Degradation of volatile chlorinated aliphatic priority pollutants in groundwater

Environment International, Vol. 10, pp. 291-298, 1984

0160-4120/84 $3.00 + .00 Copyright © 1985 Pergamon Press Ltd.

Printed in the USA. All rights reserved.

DEGRADATION OF VOLATILE CHLORINATED ALIPHATIC PRIORITY POLLUTANTS IN GROUNDWATER Leverett R. Smith a and James Dragun b Kennedy/Jenks Engineers, 657 Howard Street, San Francisco, CA 94105, USA (Received 16 April 1984; Accepted 15 September 1984) The known degradation products and degradative pathways involving selected volatile aliphatic chlorinated hydrocarbons in soil-groundwater systems are summarized. Current data indicate that the most commonly found products of microbial degradation of these compounds come from reductive dehalogenation, while nonmicrobial degradations tend to involve hydrolysis and/or oxidation. However, conclusions must be regarded as tentative, since most of the available studies have involved model systems and extrapolations, rather than direct studies of compounds in groundwater systems. Other potentially complicating factors, such as mobility and volatilization, are mentioned. Current knowledge is discussed in the context of data that would be desirable to obtain in future studies.

Introduction

The widespread use of chlorinated organic solvents has resulted, in recent years, in numerous cases of groundwater contamination (Conservation Foundation, 1982; Council on Environmental Quality, 1982; Environment Reporter, 1984; Page, 1981; Zoeteman et al., 1980). There has accordingly been a growing interest in studying the fate of such compounds in groundwater and in the detailed mechanisms by which these materials break down in the environment (McCarty et al., 1981; Wilson and McNabb, 1983). The latter interest has been increased by the occasional discovery in groundwater of "exotic" chlorinated hydrocarbons, non-naturally occurring compounds not known to have been introduced by human activity in the areas of discovery. For those engaged in monitoring water quality and in analyzing trace organics, it has therefore become important to gain familiarity with transformations that can occur in groundwater, the better to predict the possible spectrum of halogenated hydrocarbons that may be expected where a particular chlorinated compound has been introduced into the soil-groundwater system. Over the past two or three decades, enormous effort has been devoted to the study of synthetic chemicals in the environment. Concern about the potentially harmaAuthor to whom correspondence should be sent. bCurrent address: E. C. Jordan Co., 17515 W. Nine Mile Rd., Suite 225, Southfield, MI 48075.

ful effects of such substances led to the U.S. Environmental Protection Agency's designation of a list of "priority pollutants," encompassing a wide range of pesticides, carcinogens, and other compounds (Federal Register, 1979). Numerous items of legislation address concerns about groundwater pollution (Lehr et al., 1984). Studies of the degradation of chemicals in the environment have shown that a wide range of transformations are possible (Alexander, 1981; Chakrabarty, 1982; Kobayashi and Rittman, 1982; U.S. EPA, 1979). However, because the common one- and two-carbon chlorinated solvents have only in recent years come to be viewed as pressing concerns (Environment Reporter, 1984), the scientific literature on their environmental fate is quite sparse. It was only in 1980, for example, that firm data on the microbial degradation of so common a solvent as dichloromethane was first published (Rittmann and McCarty, 1980; Brunner et al., 1980). While the technical and economic importance of chlorinated solvents has resulted in extensive coverage in the chemical literature, the known chemistry of industrial processes (Lowenheim and Moran, 1975) may offer little guidance under the radically different conditions encountered at trace levels in groundwater. References on potential chemical hazards (Sittig, 1981), while underscoring the potential importance of environmental fate data, offer only limited guidance in this aspect. There appear to be a number of reasons for the neglect of this area of environmental research. First, the 291

