are contaminated with one- and two-carbon, volatile, chlorinated aliphatic hydrocarbons [e.g-, TCE), many of which are suspected carcinogens (U.S. EPA 1981; Imfante and Tsongas 1982; Westrick, Mello, and Thomas 1984; U.S. EPA 1984). Due to the natural flow of groundwater and dispersive transport processes, aquifer contamination at such sites has usually resulted in large, dilute, migrating plumes that can extend tens of meters in depth (Layton 1989). The contaminants are present in the aqueous phase and they are sorbed on the mineral phases that constitute a heterogeneous soil or rock, """hese physical characteristics of aquifer contamination make groundwater remediation to the concertttatiaas (uatviated by the EPA ^<5 ppb for TCE\ a challenging problem.

Pump-and-treat is currently the standard method for remediating volatile organic compounds. In this process, the plume is penetrated by a number of wells, contaminated groundwater is extracted from the wells, and the water is treated and disposed of at the surface (Russel, Matthews, and Sewell 1990). Pump-and-treat has great uncertainty in the ultimate level of decontamination (Travis and Doty 1990; Baker and Herson 1991). The main cause of these uncertainties is the highly heterogeneous nature of the subsurface medium, which creates preferential flow paths for the extracted fluids. Less permeable subsurface regions receive less remediation and remain as sources of residual contamination to recontaminate the cleaned-up regions.

In situ microbial bioremediation of aquifers contaminated with volatile organic

3 compounds has received increasing attention as an alternative to pump-and-treat (Wilson etal. 1986; Lee etal. 1988; Baker and ^ irson 1991). It has the potential advantages that many common organic contaminants can be biodegraded to innocuous compounds by naturally occurring microorganisms, that it is a contaminant-destructive as opposed to a contaminant-relocative process, and that it is carried out in situ, obviating the need for disposal of the treated groundwater. Unfortunately, clear-cut proof that biodegradation of the undesired organic contaminant has taken place in an aquifer has been difficult to substantiate (Madsen, Sinclair, and Ghiorse 1991; Madsen 1992), even for TCE added in known amounts to a very small, shallow aquifer (Roberts etal. 1990; Sempririi etal. 1990; Semprini etal. 1991; Semprini and McCarty 1991).

The usual approach to in situ aquifer bioremediation is to pump a suite of nutrients into the subsurface to stimulate the growth of indigenous bacterial populations (Wilson etal. 1986; Lee etal. 1988; Flathman, Jerger, and Bottanley 1989; Roberts etal. 1990; Semprini etal. 1990; Semprini etal. 1991; Semprini and McCarty 1991; Baker and Herson 1991). Ideally, the resulting increase in biomass causes the dej.red biodegradation at an acceptable rate. However, three complications arise: (1) heterogeneous permeability of the subsurface environment makes it difficult to deliver nutrients throughout the contaminated plume, (2) nutrient pumping often causes preferential growth near the injection wells and can lead to biofouling, and (3) the biotransformation of chlorinated hydrocarbons is generally a cometabolic phenomenon.

Cometabolism, the fortuitous microbial transformation of compounds such as chlorinated aliphatic hydrocarbons, cannot provide the energy or carbon needed for cell division. Growth stimulation, therefore, depends on another carbon source that may competitively inhibit the contaminant biodegradation. In addition, the stimulation of aerobic microbial growth utilizes dissolved oxygen that is needed for contaminant cometabolism. Methanotrophs, which are ubiquitous aerobic bacteria that can be cultured on methane as a sole carbon source, typify this situation. TCE cometabolism has been reported for methanotrophic mixed cultures (Wilson and Wilson 1985; Fogel, Taddeo, and Fogel 1986; Janssen ete'. 1988; Henson etal. 1988; Henson, Yates, and Cochran 1989; Strandberg, Donaldson, and Farr 1989; Lanzarone and McCarty 1990; Strand, Bjelland, and Stensel 1990; Broholm etal. 1990; Henry and Grbic-Galic 1991; Alvarez-Cohen and McCarty 1991a; Alvarez-Cohen and McCarty 1991b) and pure cultures (Little etal 1988; Tsien etal 1989; Oldenhuis etal. 1989; Brusseau etal. 1990; Henry and Grbic-Galic 1990; Henry and Grbic-Galic 1991; Oldenhuis etal 1991) under an assortment of laboratory test conditions. With one exception (Strand, Bjelland, and Stensel 1990), the growth substrate (methane) was found to be a competitive inhibitor of methanotrophic TCE degradation in laboratory studies (Oldenhuis etal. 1989; Lanzarone and McCarty 1990; Oldenhuis etal. 1991) and an aquifer field study (Roberts et al. 1990; Semprini et al. 1990; Semprini etal. 1991; Semprini and McCarty 1991; Semprini and McCarty 1992). This is not surprising, because it has been known for some time that the methane monooxygenases (MMOs) produced by methanotrophs can exhibit oxidative activities with a broad range of organic substrates, particularly alkene epoxidations (Dalton 1980; Hou 1984). By analogy to the MMO-catalyzed conversion of propene to propene cxide, it is generally assumed (Fogel, Taddeo, and Fogel 1986; Little et al. 1988; Alvarez-Cohen and McCarty 1991b) that methanotrophs initially biotransform TCE to its corresponding

4 epoxide, which is highly reactive and rapidly hydrolyzed (half-life -12 sec) in aqueous systems (Miller and Guengerich 1982; Donaldson, Chilingarian, and Yen 1989).

