2. PHYSICAL AND CHEMICAL PROPERTIES AND FATE 2-1

2.1. INTRODUCTION 2-1

2.2. GENERAL INFORMATION 2-2

2.3. PHYSICAL/CHEMICAL PROPERTY EVALUATION METHODOLOGY 2-4

2.4. PHYSICAL/CHEMICAL PROPERTIES - CHLORINATED COMPOUNDS 2-7

2.4.1. Water Solubility 2-7

2.4.2. Vapor Pressure 2-13

2.4.3. Henry's Law Constant 2-15

2.4.4. Octanol/Water Partition Coefficient 2-16

2.4.5. Organic Carbon Partition Coefficient 2-18

2.4.6. Photo Quantum Yields 2-19

2.5. PHYSICAL CHEMICAL PROPERTIES - BROMINATED COMPOUNDS 2-20

2.6. ENVIRONMENTAL FATE - CHLORINATED COMPOUNDS 2-21

2.6.1. Environmental Fate of Chlorinated Dibenzo-p-dioxins (CDDs) and Chlorinated Dibenzofurans (CDFs) 2-21

2.6.1.1. Summary 2-21

2.6.1.2. Transport Mechanisms 2-21

2.6.1.2.1. Transport Mechanisms in Air 2-21

2.6.1.2.2. Transport Mechanisms in Soil 2-26

2.6.1.2.3. Transport Mechanisms in Water. 2-28

2.6.1.3. Transformation Processes 2-30

2.6.1.3.1. Photodegradation 2-30

2.6.1.3.2. Oxidation. 2-35

2.6.1.3.3. Hydrolysis. 2-36

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 2-38

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 2-45

2.7.3.3. Hydrolysis 2-45

2.7.3.4. Biotransformation and Biodegradation 2-45

This chapter summarizes available information regarding the physical and chemical properties and fate of the CDDs, CDFs, BDDs, BDFs, and coplanar PCBs, with an emphasis on the subset of these chemicals defined as dioxin-like in Chapter 1. Physical/chemical properties addressed in this chapter include melting point, water solubility, vapor pressure, Henry's Law constant, octanol/water partition coefficient, organic carbon partition coefficient, and photochemical quantum yield. Fate and transport processes addressed include photolysis, oxidation, hydrolysis, biodegradation, volatilization, and sorption. Biologically-mediated transport properties (i.e., bioconcentration, plant uptake, etc.) are covered in the companion volume to this report, Volume 3: Site-Specific Assessment Procedures.

Knowledge of physical and chemical properties is essential to understanding and modeling the environmental transport and transformation of organic compounds such as the dioxin-like compounds. The properties most important for understanding the environmental behavior of the dioxin and dioxin-like compounds appear to be water solubility (WS), vapor pressure (VP), octanol/water partition coefficient (K_ow), organic carbon partition coefficient (K_oc), and photochemical quantum yield. The ratio of VP to WS (VP/WS) can be used to calculate the Henry's Law constant (H_c) for dilute solutions of organic compounds. Henry's Law constant is an index of partitioning for a compound between the atmospheric and the aqueous phase (Mackay et al., 1982).

To maximize and optimize the identification of information on the physical/chemical properties of these compounds, a thorough search of the recent literature was conducted. A computer literature search was conducted using the on-line Chemical Abstracts (CA) data base maintained by the Scientific Technical Network (STN). Printed abstracts were obtained and screened, and selected literature were retrieved and critically evaluated. The most definitive value for each physical/ chemical property for each congener was selected. The evaluation method used to select the most definitive physical/chemical property values is detailed in Section 2.3. The property values obtained from the scientific literature are summarized in Appendix A. Sections 2.4 and 2.5 present the property values for the dioxin-like compounds that are considered to be the most definitive. These values are utilized in the modeling equations in the companion volume to this report, Volume 3-Site-Specific Assessment Procedures. Appendix A lists all reported property values for the CDDs, CDFs, and coplanar PCBs. Where technically feasible, estimation procedures have been used to provide values where measured data are not available. For those compounds for which data could not be found and estimates are not appropriate, the field is left blank and a congener group average is presented as the property value for that congener group.

The values suggested in this document as most definitive are, in the authors' opinion, the best values derivable from current data. Since the document has undergone extensive review inside the Agency, by scientific community outside the Agency, and by the Science Advisory Board, the values can be interpreted as generally representative of the Agency and scientific community. The authors recommend that document users consider the values as defaults in the sense that users are encouraged to accept them as a starting point but should feel free to modify them as new data become available.

Brief summaries of the recent and relevant scientific literature on the environmental fate of the polychlorinated and polybrominated dibenzodioxins, dibenzofurans, and biphenyls are provided in Sections 2.6 and 2.7.

Polychlorinated dibenzodioxins (CDDs), polychlorinated dibenzofurans (CDFs), and polychlorinated biphenyls (PCBs) are chemically classified as halogenated aromatic hydrocarbons. CDDs and CDFs can be formed as unintentional by-products through a variety of chemical reactions and combustion processes. Both compound classes have a triple-ring structure that consists of two benzene rings connected by a third oxygenated ring. For CDDs, the benzene rings are connected by a pair of oxygen atoms. CDFs are connected via a single oxygen atom. (See structures below.) PCBs are a class of compounds formed by the chlorination of a biphenyl molecule.

There are 75 possible different positional congeners of CDDs and 135 different congeners of CDFs. Likewise, there are 75 possible different positional congeners of BDDs and 135 different congeners of BDFs. (See Table 2-1.) The basic structure and numbering of each chemical class is shown below.

Table 2-1. Possible Number of Positional CDD (or BDD) and CDF (or BDF) Congeners

There are 209 possible PCB congeners. (See Table 2-1.) The physical/chemical properties of each congener vary according to the degree and position of chlorine substitution. The list of coplanar PCBs can be found in Table 1-2. PCBs assume a coplanar structure when the two benzene rings rotate into a position where the two rings are in the same plane. The PCBs assume a dioxin-like structure when the substituent chlorines occupy the 3, 3', 4, 4', 5, or 5' positions, or possibly, one of the 2 or 2' positions, and the structure is not hindered from assuming the preferred planar configuration. The basic structure and numbering of each chemical class is shown below.

