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A Risk-Management Strategy for PCB-Contaminated Sediments Appendix E PCB Biodegradation PCBs are not a single compound. They are a class of compounds, all of which contain the biphenyl two-ring structure with 1 to 10 chlorine atoms attached to each molecule. The chlorine atoms replace some or all of the 10 hydrogen atoms on the biphenyl molecule. As such, there are 209 possible PCB congeners, each having its own physical and chemical properties and potential for biodegradation. In general, those with fewer chlorine atoms tend to be more readily biotransformed under aerobic conditions, and the higher chlorinated congeners are more readily biotransformed under anaerobic conditions. The potential for biodegradation is a function not only of the number of chlorine atoms on a given PCB molecule but also of their placement. PCB congeners with the chlorine atom on the ortho carbon (that ring position closest to the bond connecting the two rings) tend to be more difficult to biotransform than those with the chlorine atom in the meta or para positions, the ones farther away from the connecting bond. Various reviews have been written on PCB biodegradation (Brown et al. 1987; Abramowicz 1990; Boyle et al. 1992; Higson 1992; Mohn and Tiedje 1992; Haluska et al. 1993; Tiedje et al. 1993). Aerobic Biotransformation Aerobic biotransformation of PCBs is similar to aerobic biodegradation of the biphenyl molecule itself. The first step in this process is oxidation of the ring by a 2,3-dioxygenase, which substitutes two hydrogens with two hydroxyl groups at adjacent ortho and meta positions on the molecule. The
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A Risk-Management Strategy for PCB-Contaminated Sediments ring is then cleaved through the assistance of another dioxygenase at either of two locations. The opened ring is then further metabolized by well-known pathways. PCB congeners containing several chlorine groups, especially in the 2 (ortho) or 3 (meta) positions, can block the first oxygenation of a PCB. Other researchers (Bedard and Haberl 1990) have shown that a 3,4-dioxygenase might also be effective, thus permitting somewhat broader susceptibility of PCB congeners to aerobic biodegradation. Because chlorine atoms on the PCB ring effectively block the action of the oxygenating enzymes, only PCBs with relatively few chlorine atoms are readily susceptible to aerobic biodegradation. Anaerobic Biotransformation Anaerobic biotransformation of PCBs is significantly different from aerobic biodegradation (Tiedje et al. 1993) and is most effective with more highly chlorinated PCBs. Under anaerobic conditions, PCBs are transformed by reductive dehalogenation. Here, a chlorine atom is removed from the molecule and substituted with hydrogen. Reductive dehalogenation of organic molecules has become recognized in recent years as a general process that is effective for dehalogenating a variety of halogenated organic compounds, from pesticides and many aromatic compounds, such as PCBs, to aliphatic compounds, such as chlorinated solvents (Holliger et al. 1998; Tiedje et al. 1993). In such reductions, the haloorganic serves as an electron acceptor, the role that oxygen serves under aerobic conditions. Such dehalogenation might be a fortuitous event brought about by enzymes that are designed for other purposes. In this case, the process is called a cometabolic one. However, in some cases, organisms can use the electron-accepting potential of halo-organics in energy metabolism, in which case the process is called dehalorespiration (Holliger et al. 1998) or simply halorespiration. Dehalorespiration has been demonstrated in the case of biotransformation of chlorobenzoates (Dolfing and Tiedje 1987) and tetrachloroethene (Holliger et al. 1993). The electron donor for these transformations frequently is molecular hydrogen (H2), which is commonly formed as an intermediate in normal anaerobic fermentation of complex organic materials. In other cases, simple organic compounds, such as acetate or lactate, might serve as the electron donor. Dehalorespiration tends to be a more efficient dehalogenating process because the organisms can grow catalytically by the process; reaction rates tend to be higher, and the use of available electron donors is more efficient. Whether cometabolic or through dehalorespiration, reductive dehalogenation generally requires the presence of other organic matter, such as decaying vegetable matter, to provide the electron donor required for the process to occur.
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A Risk-Management Strategy for PCB-Contaminated Sediments Reductive dehalogenation of PCBs was first recognized as an important process by Brown et al. (1987). Whether the reductive dehalogenation that occurs is by cometabolism or dehalorespiration, or a mixture of the two, is not well known. Dehalogenation of meta chlorines is generally favored over para chlorines, which in turn are favored over ortho chlorines (Tiedje et al. 1993). There are reports of ortho dechlorination (Natarajan et al. 1996). However, there is general doubt that reductive dehalogenation will lead to complete chlorine removal from congeners with chlorine atoms in the ortho position. However, congeners with chlorine atoms only in the meta or para positions can be completely dehalogenated under proper conditions to biphenyl. Natarajan et al. (1999) studied anaerobic biodegradation of biphenyl by a methanogenic consortium. They reported complete conversion of the biphenyl molecule to carbon dioxide and methane. This conversion occurred whether biphenyl was added as the sole carbon source or as a cosubstrate along with other glucose and methanol. Complete dehalogenation of all congeners in typical PCB formulations has not yet been demonstrated. Different PCB dehalogenating organisms have different abilities to dehalogenate, and thus it is perhaps not surprising that PCB dehalogenating patterns have been found to be different with different sediments (Tiedje et al. 1993). Some were found primarily to remove meta chlorines, others para chlorines, and some both meta and