As an alternative to the approach based on nutrient stimulation, we are investigating an in situ microbial filter strategy using resting (;'.e., non-dividing) cells. In this strategy, microcrganisms are injected into the subsurface ahead of migrating contaminant plumes. The inocula, obtained from surface bioreactors, could be either a pure strain or a mixture of bacteria that collectively possess the desired metabolic properties. An important requirement is that a portion of the injected microorganisms attach to the soil or rock and form a zone of enhanced biodegradative activity. Ideally, contaminated groundwater flows into this zone, the attached microbial population biodegrades the contaminant(s) at a rate that keeps pace with the rate of transport, and groundwater then exits clean. Such a strategy also allows for concurrent treatment of the contaminant source by microbial inoculation and for the emplacement of multiple biofilters across the expanding dimensions of the plume to accelerate the remediation process. This combined in situ microbial filter plu.- source treatment approach is based, in part, on techniques that have been employed for microbial enhanced oil recovery (Donaldson, Chilingarian, and Yen 1989).

RESULTS AMD DISCUSSION

Test Bed and Prototype Microorganism

To test this biofilter strategy in the laboratory, we developed a meter-scale test bed (Fig. 1a). The internal dimensions of the frame containing the test bed were 1.1-m in length, 0.4-m in height, and 0.1-m in width. Flow of water in the test bed was horizontal, along the long dimension, and was strictly one-dimensional. Except for the three glass inspection windows shown in Figure 1b, the inside surfaces of the frame were nickel metal, mostly pure nickel foil covering nickel-plated stainless steel. The frame was filled with water and then with a commercially available sand (Oklahoma No. 1). It was a high-purity quartz sand with trace amounts of hematite, illite, and calcite. The grain size distribution was narrow, with 84% of the grains having a diameter between 0.106-mm and 0.25-mm. The sand pack was completely saturated (no visible gas bubbles), and had a bulk permeability of approximately 8.5 Darcys and a porosity of 0.32 + 0.02. Once the test bed was sealed, a horizontal flow of 10 mM Higgin's phosphate buffer (Park etat. 1991; Park etal. 1992; Shah era/. 1992) was initiated and maintained at a constant rate of 200 mLVh (flow velocity -1.5 cm/h, a typical value for the flow of natural groundwater). All experiments were conducted at 21-22°C in a temperature controlled room. A colored tracer (phenol red) and time-lapse photography through the glass windows revealed a 25% increase in permeability from the top to the bottom of the test bed, but the increase was roughly constant, without any undesirable "fingering" of flow, for example. Tracer tests also showed convincingly that there were no permeability gradients in the horizontal plane and that there were no short circuits for flow through the bed.

Fluids within the wet sand pack were sampled through side-wall ports (rig. 1a) attached to the ends of porous, 6-mm outer diameter, thick-walled fritted nickel tubes that extended across the full 0.1-m width of the sand pack. The nickel tubes had roughly the same porosity and permeability as the sand pack and a pore size of

5 approximately 20 urn. Tracer tests demonstrated that fluids taken from the ports were sampled uniformly across the sand pack.

Methylosinus trichosporium OB3b was selected as the prototype methanotroph for these tests for several reasons. First, it is one of the two best characterized methane-oxidizing bacterial strains currently available. Second, it can be easily grown in bioreactors at 30°C (doubling-time ~7 h) to yield high densities of bacteria that contain only the soluble form (sMMO), as opposed to the membranous particulate form (pMMC) of intracellular methane monooxygenase (Park ef a/. 1991; Park et al. 1992; Shah era/. 1992). Third, M. trichosporium OB3b cells producing sMMO are among the most active methanotrophs with respect to TCE biodegradation catalysis (Tsien et al. 1989; Oldenhuis etal. 1989; Brusseau era/. 1990; Oldenhuis era/. 1991). In related experiments, fresh cells containing sMMO converted propene to propene oxide at rates of 100-150 nmol/min/mg dry cell wt at 30°C in the presence of 20 mM formate as an electron donor (Park etal. 1991; Park era/. 1992; Shah etal. 1992). With 0.5 umol of unlabeled or [1,2-14C]TCE (2,000 cpm/nmol) substituted for propene in the same standard assay mixture (Park etal. 1991; Park etal. 1992; Shah etal. 1992), the initial steady-state rate of biotransformation (measured either by TCE disappearance or the appearance of non-volatile, water-soluble radioactive products) is routinely 40-60 nmol/min/mg dry cell wt. We agree with other investigators (Tsien et al. 1989; Oldenhuis etal. 1989; Brusseau etal. 1990; Oldenhuis et al. 1991) that for M. trichosporium OB3b, whole-cell biodegradation of TCE is decidedly a catalytic property of the sMMO as opposed to the pMMO. The TCE biotransformation rate for cells containing sMMO is 75-100 times faster than for cells containing pMMO, under both saturating and limiting concentrations of the TCE substrate (R. T Taylor and M. L. Hanna, unpub'ished data).