292

volatile chlorinated hydrocarbons have been in use for a long time. Their properties were considered well established long before environmental pollution became a major field of scientific interest. Second, compounds such as insecticides, PCBs, and polynuclear aromatic hydrocarbons presented more acutely obvious environmental problems for many years. Third, the chemical stability, limited functionality, and high volatility of the chlorinated solvents may initially have made them less attractive research topics than the structurally more complex pollutants. Improvements in chemical analysis methods and the promulgation of such methods by regulatory agencies such as the U.S. Environmental Protection Agency (U.S. EPA) (Federal Register, 1979) have helped to spur environmental research on the volatile chlorinated compounds. The purpose of this paper is to summarize and discuss what is known about the transformations and transformation products of several volatile aliphatic chlorinated organic priority pollutants frequently encountered in soil-groundwater systems. The environmental transformations of any organic chemical in groundwater can occur by chemical reactions in solution, by chemical reactions with soil constituents, and by microbial action. This paper will primarily consider the microbial and chemical transformations of these volatile priority pollutants: tetrachloroethylene (PCE), trichloroethylene (TCE), 1,1-dichloroethylene (1,1-DCE), 1,2-dichloroethylene (1,2-DCE), vinyl chloride (chloroethylene), 1,1,2,2-tetrachloroethane, 1,1,1-trichloroethane (1,1,1TCA, or methyl chloroform), 1,2-dichloroethane (1,2DCA), chloroethane (CA), tetrachloromethane (carbon tetrachloride), trichloromethane (chloroform), and dichloromethane (methylene chloride). Since the high volatility and appreciable water solubility of these compounds can result in losses by transport and stripping, it seems appropriate to discuss those possibilities briefly. Long-term studies of the environmental degradation of these low-molecular-weight compounds can be severely hampered by physical losses that would not present difficulties with less volatile compounds. One model ecosystem study of vinyl chloride, for example, found that vinyl chloride's volatility prevented significant degradation (Lu et al., 1977).

Sorption, mobility, and volatilization In discussing the possible ways in which a compound might degrade in groundwater, it is important to be able to estimate the extent to which a compound will indeed be present in the aqueous phase of the soil-water system. Where significant amounts of organic material are present in water-saturated soil, the sorption coefficient K, of organic solute toward the soil can be calculated by the relation (McCarty et ai., 1981; Schwarzenbach and Westall, 1981)

Leverett R. Smith and James Dragun

goc'Yoc =gp, where Ko~ is the solute's partition coefficient toward the soil organic phase and fo~ is the fraction of organic carbon in the soil. This relation holds well as long as the solute is present at levels below about one-half of its water solubility. Table 1 lists water solubilities and Ko~ values for a number of volatile aliphatic chlorinated solvents. The fraction of chemical in solution can be calculated by the relation: Concentration (in water) Concentration (total)

1 Kp[P] + 1 '

where [P] is the mass (in grams) of soil per cc of water. For a soil with a total organic carbon fraction of 1°70, the tabulated Ko¢ values give (for example) a Kp value of 1.26 for trichloroethylene, and 0.44 for trichloromethane. Taking as typical a subsoil bulk density of about 1.6, with a particle density of 2.65 (Brady, 1974), a soil's pore space will be about 40°70. On this basis, a watersaturated subsoil containing 1070organic carbon and low levels of solute would have 1707oof total TCE or 3607oof total trichloromethane in solution. Such results as these suggest a fairly high mobility for this class of compounds in ground water. In contrast, DDT, a higher molecular weight chlorinated hydrocarbon with a Ko¢ of 3.9 x 106, would be relatively immobile. Where water contaminated with volatile organic solvents is exposed to air, loss of organics can be expected to be quite rapid. For a dilute solution contained in a 200 mL beaker, with efficient mixing, this loss is a first-order process with a half-life of roughly 0.5 h at 25 °C (Dilling, 1977). In general, one can expect that volatile compounds with limited water solubilities will have significant vapor pressures even in quite dilute aqueous solution (Mackay and Leinonen, 1975). Air stripping readily removes several low-molecular-weight

Table 1. Solubility and soil organic phase partition coefficients of several volatile chlorinated organic priority pollutants) Volatile priority pollutant

Water solubility (mg/L)

Ko¢

Dichloromethane Trichloromethane Tetrachloromethane 1,2-Dichloroethane 1,1,1 -Trichloroethane 1,1,2,2-Tetrachloroethane Vinyl chloride 1,1-Dichloroethylene 1,2-Dichloroethylene Trichloroethylene Tetraehloroethylene

20,000 8,200 785 8,700 720 2,900 2,700 400 600 1,100 200

8.8 44 439 14 152 118 8.2 65 59 126 364

aReference: U.S. EPA (1982).