As discussed above, a thorough search of the recent published scientific literature was conducted to maximize and optimize the identification of measured physical/chemical properties. For the purpose of identifying the most definitive of two or more physical/chemical property values reported in the literature for a given dioxin-like compound, a ranking methodology was developed to evaluate the degree of confidence in the reported values. A property value with a ranking of 1 is considered to have the highest level of confidence; a property value with a ranking of 6 is considered to have the lowest level of confidence. The ranking scheme assumes that measured values are more definitive than estimated values. The ranking scheme is based on five ranking criteria or factors. These factors are described below:

Factor 1: Confirmation. Value, measured or derived, confirmed by at least one other laboratory, or different experimental technique. Confirmation was assumed if the reported values were within 50 percent of the highest value (within 5 percent for values reported in logarithmic units).

Factor 2: Measurement Technique. Direct measurement technique used. No measurements reported less than 10 times the method detection limit.

Factor 3: GLP Followed. Good Laboratory Practice was followed in the experimental work. This includes the use of traceable, pure standards; sensitive, selective detection technique was employed; repeatability of measurements demonstrated; all experimental details sufficiently documented so others could reproduce experiments; sources of determinate error considered - error analysis conducted.

Factor 4: Derived Value. Value derived from other directly measured physical/chemical properties by use of known physical/chemical relationships developed for structurally similar chemicals (e.g., other dioxin, furan, and PCB congeners, multiple-ring halogenated compounds). The input value (i.e., the independent variable) used to derive the property value of interest from the equation (i.e., the physical/chemical relationship) is a directly measured value.

Factor 5: Estimated Value. Value estimated using a physical/chemical relationship that was developed using estimated values or a combination of estimated and measured values; this includes QSAR (Quantitative Structure Activity Relationship) methods. Also includes values derived from other directly measured physical/chemical properties by use of known physical/chemical relationships developed, in large part, for structurally dissimilar compounds.

Although this ranking scheme is subjective in nature, it is a reasonable method for identifying the most definitive physical/chemical property value. The ranking scheme has several advantages. First, it identifies where more work is needed to obtain a more definitive p-chem property value. Second, it allows for later adjustments in these values when more definitive studies are conducted. A low ranking for a study does not mean that a particular reported value is incorrect - only that insufficient evidence exists to determine its accuracy. The ranking scheme is as follows:

Rank 1: Confirmed Measured Values. The reported value has met Factors 1, 2, and 3. (See Table 2-2.) This value is considered definitive.

Rank 2: Unconfirmed Measured Values. The reported value has met Factors 2 and 3. The value is considered accurate; it could be definitive subject to confirmation.

Rank 3: Confirmed Derived Values. The reported value has met Factors 1, 3, and 4. The value is considered to be a close approximation.

Rank 4: Unconfirmed Derived Value. The reported value has met Factors 3 and 4. The value is considered to be an approximation.

Rank 5: Estimated Value. The reported value has met Factor 5 only. The value is considered to be an "order-of-magnitude" estimate.

Table 2-2. Ranking Scheme for P-Chem Property Evaluation

Notes: _ indicates all specifications of the Factor have been met.

x indicates the specifications of the Factor have not been met, or the Factor does not apply.

 

If two or more values have the same ranking, then the value that has been peer reviewed by other EPA offices, other government agencies, or scientific data bases (e.g., the Syracuse Research Corporation Environmental Fate Data Bases) and chosen by that office, agency, or data base as the most accurate, was deemed to be the most definitive value for this document. If two or more values with the same ranking have not been peer reviewed as above, typically the most current value was chosen as the most definitive value. This decision was made on the assumption that the most current value would have been developed by the latest scientific method. If two or more values had the same ranking, then some evaluation of the techniques used to derive the value were also considered in choosing the more definitive value. The ranking of the literature can be found in Table A-2 in Appendix A. Table 2-3 lists the property values for the dioxin-like compounds that are considered to be most definitive.

Limited research has been conducted to determine physical and chemical properties of CDFs and CDDs. The congeners having 2,3,7,8-chlorination have received the most attention, with 2,3,7,8,-TCDD being the most intensely studied compound. All 2,3,7,8-substituted congeners are now available commercially, but many of these isomers have not been prepared in pure form. Some of the isomers that have been prepared may not be available in sufficient quantities for testing. Another factor which is likely to have limited research on these compounds is the high toxicity of these compounds, which necessitates extreme precautions to prevent potential adverse effects.

Although water solubility data are not directly used in the exposure scenario equations in Volume 3, water solubility data can be used to estimate Henry's Law constants (using the VP/WS ratio technique) that are used in the equations in Volume 3. Very few measured water solubility values are available in the literature. Marple et al. (1986a) reported the water solubility of 2,3,7,8-TCDD as 19.3 ± 3.7 parts per trillion (nanograms per liter, ng/L) at 22° C. Marple et al. (1986a) used a procedure of equilibrating thin films of resublimed 2,3,7,8-TCDD with a small volume of water followed by gas chromatography (GC) analysis with 63Ni electron capture detection. Other water solubility values for 2,3,7,8-TCDD have been reported in the literature and are summarized in U.S. EPA (1990). Values ranging from 7.9 ng/L to 483 ng/L are reported in U.S. EPA (1990) with 19.3 ng/L selected as the recommended value. The value of 19.3 ng/L was confirmed by Marple et al. (1987) using both radio-labeled and unlabeled 2,3,7,8-TCDD. Marple et al. (1987) reported values of 10.6 ng/L and 10.4 ng/L for the labeled and unlabeled compounds respectively. Because the value of 19.3 ng/L was confirmed by other techniques and was recommended by U.S. EPA (1990), it was chosen as the most definitive value.