The M. trichosporium OB3b bacteria used in the experiments reported here were batch cultivated in a 5-L bioreactor on minus-CuS04 Higgin's minimal nitrate salts medium that was fortified with 2x FeS0 4-7H 20 and 2x nitrate as recommended previously for forming exclusively intracellular sMMO (Park et al. 1991). However, to enhance the resting-state storage stability of this cellular sMMO, for the test bed TCE biodegradation experiments reported in this paper, the culture medium NaMo0 4-2H 20 was raised 40x to 16 uM and NiCI2 was included at 7.5 jiM (Shah, Taylor and Droege 1992). After 80 h of culturing, additional 1x amounts of both M&S0 4-7H 20 and FeS0 4-7H 20 were added to the cultures. The cells were harvested after 95-105 h at the end of their transition from a fast to a slow growth rate imposed by nitrate depletion in the culture medium.

Biodegradation Experiments

At the start of the experiments M. trichosporium OB3b bacteria, which had been batch cultured in a 5-L bioreactor to produce sMMC, were harvested by centrifugation, washed, and resuspended in Higgin's phosphate buffer (pH 7.0) (Park etal. 1991). As detailed in the legend to Figure 1a, they were then pumped into the test bed to create an attached biofilter that was about 0.1-m thick in the flow direction. Cell-count-balance data revealed that 10 % of the injected bacteria attached to the sand. After this

6 inoculation, a 24 h pulse of TCE (528 u.g) at a concentration of 109 ppb was pumped into the test bed at the inlet port. Analyses by gas chromatography of 1-mL aliquots withdrawn from ports immediately ahead of the emplaced microbial filter showed the arrival of the TCE pulse (Fig. 2); whereas, no TCE was detected immediately downstream from the filter (Fig. 3). The expected TCE arrival times for no TCE biodegradation were simulated by using a one-dimensional mathematical model for advective-dispersive transport in a porous medium witr> linear equilibrium sorption (Carslaw and Jaeger 1959). The data in Figures 2 and 3 indicate complete biodegradation of the TCE pulse (>99%). This experiment was repeated 7 days later with a second 24 h TCE pulse (416 u.g) at a concentration of 85 ppb, pumped through the originally emplaced microbial filter. The result was comparable to the first experiment in that no TCE could be detected downstream of the microbial filter (Figs. 4 and 5).

These two biofilter experiments were preceded by two mass-balance control experiments in which 24 h, ~100 ppb pulses of TCE were passed through the test bed before the M. trichosporium OB3b cells were introduced. The sequential appearance of the TCE pulse at each port of the middle row (row "C" in Fig. 1a) along the entire 1.1 -m length of the test bed is in vivid contrast to the quantitative removal of the TCE pulse by the attached microbial filter (Fig. 6). Total recoveries of TCE at the outlet port for the two mass-balance control experiments were 76% and 84 %, respectively.

Permeability measurements of the sand pack were made before the TCE mass-balance control experiments and again after the two experiments with the emplaced microbial fiter. Durinn this interval of approximately 6 weeks, fluid had been flowing continuously through ihe test bed, except during inoculation. No measurable change in permeability was observed, confirming the absence of any biofouling. Likewise, none was predicted, based on our calculations that in the microbial filter zone only 3 % of the surface area of the sand grains were covered by the attached cells.

Afte'- the post-experimental permeability measurements, 19 days after the bacteria were emplaced, the test bed cover-mounting and seal were removed and the sand pack was systematically sampled in the Plate I area (Fig. 1a), which included the injected microbial filter zone. Measurements of the attached cell population and the remaining levels of whole-cell sMMO catalytic activity (for labeled TCE) were performed with these samples (Fig. 7). The maximal attached cell population density was on the order of 10 8 cells/g of dry sand and the maxima were approximately centered around the column of ports at which they were injected; a small amount of downstream spreading can be inferred. The residual whole-cell sMMO activity averaged 0.2 nmol of TCE oxidized/min/g of dry sand; it was skewed somewhat downstream from the remaining attached bacterial population. This suggests some preferential loss of catalytic activity in the upstream portion of the metabolic filter, where most of the TCE biodegradation likely occurred. No TCE or other volatile halogenated compounds were detected when representative samples of the sand pack from throughout the excavated test bed were heated to 60°C in sealed purging vials and then analyzed by gas chromatography.

