Volume II Chapter 2.0 Pages 3 of 3 page    

2.6.1.3.2. Oxidation. 2-352.6.1.3.3. Hydrolysis. 2-35

2.6.1.3.4. Biotransformation and Biodegradation. 2-36

2.6.2 Environmental Fate of Coplanar PCBs 2-37

2.6.2.1. Summary 2-37

2.6.2.2. Transport Mechanisms 2-37

2.6.2.3. Transformation Processes 2-38

2.6.2.3.1. Photodegradation 2-38

2.6.2.3.2. Oxidation 2-39

2.6.2.3.3. Hydrolysis 2-40

2.6.2.3.4. Biotransformation and Biodegradation 2-40

2.7. ENVIRONMENTAL FATE - BROMINATED COMPOUNDS 2-42

2.7.1. Summary 2-42

2.7.2. Transport Mechanisms 2-43

2.7.3. Transformation Processes 2-43

2.7.3.1. Photodegradation 2-43

2.7.3.2. Oxidation 2-44

2.7.3.3. Hydrolysis 2-45

2.7.3.4. Biotransformation and Biodegradation 2-45

REFERENCES FOR CHAPTER 2 2-47

2.6.1.3.2. Oxidation.
Stehl (1973) has suggested that 2,3,7,8-TCDD is probably stable to oxidation in the ambient environment. The reaction rates of hydroxyl (OH) radicals with CDDs and CDFs have not been measured because, in part, the low vapor pressures of these compounds make direct measurements very difficult with currently available techniques. However, Podall et al. (1986) estimated the half-life of 2,3,7,8-TCDD vapor via OH oxidation in the atmosphere to be 8.3 days. Atkinson (1987) estimated the atmospheric lifetime of about 3 days for 2,3,7,8-TCDD due to the OH radical reaction.

2.6.1.3.3. Hydrolysis.
There is no available evidence indicating that hydrolysis would be an operative environmental process for degradation of CDDs or CDFs (Leifer et al., 1993; Miller and Zepp, 1987).

2.6.1.3.4. Biotransformation and Biodegradation.
Investigations on the biodegradability of CDDs and CDFs have focused on the microbial degradation of 2,3,7,8-TCDD. Arthur and Frea (1989) provided a comprehensive review of studies conducted during the 1970s and 1980s. Arthur and Frea (1989) concluded that 2,3,7,8-TCDD is recalcitrant to microbial degradation. Several major studies conducted during this period are discussed below.

Matsumura and Benezet (1973) tested approximately 100 strains of micro-organisms that were shown previously to degrade persistent pesticides; only five strains showed any ability to degrade 2,3,7,8-TCDD, based on autoradiographs of thin-layer chromatograms. Although it is possible that the less chlorinated dioxins are more susceptible to biodegradation, microbial action on 2,3,7,8-TCDD is very slow under optimum conditions (Hutter and Philippi, 1982).

Long-term incubations of radiolabeled 2,3,7,8-TCDD yielded no radioactivity in carbon dioxide traps after 1 year, and analyses of the cultures showed that at most, 1 to 2 percent of a potential metabolite (assumed to be a hydroxylated derivative of 2,3,7,8-TCDD) could be detected. Camoni et al. (1982) added organic compost to contaminated soil from the Seveso area to enrich the soil and enhance the 2,3,7,8-TCDD biodegradation rate; however, the soil amendment had no clear effect on degradation.

Quensen and Matsumura (1983) reported that low concentrations (5 ppb) of radiolabelled 2,3,7,8-TCDD were metabolized by pure cultures of Nocardiopsis spp. and Bacillus megaterium that had been isolated from farm soil.

The extent of metabolism after 1 week incubation was strongly dependent on the carrier solvent used to dissolve and introduce the 2,3,7,8-TCDD to the culture medium. The solvent ethyl acetate gave the best results; 52 percent of 14C was recovered as 2,3,7,8-TCDD out of a total of 77 percent 14C recovered. However, incubation of 2,3,7,8-TCDD in farm soil, garden soil, and forest soil resulted in little, if any, metabolism of 2,3,7,8-TCDD.

Bumpus et al. (1985) tested the white rot fungus, Phanerochaete chrysosporium, which secretes a unique H202-dependent extracellular lignin-degrading enzyme system capable of generating carbon-centered free radicals.

Lignin is resistant to attack by all micro-organisms except some species of fungi and a relatively small number of bacteria species. Radiolabeled 2,3,7,8-TCDD was oxidized to labeled C02 by nitrogen-deficient, ligninolytic cultures of P. chrysosporium; since the label was restricted to the ring, it was concluded that the strain was able to degrade halogenated aromatic rings. In 10 ml cultures containing 1,250 pmol of substrate, 27.9 pmol of 2,3,7,8-TCDD were converted to labeled-CO2 during the 30-day incubation period; thus, only about 2 percent of the starting material were converted.