Degradation of chlorinated pollutants in groundwater chlorinated organics from activated sludge, for example (Lurker et al., 1982; Stover and Kincannon, 1983). Volatilization may well be the most rapid route for removal of the volatile chlorinated aliphatic compounds from surface waters. However, volatilization appears unlikely to be a major pathway in the disappearance of such compounds from groundwater which is not directly exposed to air.

Microbial Degradation There are a number of significant distinctions to be made in discussing microbial degradation of xenobiotic compounds. Highly chlorinated compounds tend to be more persistent than many other substances. Depending on the substrate, aerobic or anaerobic media may result in more effective destruction of a particular compound type. While many synthetic compounds can serve as sole carbon sources for growth of microorganisms, others do not support growth and metabolize only in the presence of substrates which do support microbial growth. The latter process, known as "co-metabolism" (Alexander, 1981; Dalton and Stirling, 1982) can be quite important when the chemical structure or low concentration of a pollutant would otherwise prevent its degradation by microorganisms. One may also find at times that mixed cultures effectively destroy substances that resist pure cultures. While high concentrations of many pollutants can inhibit microbial activity, groundwater tends to be low enough in organics that rates of growth and biodegradation may at times be more limited than might be expected based on laboratory tests (Boethling and Alexander, 1978, 1979). For elucidation of particular degradative pathways, uncomplicated laboratory models may be most straightforward, but typical groundwater systems probably correspond more closely to mixed cultures using mixed substrates at fairly low concentrations. Table 2 lists the reaction half-lives, reaction pathways, and reaction products for nine volatile organic priority pollutants. Data listed in Table 2 was retrieved from the scientific literature, as referenced, and includes both microbial and chemical degradation routes. The primary microbial degradation products of volatile chlorinated hydrocarbons so far described appear to result from reductive dechlorination, i.e., the replacement of chlorine by hydrogen, under anaerobic conditions. For example, PCE is converted principally to TCE (Bouwer and McCarty, 1983a; McCarty, 1984; Parsons et al. 1982, 1983, 1984); smaller amounts of products from further reduction have also been reported, including 1,2-DCE (Parsons et al., 1982, 1983) and (tentatively) vinyl chloride (McCarthy, 1984; Parsons et al., 1982, 1984). TCE has been found to degrade mainly to DCE (Parsons et ak, 1982, 1983), with traces of vinyl chloride also evident (Parsons et al., 1982, 1984). Tetra-