The results presented in Figures 2-7 demonstrate that a microbial filter was established by injecting resting cells and that this filter removed -100-ppb levels of TCE to below a detectable limit of 0.5 ppb. Yet, TCE removal alone does not constitute

7 proof that this contaminant disappearance is primarily due to M. trichosporium OB3b biodegradation. To address this point, an aliquot of the cells utilized for injection was set aside and later challenged with a low level of [1,2-14C]TCE. The resting-stage age of these cells was identical to that of the bacteria that earlier were attached to the sand pack and later excavated and detached (Fig. 7). In Figure 8, the concentration of [1,2-1 4C]TCE (131 ppb) was chosen to be similar to the previous test bed pulses of unlabeled TCE. Moreover, the total incubation period (Fig. 8) was selected to span the time required for a TCE wave to traverse Plate I of the test bed (Figs. 2, 4, and 6). It can be seen (Fig. 8) that 80-85% of the [1,2-14C]TCE was biotransformed to water-soluble products after -12-17h. About 10 % of the radiolabeled TCE that was added to the control incubation vials could not be recovered from the organic phase; this may account for the apparent incomplete cellular conversion. At 17 h and 35 h, all of the water-soluble 14C-product material was non-volatile (stable to freeze drying). A preliminary analysis of this mixture showed that it consisted of [14C]bicarbonate, [14C]formate, and several other components (R. T. Taylor and M. L. Hanna, unpublished data). Also, a small amount of the added 1 4 C (5-8 %) was firmly associated with the cell pellet. Similar findings have been reported by others, when [1,2-14C]TCE was incubated with either a mixed (Fogel, Taddeo, and Fogel 1986) or pure (Little era/. 1988; Oldenhuis etal. 1991) methanotrophic culture. Collectively, the data in Figure 8, in combination with the results in Figures 2-7, strongly support the view that virtually complete biodegradations of the approximately 100 ppb TCE pulses in the test bed were effected by the microbial filter.

Scale-Up to Field Site Considerations

It is interesting to note that the 1.1-m length of our flow-through test bed system (Fig. 1) does not differ markedly from the distances over which the partial biodegradation of several chlorinated ethenes was achieved in a recent field study (Roberts etal. 1990; Semprini etal. 1990; Semprini etal. 1991; Semprini and McCarty 1991). At a carefully designed and instrumented field site located in Mountain View, CA, McCarty and his colleagues evaluated in situ methanotrophic bioremediation in a shallow, semiconfined aquifer that was about 1.5-m thick and consisted of fine- to coarse-grained sands and gravel. Dissolved methane and oxygen were injected into the aquifer in alternating pulses to stimulate the growth of indigenous methanotrophs. Monitoring wells were located at 1-m, 2.2-m, and 3.8-m and an extraction well was situated 6-m downstream from the injection well. After several weeks of gas injection, biostimulation of the aquifer was achieved, but the methane-oxidizing bacteria grew preferentially within 2-m of the injection well (Roberts etal. 1990). When TCE (97 ppb and later 51 ppb) was then injected, along with more methane and oxygen, the maximum biodegradation attained within 2-m of the downstream travel was only 20-30% (Roberts et al. 1990; Semprini et al. 1990; Semprini etal. 1991; Semprini and McCariy 1991). Although our objective with a resting cell microbial filter is to eliminate growth-substrate competition with contaminant biodegradation, as well as growth-dependent oxygen exhaustion, the periodic pumping of a limited supply of methane plus air could be one method to replenish a spent biofilter. The above field study suggests that it promotes localized methanotrophic bacterial growth. The more obvious replenishment method would be simply to inject additional bioreactor-grown cells.

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In contemplating a scale-up of our resting cell methanotrophic filter to a field site near the Lawrence Livermore National Laboratory, a bioactive zone 50-m long, 5-m high, 1-rn thick, and emplaced in a sand-filled trench is estimated to be adequate to capture a migrating contaminant plume. The aquifer has a pH of about 7.5, a conductivity and ionic strength about the same as 10 mM Higgin's phosphate buffer, and contains TCE (-100 ppb) as the predominant contaminant. In planning such a scale-up, three fundamental considerations are crucial with respect to the bacteria: the efficiency of bacterial attachment in the porous subsurface media, the longevity or stability of the whole-cell sMMO catalytic activity, and the absolute quantity of TCE that a given number or mass of cells can biotransform. Since the attachment efficiency will determine the amount of biomass that must to be grown to create the filter, work is in progress to define some of the key variables that influence this process for the mineralogy characteristic of this subsurface aquifer. Much progress, however, has already been made recently toward maximizing the stability of M. trichosporium OB3b sMMO in resting cells. The bioreactor culture medium and cell harvest time have a profound influence on the subsequent sMMO longevity. If the cells are grown on standard Higgin's medium (Park etal. 1991) and harvested during the iogarithmic phase, the whole-cell sMMO half-life is only -3 days; if they are collected in the slow growth phase (Park etal. 1991), it is increased to ~6 days. But if they are cultured in a modified medium and manner and harvested as described for the test bed experiments with -100 ppb TCE, the sMMO half life (using [1,2-14C]TCE as the substrate) is extended to -35 days (Fig. 9). Figure 9 depicts the resting cell storage data for two separate batch cultivations. Similar results have been obtained when the cells were cultured in the same manner and then stored in small tubes of buffer-saturated, quartz Oklahoma No. 1 sand. Although the cause of this markedly enhanced stability is currently unknown, the consistent biphasic shape of the data (Fig. 9) suggests that the whole-cell sMMO is fully retained or spared, until some other intracellular component, perhaps NADH regenerating or energy-yielding, is consumed.