Hoffman et al. (1992) demonstrated that the fungi, Fusarium redolens, could degrade 3-chlorodibenzofuran and, to a lesser degree, mono- and di-CDDs. Hoffman et al. (1992) also identified 14 other strains of fungi that demonstrated the capability to degrade dibenzofuran (nonchlorinated).

The strains are members of the following genera:
Mucor, Chaetomium, Phoma, Fusarium, Paecilomyces, Papulaspora, Inonotus, Lentinus, Phanerochaete, Polyporus, Pycnoporus, Schizophyllum
, and Trametes.

2.6.2. Environmental Fate of Coplanar PCBs

2.6.2.1. Summary

Little specific information exists on the environmental transport and fate of the 11 coplanar PCBs. However, the available information on the physical/chemical properties of coplanar PCBs coupled with the body of information available on the widespread occurrence and persistence of PCBs in the environment indicates that these coplanar PCBs are likely to be associated primarily with soils and sediments, and to be thermally and chemically stable.

Soil erosion and sediment transport in waterbodies and volatilization from the surfaces of soils/water bodies with subsequent atmospheric transport and deposition are believed to be the dominant current transport mechanisms responsible for the widespread environmental occurrence of PCBs. Photodegradation to less chlorinated congeners followed by slow anaerobic and/or aerobic biodegradation is believed to be the principal path for destruction of PCBs.

2.6.2.2. Transport Mechanisms
Based on their low vapor pressures, low water solubilities, and high Koc values, coplanar PCBs are expected primarily to be associated with soils, sediments, and particulates; however, due to the stability and persistence of coplanar PCBs via other transformation and transport pathways, volatilization is likely to be a significant transport mechanism from a global perspective.

It should be noted that although coplanar PCBs have low vapor pressures and water solubilities, the Henry's Law constants for the similarly substituted CDDs and CDFs are expected to be one to two orders of magnitude lower. Therefore, it can be expected that volatilization, as well as desorption of PCBs from particulate matter into air and water, is likely to be more significant transport mechanisms for PCBs than for CDDs and CDFs.

For example, Murray and Andren (1992) studied the precipitation scavenging of PCBs in the Great Lakes region. They reported that atmospheric PCBs are largely in the gas phase (typically >90 percent) rather than bound to particulates. Similarly, the results of their study support the hypothesis that precipitation provides episodic inputs of PCBs to the Great Lakes, which are volatilizing the PCBs back to the atmosphere for much of the year, particularly during the summer (Baker and Eisenreich, 1990).

2.6.2.3. Transformation Processes

2.6.2.3.1. Photodegradation.
Based on the data available in 1983, Leifer et al. (1983) concluded that all PCBs, especially the more highly chlorinated congeners and those that contain two or more chlorines in the ortho position, photodechlorinate. In general, as the chlorine content increases, the photolysis rate increases. The products of photolysis are predominantly lower chlorinated PCBs.

More recently, Lepine et al. (1992) exposed dilute solutions (4ppm) of Aroclor 1254 in cyclohexane to sunlight for 55 days in December and January. Isomer-specific analysis indicated that the amounts of many higher chlorinated congeners decreased while those of some lower chlorinated congeners increased.

These results are consistent with the studies reviewed in Leifer et al. (1983) that indicated photodegradation of PCBs proceeds through successive dechlorination of the biphenyl molecule. The results for the coplanar PCBs indicated a 43.5 percent decrease in the amount of 2,3,4,4',5-PeCB, a 73.5 percent decrease in the amount of 2,3,3',4,4',5-HxCB, and a 24.4 percent decrease in the amount of 2,3,3',4,4',5'-HxCB. However, 3,3',4,4'-TeCB and 3,3',4,4',5-PeCB, which were not detected in unirradiated Aroclor 1254, represented 2.5 percent and 0.43 percent, respectively, of the irradiated mixture.

The authors postulated that formation of these two congeners probably occurred, at least in part, from dechlorination at the ortho position of their mono-ortho-substituted precursors, considering the greater reactivity of PCB ortho chlorines toward photodechlorination.

2.6.2.3.2. Oxidation.
Reaction of PCBs with common environmental oxidants such as hydroperoxy radicals (HO2) and ozone (O3) has not been reported and are probably not very important because only very strong oxidant species can react with PCBs (Sedlak and Andren, 1991). However, reaction of gas-phase PCBs in the atmosphere and dissolved PCBs in certain surface waters with hydroxyl radicals (OH) (one of the strongest environmental oxidants known) may be an important degradation mechanism.