para chlorines. Combinations of sediments with their different organism types would thus appear to be capable of more complete dehalogenation. However, mixing two sediments with different dehalogenating abilities did not prove more efficient except when they were used sequentially (Tiedje et al. 1993). Factors Affecting Transformation Rates Tiedje et al. (1993) suggested various factors that might affect transformation rates: PCB concentration, bioavailability, inhibitors, temperature, and nutrients. Of course, the potential and rate also depend on the microorganisms that happen to be present at a given site. Bioavailability refers partly to the fact that PCBs are highly insoluble and sorb strongly to sediments, making them less available to bacteria for dehalogenation. At times, PCBs are mixed with mineral oils that also make them less bioavailable. More highly chlorinated congeners tend to partition more strongly onto soils so that solution concentrations are less than those for less highly chlorinated species. This partitioning process also means that the more highly chlorinated species have less tendency to move from sediments to overlying waters. Transformation rates are generally directly proportional to solution concentrations rather than total sediment concentrations. Solution concentrations are generally in the
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A Risk-Management Strategy for PCB-Contaminated Sediments low nanogram per liter concentration for most individual congeners. For this reason, transformation rates can be very slow. Quensen et al. (1988) indicated that optimal rates of reductive dehalogenation of PCB occurred for concentrations in the range of several hundred to thousands of parts per million in sediments. Below 50 ppm, rates often are very slow or negligible. A similar lower threshold concentration of 35 to 45 ppm was reported by Sokol et al. (1998). The latter indicated that at higher concentrations, removal rates of complex PCB mixtures, such as Arochlor 1248, were linear initially, but eventually a plateau was reached at which no further dechlorination resulted. This plateau appeared to occur once the meta chlorines had been removed, leaving the more resistant ortho chlorines. It is unclear whether there is an absolute threshold for PCBs. If cometabolism were involved and sufficient food for growth of dehalogenating organisms were available, dehalogenation might continue. It would continue, however, at rates that are so slow that they are difficult to measure, but they might still be significant over periods of years. In any event, complete anaerobic dehalogenation of PCBs in sediments has not yet been reported in the peer-reviewed literature. Combined Anaerobic and Aerobic Biodegradation Because anaerobic conditions are best for decomposition of more-chlorinated PCBs and aerobic conditions for less-chlorinated PCBs, it would appear that sequential anaerobic and aerobic treatment could result in complete PCB degradation. For example, anaerobic degradation in sediments would lead to the formation of less-chlorinated congeners that are more mobile and would diffuse into overlying aerobic waters in a river, where aerobic degradation could occur (Tiedje et al. 1993). Because of the slowness of the process, this route of complete decomposition has not been adequately demonstrated at given sites. Generally, when PCB concentrations are high in surface sediments, they are also high in fish. Although complete destruction of some congeners might result from this combination of processes, success in protecting fish from PCB contamination has not been demonstrated adequately. REFERENCES Abramowicz, D.A. 1990. Aerobic and anaerobic biodegradation of PCBs: a review. Crit. Rev. Biotechnol. 10(3):241–251. Bedard, D.L., and M.L.Haberl. 1990. Influence of chlorine substitution pattern on the degradation of polychlorinated-biphenyls by 8 bacterial strains. Microb. Ecol. 20(2):87–102.
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A Risk-Management Strategy for PCB-Contaminated Sediments Boyle, A.W., C.J.Silvin, J.P.Hassett, J.P.Nakas, and S.W.Tanenbaum. 1992. Bacterial PCB biodegradation. Biodegradation 3(2/3):285–298. Brown, J.F., Jr., R.E.Wagner, H.Feng, D.L.Bedard, M.J.Brennan, J.C.Carnahan, and R.J.May. 1987. Environmental dechlorination of PCBs. Environ.Toxicol. Chem. 6(8):579–594. Dolfing, J., and J.M.Tiedje. 1987. Growth yield increase linked to reductive dechlorination in a defined 3-chlorobenzoate degrading methanogenic coculture. Arch. Microbiol. 149(2):102–105. Haluska, L., S.Balaz and K.Dercova. 1993. Microbial degradation of polychlorinated-biphenyls. Chemicke Listy 87(10):697–708. Higson, F.K. 1992. Microbial degradation of biphenyl and its derivatives. Adv. Appl. Microbiol. 37:135–164. Holliger, C., G.Schraa, A.J Stams, and A.J.Zehnder. 1993. A highly purified enrichment culture couples the reductive dechlorination of tetrachloroethene to growth. Appl. Environ. Microbiol. 59(9):2991–2997. Holliger, C., G.Wohlfarth, and G.Diekert. 1998. Reductive dechlorination in the energy metabolism of anaerobic bacteria. FEMS Microbiol. Rev. 22(5):383–398. Mohn, W.W., and J.M.Tiedje. 1992. Microbial reductive dehalogenation. Microbiol. Rev. 56(3):482–507. Natarajan, M.R., W.-M.Wu, J.Nye, H.Wang, L.Bhatnagar, and M.K.Jain. 1996. Dechlorination of polychlorinated biphenyl congeners by an anaerobic microbial consortium. Appl. Microbiol. Biotechnol. 46(5–6):673–677. Natarajan, M.R., W.-M.Wu, R.Sanford, and M.K.Jain. 1999. Degradation of biphenyl by methanogenic microbial consortium. Biotechnol. Lett. 21(9):741–745. Quensen, J.F.III, J.M.Tiedje, and S.A.Boyd. 1988. Reductive dechlorination of polychlorinated-biphenyls by anaerobic microorganisms from sediments. Science 242(4879):752–754. Sokol, R.C., C.M.Bethoney, and G.Y.Rhee. 1998. Effect of Aroclor-1248 concentration on the rate and extent of polychlorinated biphenyl dechlorination. Environ. Toxicol. Chem. 17(10):1922–1926. Tiedje, J.M., J.F.Quensen III, J.Chee-Sanford, J.P.Schimel, and S.A.Boyd. 1993. Microbial reductive dechlorination of PCBs. Biodegradation 4(4):231–240.
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