293 chloromethane is reduced to trichloromethane (Bouwer and McCarty, 1983b; McCarty, 1984; Parsons et aL, 1982, 1983). 1,1,2,2-Tetrachloroethane gives 1,1,2-trichloroethane (1,1,2-TCA) (Bouwer and McCarty, 1983a; McCarty, 1984), and 1,1,1-TCA has been found to yield 1,1-DCA (Parsons et aL, 1983) and chloroethane (tentatively reported in Parsons et aL, 1983). Other researchers, using biochemical oxidation demand water in a static-culture procedure, also found degradation of chloroethanes, chloroethylenes, and halomethanes, but products were not identified (Tabak et al., 1981). Depending on conditions, half-lives for such degradations can be on the order of days to months, based on laboratory results. One field study found "no evidence of biological transformation" for carbon tetrachloride, chloroform, PCE, or TCE (Schwarzenbach et al., 1983); however, another study found degradation of halomethanes, PCE, and TCE in an aquifer (Bouwer et al., 1981; McCarty et al., 1984). Differences between the field studies may have come from differences between locations (Switzerland vs. California) or may have resulted from a divergence in methodology. The former study involved seepage to groundwater from a polluted river, while the latter involved injection of wastewater into an aquifer. In each case, the groundwater environment was anaerobic. Direct evidence for microbial reductions such as 1,2-DCE to vinyl chloride, vinyl chloride to ethylene, and trichloromethane to dichloromethane was not found in the literature; however, these reactions could reasonably be expected to occur in some cases. It is possible that the relatively high volatility of the less chlorinated hydrocarbons resulted in greater difficulty in detecting them during degradation studies. Another type of reduction, the addition of hydrogen to the double bond, has not been observed in microbial incubations of the chlorinated ethylenes. In the microbial breakdown of ], l, 1-TCA, the reduction products noted earlier can be accompanied by dehydrohalogenation products. Reported products include 1,2-DCE (evidently resulting from dehydrohalogenation with rearrangement) and (most likely) vinyl chloride, which could derive from dehydrohalogenation of 1,1DCA (Parsons et al., 1982). In a different sort of process, dichloromethane was found to be converted hydrolytically to formaldehyde by cell extracts of H y p h o m i c r o bium DM2, a species that was found to grow using dichloromethane as its sole carbon source (Stucki et al., 1981). Mixed cultures and P s e u d o m o n a s species have been found to degrade dichloromethane aerobically, both as a co-metabolite and as a growth-supporting carbon source (Brunner et aL, 1980; Rittmann and McCarty, 1980; LaPat-Polasko et al., 1984). In many cases, volatile chlorinated hydrocarbons have been found to degrade to inorganic compounds during microbial incubation studies, but it has not as yet

Leverett R. Smith and James Dragun

294

Table 2. Laboratory-derived reaction half-life (t½), reaction pathway, reaction products, and related literature references for nine volatile chlorinated organic priority pollutants. Volatile Priority Pollutant PCE

TCE

Vinyl Chloride

tv2

Pathway

Products TCE c-1,2-DCE t- 1,2-DCE Vinyl chloride

< 2 days

Microbial degradation

0.73 yr

Hydrolysis

0.3 yr at high conc.; minimal at lower cone.

Microbial degradation

0.9 yr

Hydrolysis

Dilling et al. (1975)

< 10 yr

Hydrolysis

Dilling et aL (1975); Mabey and Mill (1978); U.S. EPA (1979)

Bouwer (1983); Bouwer and McCarty (1983a); McCarty (1984); Parsons et aL (1982, 1983, 1984) Dilling et al. (1975)

c- 1,2-DCE t-I,2-DCE Vinyl chloride

slow

Bouwer et al. (1981); Parsons et al. (1982, 1984)

Microbial degradation

1,1,2-TCE

Bouwer and McCarty (1983a); McCarty (1984)

0.5 yr 0.7-0.8 yr

Hydrolysis

Acetic acid 1,1-DCE

Dilling et al. (1975); Mabey et al. (1983)

< 2 days

Microbial degradation

1,1-DCA c- 1,2-DCE t- 1,2-DCE Chloroethane Vinyl chloride

1,1,2,2-Tetrachloroethane

I,I,I-TCA

References

Bouwer (1983); Parsons et al. (1982, 1983)

Chloroethane

38 days

Hydrolysis

ethanol

Mabey and Mill (1978)

Tetrachloromethane

7 yr at 1000 mg/L 7,000 yr at 1 mg/L;

Hydrolysis

CO2 + HCI Trichloromethane, CO2

Denitrifying bacteria; Microbial degradation

Trichloromethane

Bouwer (1983); Bouwer and McCarty (1983b); Mabey and Mill (1978); Parsons et al. (1982)

< 14 days < 2 days Trichloromethane

3,500 yr. 1.25 yr <0.12 yr

Hydrolysis

Bouwer (1983); Bouwer et al. (1981); Dilling et al. (1975); Mabey

Microbial degradation

and Mill (1978); Parsons et al. (1983)

< 2 days < 14 days Dichloromethane

700 yr 1.5 yr

Hydrolysis and accompanying oxidation-reduction Microbial degradation

Methyl chloride Methanol Formic acid Formaldehyde

Dilling et aL (1975); Mabey and Mill (1978); U.S. EPA (1979) Brunner et aL (1980); LaPat-Polasko et al. (1984); Rittman and McCarty (1980); Stucki et al. (1981)