Atkinson (1987) and Leifer (1983), using assumed steady-state atmospheric OH concentrations and measured oxidation rate constants for biphenyl and monochlorobiphenyl, estimated atmospheric decay rates and half-lives for gaseous-phase PCBs. Atmospheric transformation was estimated to proceed most rapidly for those PCB congeners containing either a small number of chlorines or those containing all or most of the chlorines on one ring.

Diagram V2 2-6Sedlak and Andren (1991) demonstrated in laboratory studies that OH radicals, generated with Fenton's reagent, rapidly oxidized PCBs (i.e., 2-mono-PCB and the DiCBs through PeCBs present in Aroclor 1242) in aqueous solutions.

The results indicated that the reaction occurs via addition of a hydroxyl group to one nonhalogenated site;reaction rates are inversely related to the degree of chlorination of the biphenyl. The results also indicated that meta and para sites are more reactive than ortho sites due to stearic hindrance effects.

Based upon their kinetic measurements and reported steady-state aqueous system OH concentrations or estimates of OH radical production rates, Sedlak and Andren (1991) estimated environmental half-lives for dissolved PCBs (mono-through octa-PCB) in several water systems as listed below.

Diagram V2 2-7

Estimates for dissolved PCBs in marine surface water are in excess of 1,000 days due to the very low concentration of OH radicals in these waters (10-18M or about two orders of magnitude lower than in freshwater systems).

The results of studies to date indicate that, in the atmosphere, OH oxidation of gas- phase PCBs and PCBs dissolved in cloud water maybe important, although not very fast, degradation mechanisms for PCBs from a global perspective.

However, additional measurements of gas-phase oxidation rates, the ratio of dissolved to sorbed PCBs in cloud water, and OH production and loss rates in cloudwater may provide information that will enable an evaluation of the importance of this mechanism to other degradation mechanisms (Sedlack and Andren, 1991).

2.6.2.3.3. Hydrolysis.
PCBs are unlikely to be affected by hydrolysis under environmental conditions because the attachment of chlorines directly to the aromatic ring in PCBs confers hydrolytic stability. Specifically, SN1 and SN2 reactions do not take place readily at sp2 hybridized carbons (U.S. EPA, 1988; Leifer et al., 1983).

2.6.2.3.4. Biotransformation and Biodegradation.
Leifer et al. (1983) and Brown and Wagner (1990) summarized the available information on the aerobic degradation of PCBs by micro-organisms. Laboratory studies have revealed that there are more than two dozen strains of aerobic, terrestrial micro-organisms widely distributed in the environment that are capable of degrading most PCB congeners with five or fewer chlorines.

In general, the rate of aerobic biodegradation decreases with increasing chlorination. For example, the half-lives for biodegradation of tetra-PCBs in fresh surface water and soil are 7 to 60+ days and 12 to 30 days, respectively. For penta-PCBs and higher chlorinated PCBs, the half-lives in fresh surface water and soil are likely to exceed 1 year. PCBs with all or most chlorines on one ring and PCBs with fewer than two chlorines in the ortho position tend to degrade more rapidly.

Until recent years, little investigation focused on anaerobic microbial dechlorination or degradation of PCBs even though most PCBs eventually accumulate in anaerobic sediments (Risatti, 1992). Environmental dechlorination of PCBs via losses of meta and para chlorines has been reported in field studies for freshwater, estuarine, and marine sediments including those from the Acushnet Estuary, the Hudson River, the Sheboygan River, and Waukegan Harbor (Van Dort and Bedard, 1991). The altered PCB congener distribution patterns found in these sediments (i.e., different patterns with increasing depth or distance from known sources of PCBs) have been interpreted as evidence that bacteria may dechlorinate PCBs in anaerobic sediment.

Results of laboratory studies have also been reported recently. Chen et al. (1988) found that "PCB-degrading" bacteria from the Hudson River could significantly degrade the mono-, di-, and tri-PCB components of a 20 ppm Aroclor 1221 solution within 105 days. These congeners make up 95 percent of Aroclor 1221. No degradation of higher chlorinated congeners (present at 30 ppm or less) was observed, and a separate 40-day experiment with tetra-PCB also showed no degradation.