Friesen et al. (1985) and Shiu et al. (1988) used HPLC generator column techniques to measure the water solubilities of a series of chlorinated dioxins (1,2,3,4-, 1,2,3,7-, and 1,3,6,8-TCDD; 1,2,3,4,7-PeCDD; 1,2,3,4,7,8-HxCDD; 1,2,3,4,6,7,8-HpCDD; and OCDD). Reported water solubilities ranged from 320 ng/L to 0.074 ng/L for the 1,2,3,7-TCDD and OCDD congeners, respectively. The only congener with more than one value was OCDD. The value of 0.074 ng/L (Shiu et al., 1988) was chosen because it was the most current. Friesen et al. (1990) used a gas chromatography/mass spectrometry detection (GC/MSD) generator column technique to measure the water solubilities of a series of chlorinated furans (2,3,7,8-TCDF; 2,3,4,7,8-PeCDF; 1,2,3,6,7,8- and 1,2,3,4,7,8-HxCDF; and 1,2,3,4,6,7,8-HpCDF) and reported a decrease in water solubility with an increase in the number of chlorine substituents. The reported water solubility values ranged from 1.37 x 10^-9 mol/L (419 ng/L) for the 2,3,7,8-TCDF isomer to 3.30 x 10^-12 mol/L (1.35 ng/L) for the 1,2,3,4,6,7,8-HpCDF congener. The dioxin-like furans only had one value reported for water solubility.

Values for the various congener groups ranged as follows: TCDDs 0.47-596 ng/l, PeCDDs 120-166 ng/l, HxCDDs 4.4 ng/l, HpCDDs 2.4 ng/l, OCDD 0.074-0.4 ng/l, TCDFs 4.2 ng/l, PeCDFs 236 ng/l, HxCDFs 8.2-17.7 ng/l, HpCDFs 1.35 ng/l, and OCDF 1.16 ng/l. The range for CDDs covers nearly four orders of magnitude, and for the CDFs two orders of magnitude.

The reported water solubility values for the coplanar PCB compounds are comparable to those for the CDD and CDF compounds. The reported values range from 11,400 ng/L for 3,3',4,4'-TeCB to 0.74 ng/L for 2,3,3',4,4',5-HxCB. Measured water solubility data chosen as the most definitive were those reported by Dunnivant and Elzerman (1988) and Murphy et al. (1987). The value of 549 ng/L (Dunnivant and Elzerman, 1988) for 3,3',4,4'-TCB was confirmed by Dickhut et al. (1986) with a value of 569 ng/L. Therefore, 549 ng/L was chosen as the most definitive value for 3,3',4,4'-TCB. Murphy et al. (1987) proved the only measured values for the other congeners.

For those compounds without reported measured water solubility values, estimations were calculated by the congener group-average method. For example, for the tetra-chlorinated dioxins, values reported in the literature were averaged to yield an estimated water solubility value for the tetra-chlorinated dioxin congener group. A similar procedure was used to develop the average value for each of the other CDD and CDF congener groups. The most definitive value for each isomer was used to derive the congener group average. Estimating the water solubility values from measured log K_ow values using the estimation procedure of Lyman et al. (1982) did not yield satisfactory results; the estimated water solubilities for 2,3,7,8-TCDD and 1,3,6,8-TCDD were at least two orders of magnitude greater than the measured values in Tables 2-3 and A-1. Compounds that have water solubility values in the ranges reported for these chlorinated compounds are considered to have very poor solubility in water.

Vapor pressure data are not directly used in the exposure scenario equations in Volume 3. However, vapor pressure data can be used to estimate Henry's Law constant using the VP/WS ratio technique. Very few measured vapor pressure values are available in the literature for the CDDs and CDFs. The majority of the measured vapor pressures are for the 2,3,7,8-substituted compounds.

U.S. EPA (1990) presented the range of measured vapor pressure data for 2,3,7,8-TCDD and selected a recommended value of 7.4 x 10^-10 mm Hg at 25° C. This value was reported by Podoll et al. (1986) who used radiolabeled 2,3,7,8-TCDD and a gas saturation technique with combustion to ¹⁴CO₂. Rordorf (1987, 1989) reported a higher vapor pressure value for 2,3,7,8-TCDD, 1.49 x 10^-9mm Hg. SRC (1991) reported this same value by extrapolating the vapor pressures measured by Schroy et al. (1985) at four higher temperatures, 30° , 55° , 62° , and 71° C. The value recommended in U.S. EPA (1990) is reported in Table 2-3.

Rordorf (1987, 1989) reported experimental vapor pressure values for 1,2,3,4-TCDD (4.8 x 10^-8 mm Hg), OCDD (8.25 x 10^-13mm Hg), and OCDF (3.75 x 10^-12 mm Hg) (Table A-1). These values were chosen as the most definitive because they were the most current directly measured values. Rordorf (1987, 1989) used a gas-flow method in a saturation oven, with integrated gas chromatographic analysis, to measure vapor pressure values for ten CDDs and four CDFs. Rordorf (1987, 1989) also used a vapor pressure correlation method to predict the vapor pressures of 15 other CDDs and 55 CDFs based on the measured vapor pressures for the 10 CDDs, 4 CDFs, and the deduced boiling point and enthalpy data for the larger series of CDDs and CDFs. Measured boiling point and enthalpy data are in good agreement with the deduced data used in the correlation method. Of the CDDs studied by Rordorf (1987, 1989), only three of the ten, 1,2,3,4-TCDD, 2,3,7,8-TCDD, and OCDD, are in the dioxin-like compound group of chemicals studied in this report. The other CDDs with measured values were monochloro-, dichloro-, and trichloro-dibenzo-p-dioxins.

Eitzer and Hites (1988) reported experimental vapor pressure values for several of the dioxin-like compounds utilizing GC capillary column retention time data. The values were reported as subcooled liquids and then converted to solid-phase vapor pressures. The solid-phase vapor pressures ranged from 2.16 x 10^-12 mm Hg to 9.48 x 10^-10 mm Hg for the CDDs and from 1.07 x 10^-10 mm Hg to 8.96 x 10^-9 for the CDFs. The values from Eitzer and Hites (1988) were considered the most definitive, except for OCDD, because they were the only values that were derived (i.e., Rank 4); all other values were estimated (i.e., Rank 5).