Degradationof chlorinatedpollutantsin groundwater been possible to identify all intermediates in such processes (Bouwer and McCarty, 1983a, 1983b; McCarty, 1984; U.S. EPA, 1979). Microbial degradation of the volatile chlorinated aliphatic organic priority pollutants appears best characterized under anaerobic conditions (Kobayashi and Rittmann, 1982; McCarty et al., 1984; Wilson and McNabb, 1983), although other compounds, particularly less halogenated ones, may behave differently. Microbial methane monooxygenase suspensions have been reported to oxidize trichloromethane, dichloromethane, and chloromethane to carbon dioxide, carbon monoxide, and formaldehyde, respectively (Dalton and Stirling, 1982). While laboratory studies have been and will continue to be invaluable for determining possible microbial degradation routes, it seems unlikely that the high metabolic efficiency of laboratory cultures will be matched under conditions normally prevailing in groundwater. Short laboratory-derived half-lives should be taken as lower limits, not necessarily as predicting what one can expect to find under field conditions.

Hydrolysis and Oxidation In the absence of microbial action and barring such processes as volatilization or light-induced reactions, the most commonly observed chemical reaction pathway for the volatile organic priority pollutants in water has been slow hydrolysis, at times accompanied by oxidation (see Table 1) (Dilling et aL, 1975; Mabey et al., 1983; Mabey and Mill, 1978; U.S. EPA, 1979). In such reactions, the products are usually chlorinated alcohols and/or carboxylic acids, which are water soluble and thus difficult to detect at trace levels. PCE and TCE have been reported to degrade in water, with half-lives of about 0.75 and 0.9 yr, respectively, at room temperature (about 25 °C) (DiUing et al., 1975). The degradation apparently involved oxidation as well as hydrolysis, although products were not identified. Vinyl chloride has been reported to hydrolyze slowly: one estimated half-life was less than 10 yr (U.S. EPA, 1979). The hydrolysis of 1,1,1-TCA to acetic acid (major product) and 1,1-DCE (possible minor product) has been reported to have a half-life of 0.5-0.8 yr at 25 °C (Dilling et al., 1975). The hydrolysis rate for tetrachloromethane has been reported to depend on concentration, with a half-life of approximately 7 yr at 1000 mg/L, but with one of 7,000 yr at 1 mg/L (Mabey and Mill, 1978). The hydrolysis half-life of trichloromethane has been variously reported as 1.25 yr (Dilling et aL, 1975) and 3500 yr (Mabey and Mill, 1978), and that of dichloromethane as 1.5 yr (Dilling et al., 1975) and 700 yr (Mabey and Mill, 1978). Given such conflicting values, one must view half-lives estimated by extrapolation cautiously. If the longer hydrolysis half-lives, calculated from kinetic

295 studies carried out at elevated temperatures, are indeed reliable, the shorter experimental values must result from processes other than hydrolysis. Kinetic data under environmentally relevant conditions are limited, but hydrolyses have generally been found to be first order (or pseudo-first order) at trace levels. Allowing for the probability that more than one degradative mechanism could be at work, the shorter experimental half-lives for the chlorinated methanes might be closer to the values likely to be found under environment conditions. At trace levels, certain kinds of products are easier to detect than others. Hydrolysis and oxidation products are quite water soluble, and also tend to be more reactive than the halocarbon from which they are derived. This results in transitory, difficult-to-detect intermediates. Chlorinated reduction products retain solubility and reactivity characteristics that are rather similar to those of the starting chlorinated hydrocarbons. Although the greater volatility of reduction products may result in slightly increased difficulty of analysis, they are still much more easily detected at trace concentrations than are the polar compounds formed by oxidation or hydrolysis. Tracer studies have made it possible to show that common end results of degradative metabolism, oxidation, and hydrolysis, are incorporation into cell material and production of carbon dioxide or carbonate (Bouwer, 1983; Bouwer and McCarty, 1983a, 1983b; McCarty, 1984; U.S. EPA, 1979).