Several laboratories have reported that non-dividing resting cell suspensions of both mixed and pure methanotrophic cultures are irreversibly damaged during the catalysis of TCE degrada;,on (Henry and Grbic-Galic 1991; Alvarez-Cohen and McCarty 1991a; varez-Cohen and McCarty 1991b; Oldenhuis et al. 1991; and references therein). Cellular TCE biotransformation is accompanied by a reduced cell viability and a lowered subsequent rate of methane or TCE oxidation. These effects are thought to be imparted by the hypothetical immediate product, TCE-epoxide, which is presumed to readily form covalent adducts with intracellular macromolecules. By repeatedly exposing a mixed methanotrophic cell suspension (resting state) at 21 °C to 15 ppm TCE and measuring the disappearance of TCE after each addition until the MMO activity ceased, Alvarez-Cohen and McCarty obtained a finite TCE biotransformation capacity of 0.036 mg/mg of dry cell wt (Alvarez-Cohen and McCarty 1991a). From repeated exposures of M. trichosporium OB3b at 21 °C (no added formate) to an aqueous-phase TCE concentration of 13.1 ppm, with centrifugal reisolation of the cells between each successive addition, a finite biotransformation capacity of 0.25 mg/mg of dry cell wt (+0.03 s.d.) was found for four separate experiments. Some representative data for these experiments is plotted in Figure 10. Apparently, large differences must exist between methanotrophs with respect to their resting-cell TCE transformation capacities. In a recent follow-up paper Alvarez-Cohen

9 et al. have reported (Alvarez-Cohen et al. 1992) that the most abundant microorganism, by far, and the active TCE-degrader in their previous mixed culture (Alvarez-Cohen and McCarty 1991) was a methanotroph belonging to the genus Methylosinus. Although this newly identified Methylosinus is closely related to M. trichosporium OB3b, because of the 93.5% homology of its 5 S rRNA, it has a much lower activity, capacity, and cellular sMMO longevity than OB3b (Alvarez-Cohen and McCarty 1991a; Alvarez-Cohen and McCarty 1991b; Alvarez-Cohen et&l-1992).

A follow-up experiment was carried out to ascertain the TCE-degrading capacity of the sand-attached bacteria in the 1.1-m test bed in response to a continuous 4 ppm pulse of TCE. It can be seen from Figure 11 that this concentration of TCE was essentially eliminated by the emplaced microbial filter over the first 4-5 days. Thereafter, the level of TCE at the l-C-5 location of Plate I increased in an S-shaped manner, approximating the input concentration after -13 days, at which point the pulse was terminated. A simulation of the observed data yielded a TCE biotransformation capacity of 0.28 mg/mg of the dry cell wt, which was emplaced in the column 2-4 region of Plate I. A subsequent excavation and analysis of the test bed 22 days after the bacterial injection confirmed that cells were still attached to this region of Plate I (riot shown) in a pattern and in numbers similar to the previous experiment (Fig. 7, upper half). No residual M,2-14C]TCE degradation catalysis was detectable for the excavated test bed sand samples, following this in situ capacity experiment.

During the time course of the foregoing capacity experiment (Fig. 11), an increase in the outflow chloride concentration was measured (Fig. 12). It coincided with a decrease in the outflow TCE level (not shown). Moreover, the chloride concentration increase, which corresponds to a 90 nM increment after about 100 h (Fig. 12), represents a molar ratio of 3:1 versus the continuous TCE input concentration of ~31 nM. This provides strong direct evidence that most of the TCE that was biotransformed during the test bed capacity experiment was completely mineralized with respect to the CI" anion.

For the application of a resting methanotrophic filter to the LLNL field site mentioned above, the use of a single strain with much greater MMO longevity and a much higher TCE biotransformation capacity may offer a significant advantage over the use of a mixed culture or another given pure culture. At an estimated flow of -3.8 x 10s Uday through the sand-trench microbial filter, the TCE load would be -38 g/day. Using the bulk density of 1.69 x 10 3 kg/m3forthe sand in the microbial filter area and a M. trichosporium OB3b attached cell density of 1 x 108/g of sand, the filter biomass requirement would correspond to 21.2 kg. Based on a TCE capacity of 0-25 mg/mg of cell wt, this filter could, in principle, biodegrade the TCE flowing through the filter region for 139 days. Given these parameters, the lifetime of such a microbial filter for this particular field site would most likely be dictated by the longevity of the intracellular sMMO, the duration of the microbial attachment, and additional factors. For plumes with greater TCE loading rates, however, the finite TCE biotransformation capacity of the freshly cultured bacteria is a factor of comparable and perhaps dominant importance, when compared to enzyme longevity and attachment duration. Recent results with a mixed methanotrophic culture (Alvarez-Cohen and McCarty 1991a) and our own preliminary findings with M. trichosporium OB3b (R. T. Taylor and M. L. Hanna, unpublished data) indicate that the transformation capacity can be increased

10 by at least 2-fold with the addition of an electron donor to facilitate mere intracellular NADH regeneration. In summary, we envision a resting-cell microbial filter strategy as having the most promise for aquifers with relatively rapidly moving, very dilute plumes of contaminants such as the chlorinated ethenes and chloroform, where the operational advantages over conventional methods are potentially the greatest.