Values reported for the congeners within various congener groups are as follows:

Congener Group Vapor Pressure Range (mm Hg)

TCDD 7.4 x 10^-10 to 4.03 x 10^-6

PeCDD 4.35 x 10^-10 to 9.48 x 10^-10

HxCDD 3.60 x 10^-11 to 1.01 x 10^-10

HpCDD 5.62 x 10^-12 to 3.21 x 10^-11

OCDD 8.25 x 10^-13 to 6.54 x 10^-8

TCDF 8.96 x 10^-9 to 3.98 x 10^-8

PeCDF 1.50 x 10^-9 to 4.28 x 10^-9

HxCDF 1.80 x 10^-10 to 5.70 x 10^-10

HpCDF 3.53 x 10^-11 to 1.33 x 10^-10

OCDF 3.75 x 10^-12

The range for CDDs covers over six orders of magnitude, and for the CDFs, four orders of magnitude.

The measured vapor pressure values reported for the coplanar PCBs are comparable to those reported for the CDD and CDF compounds; the estimated values are higher by several orders of magnitude. (See Table 2-3.) The directly measured values of Murphy et al. (1987) and the derived value of Dunnivant and Elzerman (1988) were considered the most definitive. All other values were estimated. The values reported in Tables 2-3 and A-1 by Foreman and Bidleman (1985) are an average of the 0V-101 RI and Dexsil 410 RI correlation methods because both methods were determined to be equally valid. As with the other groups, the vapor pressures of the PCBs decrease with an increase in the number of chlorine substituents. The highest reported value for the coplanar PCBs is 2.90 x 10^-6 mm Hg for 3,3',4,4',5-PeCB, and the lowest value reported is 2.80 x 10^-10 mm Hg for 3,4,4',5-TeCB.

Estimated vapor pressure values for those CDDs and CDFs for which measured values were not found in the literature were calculated by the congener group-average method using the literature-reported values within a congener group. For example, the literature values for the TCDDs were averaged to obtain an estimated vapor pressure assumed to apply to the TCDD congeners that did not have literature values. A similar procedure was used to develop a congener-average for each of the other congener groups. The most definitive value for each isomer was used to derive the congener group average. Compounds with vapor pressures in the ranges reported for these compounds are considered to have very low vapor pressures.

Henry's Law constant data are used in Volume 3 to estimate the volalitization of the dioxin-like compounds from soil. They are also utilized in estimating the vapor-phase bioconcentration factor from air to plant leaves. Directly measured data for Henry's Law constant have been reported for only two compounds, one TCDD congener, and one PCB congener. The measured values for 1,3,6,8-TCDD, 6.81 x 10^-5 atm-m³/mol (Webster et al., 1985), and for 3,3',4,4'-PCB, 9.4 x 10^-5 atm-m³/mol (Dunnivant and Elzerman, 1988) were determined by the gas-purging technique. These two values were considered the most definitive. Other values reported in the literature for CDDs, CDFs, and PCBs were calculated by the vapor pressure/water solubility (VP/WS) ratio technique or by structure-activity relationship techniques. A derived VP/WS ratio value, Rank 4, was determined to be more definitive than an estimated value, Rank 5.

Group-average Henry's Law constants were estimated for each congener group based on the reported data for that group. The Henry's Law constant values for the PCBs are similar to those for the CDDs and CDFs.

Lyman et al. (1982) offers guidelines, though not specific to these compounds, for comparing the degree to which organic compounds volatilize from water. These guidelines suggest that volatilization of polycyclic aromatic hydrocarbons and halogenated aromatics (which includes all the dioxin-like compounds) from water represents a significant transfer mechanism from the aqueous to the atmospheric phase.

The octanol/water partition coefficient is used in several exposure estimation procedures in Volume 3. It is used to estimate log K_oc when measured data are not available, and it is utilized in estimating the root concentration factor (RCF). The RCF is used to estimate the uptake of contaminants by plant roots. Log K_ow is also used to estimate the vapor-phase bioconcentration factor from air to plant leaves.

Marple et al. (1986b) reported the octanol/water partition coefficient of 2,3,7,8-TCDD as 4.24 (± 2.73) x 10⁶ at 22 ± 1° C, yielding a log K_ow of 6.64 (Table A-1). Two similar experimental techniques were used, but the more reliable method involved equilibration of water-saturated octanol, containing the 2,3,7,8-TCDD, with octanol-saturated water, over 6 to 31 days. U.S. EPA (1990) reported that the available low K_ow data ranged from 6.15 to approximately 8.5. The 6.64 value reported by Marple et al. (1988b) was the value recommended in that report. The 6.64 value was confirmed by Sijm et al. (1989) with a value of 6.42, and by Marple et al. (1987) with a value of 6.69; therefore, a log K_ow value of 6.64 was considered the most definitive.

Burkhard and Kuehl (1986) used reverse-phase High Pressure Liquid Chromatography (HPLC) and Liquid Chromatography/Mass Spectrometry (LCMS) detection to determine octanol/water partition coefficients for 2,3,7,8-TCDD and a series of seven other tetrachlorinated planar molecules, including three other TCDD isomers (1,2,3,4-TCDD; 1,3,7,9-TCDD; 1,3,6,8-TCDD), 2,3,7,8-TCDF, and 3,3',4,4'-tetrachlorobiphenyl. The log K_ow values for the four TCDD isomers ranged from 7.02 to 7.20. The log K_ow for 2,3,7,8-TCDF was 5.82, and the log K_ow for 3,3',4,4'-TCB was 5.81.