Relative Rates of Reactions From the information presented in Table 2, some generalizations regarding the relative rates of microbial and chemical reactions can be made. However, the data is limited and little of it was derived from field studies in actual groundwater systems. In general, the volatile chlorinated alkanes degrade more readily than the chlorinated alkenes. In biodegration, reductive mechanisms predominate and, when they occur, are relatively fast. Anaerobic conditions favor biodegradation, while aerobic conditions favor chemical degradation. It will be of interest to see whether these patterns continue to hold as more extensive information becomes available in future environmental studies. When chemical pathways are the primary pathway of degradation, hydrolysis and oxidation are the principal reactions, but both types of chemical reaction will probably occur more slowly in groundwater than does biodegradation. However, it is most important to note that chemical and microbial reactions will usually be occurring at the same time, and both kinds of reaction will determine which degradation products are present in groundwater and what the overall degradation pathway will be. Figure 1 illustrates the transformation pathways for

296

Leverett R. Smith and James Dragun

CARBON I TETRACHLORIDEI

_1 Yl

~1 CHLOROFORM I -I I

MET"¥LENE I CHLORIDE

~1

METHYL ]

CHLORIDE

I~I c-I,2-DCE

T

PCE

TCE

._• 1,1,1-TCA

VINYL I CHLORIDE

-I

111OCE _ _1 ,.,.DCA -I

---~' CHLOROETHANE}

Fig. 1 Transformation pathways for various volatile organic priority pollutants in soil-groundwater systems.

various volatile organic priority pollutants that appear as substrates and/or products of microbial and chemical degradation. The relationships presented in this figure are based upon the information provided in Table 2 and upon the present-day understanding of the reaction mechanisms of. reductive dechlorination, hydrolysis, and oxidation, which were discussed above. Given the rudimentary state of present knowledge, such a transformation scheme must be regarded as a summary of plausible alternatives, rather than as a means of making firm predictions. For example, contamination of groundwater by 1,1,1-TCA can result in the formation of 1,2-DCE, 1,1-DCE, 1,1-DCA, or compounds not yet detected. Which degradation products will actually be observed in a given case will depend on soil conditions, types of microorganisms present, pH, temperature, and other environmental factors. Under some circumstances, degradation may be less important than physical removal by transport or volatilization.

Topics for Future Research Our knowledge of the detailed transformations of volatile chlorinated hydrocarbons in groundwater and soil remains rudimentary at present. The roles of many groundwater variables such as soil type, dissolved matter, pH, oxygen levels, temperature, microbial strains present, and pollutant levels all remain in need of detailed

further study. The foregoing discussion suggests a number of microbial and chemical pathways for which one might reasonably try to search in future studies. For example, reductive dechlorination of methylene chloride to methyl chloride appears plausible, but has not yet been confirmed. There have been intriguing reports of such compounds as vinyl chloride appearing in landfill gases (Young and Parker, 1983; Bruckmann and Muelder, 1982), which may well be due to degradation of higher-molecular-weight precursors. It will be important to perform additional studies on the dichloroethylenes, to delineate conditions under which reduction to vinyl chloride may be expected, and to assess the extent to which 1,1-DCA and 1,2-DCA may be converted environmentally to chlorethane. The possible conversion of chlorinated ethylenes to chlorinated ethanes deserves further s t u d y - a r e such transformations intrinsically difficult, or merely less favorable than the transformation routes already observed? For nonmicrobial transformations, more detailed product studies would be a helpful addition to the halflife data and the rather limited mechanistic reports now available. The relative environmental importance of hydrolytic, oxidative, and dehydrohalogenative routes requires further clarification in most cases. Above all, it will be necessary to design and carry out a variety of well-controlled field studies to determine the extent to which laboratory models can accurately predict the

Degradation of chlorinated pollutants in groundwater

behavior of volatile chlorinated aliphatic hydrocarbons under the conditions actually found in soil-groundwater systems. A c k n o w l e d g e m e n t - T h e authors wish to thank Dr. William Mabey and two anonymous reviewers for their constructive suggestions and criticism.

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