ACKNOWLEDGEMENTS

We appreciate the valuable technical input and assistance of D. J. Bishop, C. Boro, G. Cobb, F. Holdener, K. Keller, S, Martin, R. Martinelli, S. Nakano, W. Ralph, D. Ruddle, C. Rundle, C. Steffani, J. Uenu. and B. Viani in the design and construction of the 1.1-m test bed system, the injection of the bacteria, and in the collection of the TCE samples. This work was performed under the auspices of the U. S. Department of Energy (DOE) by the Lawrence Livermore National Laboratory under Contract W-7405-Eng-48. The work was supported by the DOE Office of Technology Development under the direction of Deputy Assistant Secretary Dr. C. Frank, Office of Environmental Restoration and Waste Management.

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Henson, J. M., M. V. Yates, J. W. Cochran and D. L Shackleford (1988), "Microbial removal ol hatogenateti TneVnanes, ethanes and etheytenes in an aerobic BO'I'I exposed to methane". FEMS Microbial Ecology 5 3, 193-201.

Henson, J. M., M. V. Yates and J. W. Cochran (1989), "Metabolism of chlorinated methanes, ethanes and ethylenes by a mixed bacterid culture growing on methane". Journal of Industrial Microbiollogy 4, 29-35.

Hou, C. T. (1984), "Propylene oxide production from propylene by immoDilized whole cells of Methylosinus sp CRL 31 in a gas-solid bioreactor". Applied Microbiology and Biotechnology 1 9, 1-4.

Imfante, P. F. and T. A. Tsongas (1982), "Mutagenic and oncogenic effects of chloromethanes, chloroethanes, and halogenated analogues of vinyl chloride". Environmental Science Research. 2 5, 301-327.

Janssen, D. 3., G. Grobben, h. Hoekstra, R. Cldenhuis and B. Witholt (1988), "Degradation of trans- 1,2-dichloroethene by mixud and pure cultures of methanotrophic bacteria". Applied Microbiology and Biotechnology 2 9, 3i2-399.

Layton, D. W., Ed. (1989), Baseline Public Health Assessment for CERCLA Investigations at the LLNL-Livermore Site, Technical Repcrt No. 'JOAR-10Z79, Lawrence Livermcre National Laboratory, Li^ermore, CA, June

LanzawiS, N. K awd P. L. McCarty (1990}. "Column stodtes on rretafcTCATophto degradation of frichloroethylene and 1,^-dichloroet!iane". Ground WsterZZ, 910-919.

Lee, M. D., J. M. Thomas, R. C. Borden, P. B. Bedient, C. H. Ward and J. T. Wilson (1988), "Biorestoration of aquifers contaminated with organic compounds" CRC Critica. Reviews in Environmental Control 1 8, 29-89,

Little, C. D., A. V. Palumbo, S. E. Herbes, M. E. Lidstrom, R. L Tyndall and P. J. Gilmer (1988), "Trichloroethylene biodegradation by a methane-oxidizing bacterium'1. Applied and Environmental Microbiology 5 4, 951-956.

Madsen, E. L, J. L. Sinclair and W. C. Ghiorse (1991), "In situ biodegradation: microbiological patterns in a contaminated aquifer". Science *>52, 830-833.

Madsen, E. L. (1992), "Determining in situ biodegradation". Environmental Science and Technology 2 5 1663-1673.

Miller, R. and F. P. Guengerich (1982v, "Oxidation of trichloroethylene by !iver microsomal cytochrome P-450; evidence for chlorine migration in a transition state not involviny trichloroethylene oxid^". Biochemistry 2 1 , 1090-1097.

Oldenhuis, R., R. L J. M. Vink, D. B. Janssen and 3. Witholt (1989), "Degradation of

12 chlorinated aliphatic hydrocarbons by Methylosinus trichosforium OB3b expressing soluble methane monooxygenase". Applied and Environmental Microbiology 5 5, 2819-2826.

Oldenhuis, R., J. Y. Oedz^s, J. J. van der Waarde and D. B. Janssen (1991), "Kinetics of chlorinated hydrocarbon degradation by Methylosinus trichosporium OB3b and toxicity of trichioroethylene". Applied and Environmental Microbiology 5 7, 7-14.

Park, S., M. L.. Hanna, R. T. Taylor and M. W. Droege (1991), "Batch cultivation of Methylosinus trichosporium OB3b. I: Production of soluble methane monooxygenase". Biotechnology and Bioengineering 3 8, 423-433.

Park, S., N. N. Shah, R. T. Taylor and M. VV. Droege (19^2), "Batch cultivation of Methylosinus trichosporium OF3b. II: Production of particulate methane moncoxygenase". Biotechnology and Bioengineering 4 0, I51-16"7.