Burkhard and Kuehl (1986) also re-evaluated data on 13 CDDs and CDFs previously reported by Sarna et al. (1984) under similar experimental techniques. In the re-evaluation, Burkhard and Kuehl (1986) used experimental rather than estimated log K_ow values in correlations with gas chromatographic retention times. This approach yielded log octanol-water partition coefficients ranging from about 4.0 for the nonchlorinated parent molecules to about 8.78 for the octa-chlorinated compounds, much lower than the values originally reported by Sarna et al. (1984).

Sijm et al. (1989) used a slow stirring method to obtain log K_ow values for 73 CDD and CDF congeners. Values ranged from 6.10 to 7.92.

The most definitive values chosen were either a directly measured value or the most current derived value. Only 2,3,7,8-TCDD had more than one directly measured value.

Values reported for congeners within the various congener groups ranged as follows:

Congener Group Octanol/Water Partition Coefficient

TCDD 5.91 to 8.84

PeCDD 6.2 to 9.69

HxCDD 6.85 to 10.55

HpCDD 8.2 to 11.54

OCDD 7.46 to 8.6

TCDF 5.6 to 6.73

PeCDF 6.19 to 6.92

HxCDF not available

HpCDF 7.92

OCDF 7.05 to 13.35

The range for the CDDs covers nearly six orders of magnitude, and for the CDFs nearly seven orders of magnitude.

The measured and literature-estimated log K_ow values for the PCBs are similar to those reported for the CDDs and CDFs. The values range from 5.62 (measured) for 3,3',4,4'-TeCB to 7.71 (literature-estimate) for 2,3,3',4,4',5,5'-HpCB. The log K_ow values increase with an increase in the number of chlorine substituents. The log K_ow for the 3,3',4,4'-TeCB was measured by Hawker and Connell (1988) using the generator column technique against the linear relationship of relative retention time on a nonselective gas chromatograph stationary phase. This was the only directly measured log K_ow for the PCBs; therefore, it was considered the most definitive. Log K_ow values for the HxCBs were measured by Risby et al. (1990) using a high-performance liquid chromatographic (HPLC) system. The values reported in Tables 2-3 and A-1 are an average of the two techniques because both methods were determined to be equally valid. The values for the HxCBs were ranked 3 because both methods produced similar values. The most definitive values for the other PCBs were either derived values or the most current estimated value.

Partition coefficient values were calculated for those compounds for which no measured data were reported in the literature by averaging the literature values within congener groups, as were done for vapor pressure and water solubility. Literature values for the hexachlorodibenzofurans could not be found; thus, no congener group average could be calculated. Partition coefficients in the ranges of these reported values indicate that the substances tend to adsorb strongly to organic components in the soil and may bioconcentrate in those organisms exposed to the compounds.

The organic carbon partition coefficient (K_oc) is used in several exposure estimations in Volume 3. K_oc is used in the estimation of the adsorption partition coefficient, which describes the partitioning of contaminants between suspended sediment and the water column. K_oc is also used in estimating the concentration of contaminants in below ground vegetables grown in contaminated soil.

Measured log K_oc values could be found for 2,3,7,8-TCDD in five studies. Lodge and Cook (1989) used contaminated sediments from Lake Ontario and distilled water in glass cylinders to measure the log K_oc of 2,3,7,8-TCDD. Log K_oc values ranged from 7.25 to 7.59. Jackson et al. (1986) used 10 contaminated soil samples in a batch extraction procedure to measure log K_oc. The average log K_oc of the 10 soils was reported as 7.39. Marple et al. (1986) used two uncontaminated soils spiked by two different methods with 2,3,7,8-TCDD to obtain the log K_oc value. The soil was stirred with water in 2-liter flasks. The log K_oc values ranged from 5.96 to 6.54 for both soils, with an average value of 6.40 for the red clay soil and 6.02 for the alluvial soil.

Puri et al., (1989) studied log K_oc of 2,3,7,8-TCDD with several other co-contaminants such as crankcase oils and surfactants. An average log K_oc value of 5.68 was reported for 2,3,7,8-TCDD in the presence of 0.01 percent surfactant. Walters and Guiseppi-Elie (1988) used several soils and water/methanol mixtures in a batch shake testing procedure to determine the log K_oc of 2,3,7,8-TCDD. The study resulted in a log K_oc value of 6.6.

Four studies for log K_oc of 2,3,7,8-TCDD were ranked number 1. The studies by Jackson et al. (1986) and Lodge and Cook (1989) had confirming values of 7.39 and 7.42, respectively. The studies by Walters and Guiseppi-Elie (1988) and Marple et al. (1987) had confirming values of 6.6 and 6.4, respectively. The 6.6 value reported by Walters and Guiseppi-Elie (1988) was chosen by Syracuse Research Corporation (SRC) in the CHEMFATE Database (SRC, 1991) as the most definitive. This value was determined in a mixed solvent system, water and methanol; therefore, it is not considered as appropriate as a pure water equilibration system determined value. The confirming value by Marple et al. (1987), 6.4, was determined in uncontaminated soil and with pure water; therefore, this value is considered the most definitive for this document.

Webster et al. (1986) used a modified generator column technique to measure the organic carbon partition coefficients of three dioxin-like compounds. Three dissolved humic substances were introduced into the carrier stream to measure the interaction between the contaminants and organic matter. The reported average values at 20° C were 5.97 for 1,2,3,7-TCDD, 5.68 for 1,2,3,4,7-PeCDD, and 5.92 for 1,2,3,4,7,8-HxCDD. Only one value was found for the dioxin-like PCBs, 5.7 for 2,3',4,4',5-PeCB (EPRI, 1990). Compounds that have log K_oc values in the ranges reported for these chemicals are expected to strongly sorb to particulate matter.