VanDort and Bedard (1991) reported the first experimental demonstration of biologically-mediated ortho dechlorination of a PCB and stoichiometric conversion of that PCB congener (2,3,5,6-TeCB) to less-chlorinated forms. In that study, 2,3,5,6-TeCB was incubated under anaerobic conditions with unacclimated methanogenic pond sediment for 37 weeks with reported dechlorination to 2,5-DCB (21%), 2,6-DCB (63%), and 2,3,6-TrCB (16%).

Risatti (1992) examined the degradation of PCBs at varying concentrations (10,000 ppm, 1,500 ppm, and 500 ppm) in the laboratory with "PCB-degrading" bacteria from Waukegan Harbor. After 9 months of incubation at 22° C, the 500 ppm and 1,500 ppm samples showed no change in PCB congener distributions or concentrations, thus indicating a lack of degradation. Significant degradation was observed in the 10,000 ppm sediment with at least 20 congeners ranging from TrCBs to PeCBs showing decreases.

Quensen et al (1988) also demonstrated that micro-organisms from PCB-contaminated sediments (Hudson River) dechlorinated most PCBs in Aroclor 1242 under anaerobic laboratory conditions. Aroclor 1242 contains predominantly tri- and tetra-PCBs. Three concentrations of the Aroclor corresponding to 14, 140, and 700 ppm on a sediment dry-weight basis were used.

Dechlorination was most extensive at the 700 ppm test concentration; 53 percent of the total chlorine were removed in 16 weeks, and the proportion of TeCBs through HxCBs decreased from 42 to 12 percent. Much less degradation was observed in the 140 ppm sediment, and no observable degradation was found in the 14 ppm sediment. These results and those of Risatti (1992) suggest that the organism(s) responsible for this dechlorination may require relatively high levels of PCB as a terminal electron acceptor to maintain a growing population.

Quensen et al. (1990) reported that dechlorination of Aroclor 1242, 1254, and 1260 by micro-organisms from PCB-contaminated sediments in the Hudson River and Silver Lake occurred primarily at the meta and para positions; ortho-substituted mono- and di-PCBs increased in concentration. This latter finding is significant because removal of meta and para chlorines from the coplanar PCBs should reduce their toxicity and form less chlorinated congeners that are more amenable to aerobic biodegradation.

2.7. ENVIRONMENTAL FATE - BROMINATED COMPOUNDS

2.7.1. Summary
Although there are no available published studies documenting measured fate rate constants, relatively few studies with measured physical/chemical property data, and few relevant environmental monitoring studies, it is possible to estimate the environmental transport and transformation processes for major BDDs, BDFs, and PBBs using structure activity and property estimation methods. Mill (1989) performed such an assessment and much of what is reported in this section is a summary of that review paper. Also useful are the studies undertaken by Jacobs et al. (1976, 1978) to examine the distribution and fate of PBBs in the environment following the accidental contamination of livestock feed in Michigan in 1973 with the brominated flame retardant, FireMaster BPG. FireMaster BPG (a.k.a., PBB) was found by Jacobs et al. (1976) to be comprised of 2,2',4,4',5,5'-hexabromobiphenyl as the major component, two isomers of pentabromobiphenyl, three additional isomers of hexabromobiphenyl, and two isomers of heptabromobiphenyl.

Mill (1989) concluded that the estimated physical/chemical properties of these compounds indicate they will behave in a similar fashion to their chlorinated analogues. In general, these chemicals are expected to be stable under normal environmental conditions, relatively immobile in the environment, and primarily associated with particulate and organic materials. The only environmentally significant path for destruction is photodegradation. If discharged to the atmosphere, any vapor-phase compounds will probably be rapidly photolyzed. The higher brominated congeners, as their chlorinated counterparts, may be present primarily in a particle-bound rather than gaseous phase. If so, they likely will be more resistant to photolysis and become more widely dispersed in the environment.

Upon deposition onto surfaces, there can be an initial loss due to photodegradation and/or volatilization. Once sorbed onto soils or sediments, however, they are expected to be strongly sorbed with erosion and aquatic transport of sediment the dominant physical transport mechanism. If discharged to water, they are expected to preferentially sorb to solids. Volatilization may also be a significant transport mechanism for nonsorbed chemicals even though they have negligible estimate vapor pressures.

2.7.2. Transport Mechanisms
Little information exists on the environmental transport of BDDs, BDFs, and PBBs. However, the available information on the physical/chemical properties of these compounds and their chlorinated analogs coupled with the body of information available on the widespread occurrence and persistence of the chlorinated analogs in the environment indicates that these compounds are likely to be strongly sorbed by soils, sediments, and other particulate material, and to be resistant to leaching and volatilization.