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Monitoring Review Fall, p. 56

FIGURE LEGENDS

(a) Schematic of the 1.1-m test bed showing the flow direction and locations of the 75 fluid sampling ports (black dots). Bacteria (4.7 x 10 9 cells/mL) were injected into the sand test bed through the middle column of sampling ports in the left-hand plate (Plate I) at a rate of 2 mLJmin/port, using a peristaltic pump. Fluid was simultaneously removed, at 1 mL/min/port, from the second and fourth columns of ports (second peristaltic pump) to establish a closed circulatory system. The peristaltic pumping directions were then reversed and the zone between the second and fourth column of ports was flushed with 10 mM Higgin's medium phosphate buffer (pH 7.0) to remove the unattached cells. During the bacterial injection and withdrawal procedure, the ambient flow field was stagnant. From the difference between the number of cells injected and the number withdrawn, the number of bacteria that initially attached in the zone was determined to be 1.6 x 10 1 2 . Bacterial cells were enumerated with a Coulter electronic particle counter (model ZB1) having a 30 u.m aperture; and the particles counted were verified to be bacteria by their size distribution, which was determined with a Coulter channelizer and an x-y recorder. The counting fluid was 5.0-mL of triple-filtered 4% NaCI. (b) Photograph of the sealed operational test bee! system, viewed from tne opposite side and showing the glass windows. Also shown is the top assembly, which serves to seal the upper surface <?1 the sand pack. Establishment of the microbial zone and all of the subsequent flow-through experiments were carried out at 21-22°C and under dim yellow lights.

Experimental data and simulation (for no biodegroJation) of the TCE wave at the sampling port in the upper left corner of Plate I (Fig. 1a), immediately upstream from the microbial filter, during the first experiment. Travel time shown on the x-axis in this and the subsequent four figures is relative to the start of flow for a TCE wave; it can be related to position via the 1.5-cm/h flow velocity. C/C 0 is the concentration relative to the input 109 ppb TCE pulse. TCE concentrations were determined by gas chromatography, according to the U.S. EPA test method No. 601 for volatile halogenated organic compounds, Fed. Reg. 49, No. 209, Oct. 26 (1984). Aqueous (1-mL) samples were removed from the test bed via the attached ports (Fig. 1a), placed in

glass purging vials, and then stored in a refrigerator several hours until the analyses could be performed. The analyses were carried out with a Hewlett-Fackard model 5880 gas chromatograph using a J & W DB-624 capillary column (30-m x 0.5-mm i.d.). It vras equipped with an electron capture detector, an H-P model 3396 integrator, a Dynatech Precision model PTA-30 W/S auto-sampler, and an O. I. Analytical mcdel 4460A purge and trap concentrator. The detection limit for TCE was 0.5 ppb.

Fig. 3. Experimental data and simulation (for no biodegradation) of the TCE wave at the sampling port in the upper right corner of Plate I (Fig. 1a), immediately downstream from the microbial filter, during the first experiment. C/C0 is the concentration relative to the input 109 ppb TCE pulse.

Fig. 4 Experimental data and simulation (for no biodegradation) of the TCE wave at the sampling port in the lower left corner of Plate I (Fig. 1a), immediately upstream from the microbial filter, during the second experiment. C/C0 is the concentration relative to the input 85 ppb TCE pulse.

Fig. 5 Experimental data and simulation (for no biodegradation) of the TCE wave at the sampling port in the lower right corner of Plate I (Fig. 1a), immediately downstream from the microbial filter, during the second experiment. C/C0

the concentration relative to the input 85 ppb TCE pulse.

Fig. 6. Comparison of the TCE mass balance data from the test bed control experiment (469 u.g pulse), generated prior to the injection of bacteria, with the data from the first biofilter experiment collected following establishment of the microbial filter zone. Data are from samples taken a'ong the middle row (Row C) in each plate of the 1.1-m test bed. Row C is about 0.2-m from the bottom of the sand pack. The port at location I-C-1 is approximately 0.1 -m from the inlet port; the other locations, all with respect to the inlet are: U-C-1 ~0.45-m, H-C-5 ~0.65-m, and III-C-5 ~0.99-m. The emplaced microbial filter was located between 0.16 and 0.25-m.

Fig. 7. Measurements of the attached microbial population (top) in millions of attached cells/g of dry sand, and their residual sMMO [1,2-14C]TCE degradation activity (bottom), in nmol/min/g of dry sand in Plate I. Black dots

show the location of the sampling ports in Plate I for reference. These data were obtained following the two experiments with the emplaced biofilter, 19 days after the bacteria were originally injected. The sand pack was sequentially excavated in six horizontal layers between the removable fritted nickel tubes. Ten wet sand samples per layer (each containing 2.1 g of dry sand) were taken from the center of the pack across Plate I and placed in glass screw-capped scintillation vials with f ."iL of Higgin's phosphate buffer. The attached cells were then dislodged from the sand by vigorous shaking for 30 sec. After allowing the sand particles to settle, aliquots of the suspended bacteria were analyzed for their cell number and their steady-state [1,2-1 4C]TCE bioactivity. These initial steady-state rate assays were performed at 30°C in the presence of forrrate, by replacing propene with 0.5 umole of [1,2-1 4C]TCE (2,000 cpm/nmol) in our standard (0.5-mL) whole-cell sMMO incubation mixture (Park etal. 1991; Park etal. 1992; Shah etal. 1992). The lower limit for measuring the rate of radiolabeled TCE degradation in these assays is 0.02 nmol/min/g of sand.