Photo quantum yields, which can be used to estimate the rate of photolysis in the environment, have been reported for only nine congeners:

Congener Photo Quantum Yield Reference

1,2,3,7-TCDD 5.42 x 10^-4 (Choudhry and Webster, 1989)

1,3,6,8-TCDD 2.17 x 10^-3 (Choudhry and Webster, 1989)

2,3,7,8-TCDD 2.2 x 10^-3 (Dulin et al., 1986)

2,3,7,8-TCDD 3.3 x 10^-2 (Rapaport and Eisenreich, 1984)

1,2,3,4,7-PeCDD 9.78 x 10^-5 (Choudhry and Webster, 1987)

1,2,3,4,7,8-HxCDD 1.10 x 10^-4 (Choudhry and Webster, 1987)

1,2,3,4,6,7,8-HpCDD 1.53 x 10^-5 (Choudhry and Webster, 1987)

OCDD 2.26 x 10^-5 (Choudhry and Webster, 1987)

1,2,4,7,8-PeCDF 1.29 x 10^-2 (Choudhry et al., 1990)

1,2,3,4,7,8-HxCDF 6.96 x 10^-4 (Choudhry et al., 1990)

All quantum yields were measured in a water-acetonitrile solution at 313 nm, except those reported by Rapaport and Eisinreich (1984) which were measured in the vapor phase at 250-360 nm. No values were found for the PCBs.

Homologue group averages were not calculated because photo quantum yields are very sensitive to chlorine position and the solvent system used in the experiments. Different water to acetonitrile volume ratios were used in these experiments.

Information on the physical and chemical properties of the polybrominated dioxins and furans is very limited. Dr. G. R. B. Webster, University of Manitoba is expected to will publish measured results for testing with brominated compounds in the near future.

2.6. ENVIRONMENTAL FATE - CHLORINATED COMPOUNDS

2.6.1. Environmental Fate of Chlorinated Dibenzo-p-dioxins (CDDs) and Chlorinated Dibenzofurans (CDFs)

The growing body of literature from laboratory, field, and monitoring studies examining the environmental transformation and environmental distribution of CDDs and CDFs has increased the understanding of the fate of these environmentally ubiquitous compounds. In soil, sediment, the water column, and probably air, CDDs/CDFs are primarily associated with particulate and organic matter because of their high lipophilicity and low water solubility. They exhibit little potential for significant leaching or volatilization once sorbed to particulate matter. The available evidence indicates that CDDs and CDFs, particularly the tetra- and higher chlorinated congeners, are extremely stable compounds under most environmental conditions. The only environmentally significant transformation process for these congeners is believed to be photodegradation of nonsorbed species in the gaseous phase or at the soil or water-air interface. CDDs/CDFs entering the atmosphere are removed either by photodegradation or by dry or wet deposition. Burial in-place or erosion of soil to water bodies appears to be the predominant fate of CDDs/CDFs sorbed to soil. CDDs/CDFs entering the water column primarily undergo sedimentation and burial. The ultimate environmental sink of CDDs/CDFs is believed to be aquatic sediments.

2.6.1.2.1. Transport Mechanisms in Air. Once released into the atmosphere, CDDs and CDFs become widely dispersed throughout the environment by atmospheric transport and deposition. In a recent assessment of the atmospheric transport and deposition of CDDs and CDFs for EPA, Hites and Harless (1991) generated data and analyses that support the contention that background environmental levels and congener profiles of CDDs and CDFs in soils and sediment (i.e., higher rather than lower chlorinated congener patterns predominate) can be attributed, in large part, to the atmospheric transport and transformation of CDDs and CDFs released from combustion sources.

Hites and Harless (1991) showed that during transport there is partitioning between the vapor and particle-bound phases. The two key parameters controlling the phase in which a particular congener is found are the congener's vapor pressure and the atmospheric temperature. Congeners with higher vapor pressures are found to a greater extent in the vapor phase. A comprehensive evaluation of the partitioning of dioxin-like compounds between vapor and particle phases was performed in Volume III of this three-volume document. Ambient air monitoring studies that examined the partitioning of dioxin-like compounds between vapor and particle phases were summarized in the Volume III evaluation. A theoretical approach developed by Bidleman (1988) was also discussed, and this approach was used to model the vapor/particle (V/P) partitioning for purposes of evaluating the impact of stack emissions. Table 2-4 summarizes the V/P partitioning reported in several ambient air monitoring studies and also the V/P partitioning estimated by the Bidleman (1988) model. The results are presented as V/P ratios (i.e., the ratio of the concentration of a compound in the vapor phase to the concentration of that compound in the particulate phase on a volume-to-volume basis). From the review in Volume III, the following conclusions were made:

· Ambient air sampling methods do give an approximate indication of the V/P ratio that seems to be responsive to changes in temperature and to the degree of chlorination of the CDDs/CDFs. This is in accordance with what would be expected from their individual vapor pressures. The methods present a realistic picture of partitioning under variable ambient conditions. However, the method has certain limitations that currently prevent deriving a true measurement of V/P partitioning in the ambient air. First, the glass fiber filter is designed to capture and retain particulate matter greater than or equal to 0.1 µm diameter. Particles less than this diameter may pass through the filter and be retained in the polyurethane foam vapor trap downstream. If this is the case, the amount of CDDs/CDFs observed to be particle-bound would be underestimated, and the amount observed to be in vapor phase would be overestimated. Second, the relatively high volume of sampled air passing through the system (200 to 400 m³ of air per 24 hours) may redistribute the more volatile congeners from the filter to the adsorbent trap by a process known as 'blow-off'. Again, this would lead to an overestimate of the fraction in the vapor phase.

· The theoretical construct relies on current adsorption theory, considers the molecular weight and the degree of halogenation of the congeners, uses the boiling points and vapor pressures of the congeners, and uses the availability of surface area on atmospheric particles for adsorption that correspond to a variety of ambient air shed classifications having variable particulate matter densities. Four air shed classifications are described in Bidleman (1988): "clean continental", "background", "background plus local sources", and "urban". The classification used in Volume III for evaluating impacts in a rural environment is "background plus local sources". It is noted from Table 2-4 that the V/P ratios determined theoretically indicate less compound in the vapor phase (or equivalently, more in the particle phase) than is reported in the monitoring studies. This is consistent with the discussion above suggesting that the ambient air instrumentration could overestimate the vapor fraction because of instrumentation design and performance.