Jacobs et al. (1978) reported that less than 0.2 percent of 2,2',4,4',5,5'-hexa-PBB (14µg PBB/g of soil) and 2,2',3,4,4',5,5'-hepta-PBB (7µg PBB/g of soil) volatilized from soil incubated for 1 year at 28° C.

2.7.3. Transformation Processes

2.7.3.1. Photodegradation.
Photolysis in the atmosphere appears to be a major potential pathway for loss of BDDs and BDFs based on recent studies by Buser. Buser (1988) studied the photolytic decomposition rates of the following compounds in dilute isooctane solutions and as solid phases on quartz surfaces under sunlight and artificial laboratory illumination: 1,2,3,4-TBDD; 2,3,7,8-TBDD; 2,3,7,8-TBDF; and mono- and dibrominated 2,3,7,8-TCDD and 2,3,7,8-TCDF. Under natural sunlight, estimated half-lives were very short, on the order of minutes.

Solid-phase photolysis was significantly slower, in the range of 7 to 35 hours. The major photolytic pathway was reductive dehalogenation with the formation of lower halogenated or unsubstituted dibenzo-p-dioxins and dibenzofurans. The bromo-chlorodibenzofurans degraded faster than either the brominated or chlorinated congeners. The major pathway of photolysis was debromination to form a chlorinated dibenzofuran.

Mill (1989) used the results obtained by Buser (1988) together with assumptions to overcome the lack of quantum yield data from Buser (1988) to estimate the photolysis half-lives of the three brominated-only compounds tested by Buser (1988). Mill (1989) estimated the following half-lives in water (top 1 meter) and for vapor in air (first kilometer above surface) for clear-sky conditions in mid-summer at 40 degrees north latitude:

Diagram V2 2-8

Lutes et al. (1992a, 1992b) studied the short-term photochemistry of tetra- and penta-BDDs and BDFs sorbed onto airborne soot particles in 25 m3 outdoor Teflon film chambers.

The emissions from controlled burning of polyurethane foam containing polybrominated diphenyl ether flame retardants served as the source of the BDDs and BDFs.

Initial experiments demonstrated that more than 95 percent of the BDDs/BDFs were associated with airborne particulate material; less than 5 percent were in the vapor phase. Particulate phase concentrations of tetra- and penta-CDDs/CDFs were monitored for 3 to 6 hours after introduction of the emissions from the foam burn to the chamber under winter and spring temperatures and sunlight regimes. No significant reduction in concentration was observed. The authors conclude that if photolytic degradation was occurring, then the half-lives are much greater than 3 to 6 hours. Thus, as has been observed with CDDs/CDFs and with solid phase experiments by Buser (1988) on BDDs/BDFs, particulate bound BDDs/BDFs are much less susceptible to photolysis than are gaseous-phase BDDs/BDFs. The estimated half-lives listed below indicate that OH oxidation is probably too slow to compete with photolysis.

Diagram V2 2-9

2.7.3.2. Oxidation
No reaction rate data for OH radicals with gaseous-phase BDDs, BDFs, and PBBs could be located. The low vapor pressures of these compounds make direct measurements very difficult with the current techniques. However, Mill (1989), using a structure activity relationship developed by Atkinson (1987), has estimated the half-lives of OH oxidation for the tetra- through octa- BDDs and BDFs. The estimated half-lives listed below indicate that OH oxidation is probably too slow to compete with photolysis.

2.7.3.3. Hydrolysis
There is no available evidence indicating that hydrolysis would be a significant degradation process for these compounds.

2.7.3.4. Biotransformation and Biodegradation
Although there are no data available concerning the biodegradability of the brominated analogs of CDDs and CDFs, it is expected that these brominated analogs, especially the more halogenated congeners, will be recalcitrant to biodegradation. The limited data available on PBBs (discussed below) indicates recalcitrance.

Jacobs et al. (1976) reported that PBBs are extremely persistent based on the results of aerobic and anaerobic soil incubation studies for 24 weeks with the flame retardant, PBB. Only one major PBB component, a pentabromobiphenyl isomer, showed any significant disappearance; however, Jacobs et al. (1976) were not certain whether the disappearance was due to microbial degradation, to poor soil extraction efficiency, or sorption onto glassware.

Jacobs et al. (1976) also detected components of PBB in soils from a field that had received manure from a PBB-contaminated dairy herd 10 months earlier (quantitative changes in PBB were not possible because no earlier soil samples had been obtained). Additional soil studies by Jacobs et al. (1978) found no degradation of 2,2',4,4',5,5'-hexa-PBB (14µg/25g soil) or 2,2',3,4,4',5,5'-hepta-PBB (7µg/25g soil) after incubation at 28° C for 1 year.

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