Fig. 8. Whole-cell biotransformation of a [1,2-14C]TCE concentration similar to the unlabeled TCE pulses injected into the 1.1 -m test bed for the first two microbial filter experiments. An aliquot of the same cell suspension that was used to create the test bed microbial filter was stored 19 days at room temperature in 10 mM Higgin's phosphate buffer (pH 7.0). Samples, each containing 5 x 10 s cells, were then incubated in 1.0-mL volumes of Higgin's phosphate buffer with 1.0 nmol (131 ppb) of [1,2-14C]TCE (4,000 cpm/nmol) and the amount of bioconversion to water-soluble products was determined at the times shown. The mixtures were incubated at room temperature in 5-mL amber vials that were sealed with open-top-closure screw caps and PTFE-faced rubber septa. At the times indicated, the reactions were terminated by injecting 50-u.L ot 1.0 M NaHC03 and 1.0-mL of unlabeled TCE. The sealed vials were centrifuged 1 min at top speed in a clinical centrifuge to separate the aqueous and organic liquid phases. A sample (0.6-mL) of the upper aqueous phase was removed and its raaiua^tivity was determined with 10-mL of ICN Universol counting fluid in a Packard Tri-Carb liquid scintillation spectrometer. [1,2-14C]TCE (13.1 jiCi/|imol) was purchased from the Sigma Chemical Co. and adjusted to the desired specific radioactivity and stock concentration (0.25 mM) with unlabeled TCE (99+% obtained from the Aldrich

Chemical Co.)

Fig. 9. Resting-state ceil storage of sMMO-containing bioreactor-grown M. trichosporium OB3b. The cells were batch-cultivated in minus-CuS04-modified Higgin's medium, harvested and washed, and then stored in 10 mM Higgin's phosphate buffer (pH 7.0) as a 1.0 mg/mL cell suspension in 15-mL screw-capped polystyrene tubes (Corning) at room temperature. At the times shown, cell suspension samples were removed and assayed at 30°C for their initial steady-state rates of [1,2-14C]TCE (2,000 cpm/nmol) degradation, in the presence of formate, by substituting 0.5 umol of labeled TCE for propene in a standard whole-cell sMMO incubation mixture (Park et al. 1991; Park etal. 1992; Shah etal. 1992).

Fig. 10. Resting-state finite TCE biotransformation capacity of washed sMMO-containing M. trichosporium OB3b. Cells (0.1 mg dry wt) were incubated with either 16.4 jig of [1,2-14C)TCE (2,000 cpm/nmol) or 16.4 u.g of unlabeled TCE at room temperature in 0. 25-mL of Higgin's phosphate buffer (pH 7.0) within sealed 5-mL amber vials. At time intervals of 15 min, 30 min, 1 h, and 2 h, 50-u.L aliquots were removed from the single 14C-containing vial and the amounts of [1,2-14C]TCE bioconversion to water-soluble products were measured. At each 2 h time interval, the unlaDeled TCE vials were centrifuged 2 min at 10,000xf to pellet the cells, the supernatant liquid was removed, and the bacteria were resuspended in a fresh 0.25-mL volume of Higgin's phosphate. Radiolabeled TCE, 16.4 jag, was added to one vial and unlabeled TCE, 16.4 ng, was re-added to the remaining test exposure vials and the [1,2-14C]TCE bioconversion was again determined. This cycle was carried out 6-7 times or until the subsequent [1,2-14C]TCE bioconversion ceased. At each 2 h time interval cells in additional vials, which had been identically reisolated and repeatedly exposed to either unlabeled TCE or buffer ahne (control), were removed and assayed at 30°C for their remaining initial steady-state rates of [1,2-14C]TCE degradation.

Fig. 11. Finite TCE biotransformation capacity experiment in the 1.1-m test bed. The test bed was re-poured with new sand and then injected with freshly grov/n M. trichosporium OB3b as summarized in the legend to Fig. 1a. Depicted are

the observed experimental data for the concentration of TCE at the middle row sampling port of Plate I (Fig. 1a), immediately downstream from the microbial filter and following a continuous 4 ppm pulse of TCE at the inlet port starting at day zero. Included also is a simulation of the data, based on an exponential loss of whole-cell TCE degradation rate during the TCE exposure time.

Fig. 12. Time course of the chloride ion concentration at the downstream, outlet port (Fig. 1a) of the 1.1-m test bed during the in situ TCE biotransformation capacity experiment (Fig. 11). The levels of chloride in the outflow samples were determined by a commercial laboratory (Brown and Caldwell, Laboratories) using an analytical HPLC method.

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