 

Towara et al. (1993) studied the particle size distribution of atmospheric particle-bound CDD/CDFs. Three 48-hour samples were collected in a rural area of Germany during the summer of 1992. Particles with aerodynamic diameters of less than 1.35 m m (i.e., particles that have relatively long residence times in the atmosphere) accounted for 65, 84, and 82 percent of the total particle mass in the three samples. However, these small particles accounted for 91, 90, and 85 percent of the total mass of CDD/CDFs found in all particle sizes combined.

CDDs and CDFs are removed physically from the atmosphere by wet deposition (i.e., scavenged by precipitation), particle dry deposition (i.e., gravitational settling of particles) and gas-phase dry deposition (i.e., sorption of CDD/CDFs in the vapor phase onto plant surfaces) (Rippen and Wesp, 1993; Welsch-Pausch et al., 1993). Precipitation can be very effective in removing CDDs and CDFs from the atmosphere. Listed in Table 2-5 are the average precipitation scavenging ratios for congener groups reported by Hites and Harless (1991) and Koester and Hites (1992a) for Bloomington, Indiana, and Indianapolis, Indiana, respectively. The scavenging ratio is the ratio of the concentration of a chemical in precipitation (rain in these studies) to the concentration in the atmosphere and is a measure of the effectiveness of rain in removing the chemical. Also listed in Table 2-5 are the percentages of congener groups scavenged as particles in rain rather than as dissolved solutes in rain. Total rain scavenging ratios ranged from 10,000 to 150,000; hepta- and octa- CDDs (i.e., the congeners most strongly associated with particulates) were scavenged most efficiently.

As part of their studies, Hites and Harless (1991) and Koester and Hites (1992a) also measured dry deposition of CDDs and CDFs and calculated wet and dry deposition fluxes to determine which process dominated CDD/CDF deposition. The calculated wet deposition flux for both cities was similar; 220 ng/m²-yr for Indianapolis and 210 ng/m²yr for Bloomington as might be expected based on similar rainfall patterns. However, the dry deposition fluxes differed by a factor of two between the cities (160 ng/m²-yr for Bloomington and 320 ng/m²-yr for Indianapolis). Wet deposition was calculated to be the dominant process for Bloomington; whereas, dry deposition was calculated to be the dominant process for Indianapolis. The difference was attributed to the higher total suspended particulate matter in Indianapolis air.

2.6.1.2.2. Transport Mechanisms in Soil. Upon deposition of CDDs/CDFs onto soil or plant surfaces, there can be an initial loss due to photodegradation and/or volatilization. The extent of initial loss due to volatilization and/or photodegradation is uncertain and may be controlled by climatic factors, soil characteristics, and the concentration and physical form of the deposited CDDs/CDFs (i.e., particulate-bound, dissolved in solvent, etc.) (Freeman and Schroy, 1989; Paustenbach et al., 1992). For example, observations from the Seveso incident indicated that when 2,3,7,8-TCDD was deposited on the soil surface, the levels in the surface soil decreased substantially in the first 6 months (DiDomenico et al., 1982). Similarly, Nash and Beall (1980) reported that 12 percent of the 2,3,7,8-TCDD applied to bluegrass turf as a component (7.5 ppm concentration) of an emulsifiable Silvex concentrate volatilized over a period of nine months.

Because of their very low water solubilities and vapor pressures, CDDs/CDFs below the soil surface (i.e., below the top few millimeters) are strongly adsorbed and show little upward or downward vertical migration, particularly in soils with a high organic carbon content (Yanders et al., 1989). Freeman et al. (1987) found no statistically meaningful changes in the concentration profile of 1,2,7,8-TCDD in the top 1 cm of Time Beach Soil over a 16-month period, with the exception of the top 3mm of soil exposed to water and sunlight in which 50 percent reduction in 2,3,7,8-TCDD concentration was observed. In addition, the more chlorinated congeners do not show any significant degree of degradation below the soil surface. Although for several years it was believed that near-surface (i.e., the top 1cm) CDDs/CDFs could volatilize slowly to the surface (Freeman and Schroy, 1985), recent research has indicated that CDDs/CDFs, particularly the tetra and higher chlorinated congeners, show little or no movement upward or downward in the subsurface unless a carrier such as waste oil or diesel fuel is present to act as a solvent. For example, Palausky et al. (1986) injected 2,3,7,8-TCDD dissolved in various organic solvents into soil columns to determine the extent of vapor phase diffusion; little movement due to volatilization was observed unless the soil was incubated at 40° C.

Paustenbach et al. (1992) reviewed many major published studies on dioxin persistence in soil and concluded that 2,3,7,8-TCDD probably has a half-life of 25 to 100 years in subsurface soil and 9 to 15 years at the soil surface (i.e., the top 0.1 cm). Several major studies reviewed by Paustenbach et al. (1992) and additional recent studies are summarized below. Some of these recent studies have concluded that the binding of dioxin-like compounds to soil approaches irreversibility over time due to the encapsulation of the compounds in soil organic and mineral matter (Puri et al., 1989; Puri et al., 1992).

Orazio et al. (1992) studied the persistence of di- to octa-chlorinated CDDs and CDFs in sandy loam soil held in laboratory columns under water-saturated soil conditions for a period of 15 months. Measurable upward movement was reported only for the dichlorofurans and dioxins. Downward movement was only noticeable for the dichloro- and trichloro-congeners. The mobility of the CDDs and CDFs was not significantly affected by co-contaminants (i.e., pentachlorophenol and creosote components) present at concentrations as high as 6,000 mg/kg. As much as 35 percent loss of the di- and trichloro-congeners due to degradation was observed; no significant degradation of the tetra- through octa-chlorinated congeners was reported (Orazio et al., 1992).

Hagenmaier et al. (1992) collected soil samples around two industrial plants in Germany in 1981, 1987, and 1989 at the same site and from the same depth, using the same sampling method. There was no indication (within the limits of analytical accuracy (+/- 20 percent)) of appreciable loss of CDDs and CDFs by vertical migration, volatilization, or degradation over the 8-year period. Also there were no significant changes in the congener distribution pattern (i.e., tetra- through octa-) over this time period.

Yanders et al. (1989) reported that 12 years after oil containing 2,3,7,8-TCDD was sprayed on unpaved roads at Times Beach, Missouri, no dioxin was discovered deeper than 20 cm. However, these roads were paved about 1 year after the spraying episode, thus preventing volatilization to the atmosphere. Yanders et al. (1989) excavated this soil and placed the soil in bins located outdoors, subject to the natural conditions of sunlight and precipitation. They reported no appreciable loss nor vertical movement of 2,3,7,8-TCDD from the soil, even in the uppermost sections, during a 4-year study period. Puri et al. (1992) reported no migration or loss of 1,2,3,4-TCDD, 1,2,3,7,8-PeCDD, OCDD, and OCDF from samples of this soil which were examined for 2 years in controlled laboratory column experiments.

Hallett and Kornelson (1992) reported finding 2,3,7,8-TCDD at levels as high as 20 pg/g in the upper 2 inches of soil obtained from areas of cleared forest in New Brunswick, Canada, where the pesticides 2,4-D and 2,4,5-T had been applied in one or more applications 24 to 33 years earlier.

Pereira et al. (1986) reported contamination by CDDs of the sand and gravel aquifer underlying unlined surface impoundments at a wood-treatment facility that had utilized creosote and pentachlorophenol. CDDs migrated both vertically and horizontally in the subsurface. Puri et al. (1992), using soil column experiments in the laboratory, demonstrated that pentachlorophenol and naphthalene and methylnaphthalene (components of creosote) readily transported CDDs/CDFs through soil. Puri et al. (1989) and Kapila et al. (1989) demonstrated that application of waste oil and anionic surfactant solutions to field and laboratory columns of Times Beach soil can move 2,3,7,8-TCDD through soil. Walters and Guiseppe-Elie (1992) showed that methanol/water solutions (1g/L or higher) substantially increase the mobility of 2,3,7,8-TCDD in soils.

Although few studies have evaluated quantitatively the transport of soil-bound CDDs/CDFs, the very low water solubilities and high K_ocs of these chemicals indicate that erosion of soil to water bodies appears to be the dominant surface transport mechanism for CDDs/CDFs sorbed to soil (Paustenbach et al., 1992).

2.6.1.2.3. Transport Mechanisms in Water. Most CDDs/CDFs entering the aquatic environment are associated with particulate matter (e.g., dry deposition of atmospheric particles and eroded soil) and are likely to remain sorbed to the particulate matter once in the aquatic environment. Recent studies have demonstrated that dissolved CDDs/CDFs entering the aquatic environment will, like other lipophilic, low water solubility organic compounds, partition to suspended solids or dissolved organic matter such as humic substances.

Muir et al. (1992) and Servos et al. (1992) recently reported that 48 hours after the addition of 2,3,7,8-TCDF, 1,3,6,8-TCDD, and OCDD in a sediment slurry to natural lake water/sediment limnocorrals, between 70 and 90 percent had partitioned to suspended particulates. The proportion freely dissolved in water ranged from <2 percent for 2,3,7,8-TCDF and OCDD to 10 to 15 percent for 1,3,6,8-TCDD. The remainder was associated with dissolved organic substances.

Broman et al. (1992) analyzed water collected from nine sampling points in the Baltic Sea selected to be representative of background levels. The concentration of particle-associated (>0.45mm) total CDDs/CDFs varied between 0.170 and 0.390 pg/L with an average concentration of 0.230 pg/L (or 66 percent of total CDDs/CDFs). The total CDD/CDF concentration of the "apparently" dissolved fraction varied between 0.036 and 0.260 pg/L with an average concentration of 0.120 pg/L (or 34 percent of the total). Subsequent calculations estimated that, on average, only 0.070 pg/L of the "apparently" dissolved CDDs/CDFs were truly dissolved.

The dominant transport mechanism for removal of CDDs/CDFs from the water column is believed to be sedimentation and ultimately burial in sediments; sediment resuspension and desorption of CDDs/CDFs will vary on a site-by-site basis. Servos et al. (1992) reported that the 1,3,6,8-TCDD and OCDD added as a sediment slurry to lake limnocorrals rapidly partitioned/settled to surficial sediments where they persisted over the 2 years of the study. The half-lives of 1,3,6,8-TCDD and OCDD in the water column were reported as 2.6 and 4.0 days, respectively. Based on sediment trap and mixed surface layer studies of the Baltic Sea, Broman et al. (1992) report that the mass of CDDs/CDFs in the mixed surface layer at any moment represents about 1 percent of the total flux of CDDs/CDFs to the sediment annually; this implies little recirculation of these compounds within the water column of the Baltic Sea. Broman et al. (1992) also reported that the concentration of CDDs/CDFs in settling solids (i.e., sediment trap collected material) is approximately one order of magnitude greater than the concentration in suspended particulates. They attributed this elevated concentration to the capacity of settling solids to scavenge the dissolved fraction as the solids settle through the water column. Similar findings have been reported elsewhere (e.g., Baker et al., 1991) for PCBs and PAHs in the Great Lakes.

Even though they possess very low vapor pressures, CDDs/CDFs can volatilize from water. However, volatilization is not expected to be a significant loss mechanism for the tetra- and higher chlorinated CDDs/CDFs from the water column under most non-spill scenarios. Podoll et al. (1986) calculated volatilization half-lives of 15 days and 32 days for 2,3,7,8-TCDD in rivers and ponds/lakes, respectively. Broman et al. (1992) used measured concentrations of CDDs/CDFs in ambient air (gaseous phase) and in Baltic Sea water (truly dissolved concentrations) to calculate the fugacity gradient over the air-water interface. The fugacity ratios obtained indicated a net transport from air to water (ratios between 0.4 and 0.004).

Fish and invertebrates bioaccumulate CDDs/CDFs, although the benthic and pelagic pathways by which the accumulation occurs are not well understood. Organisms have been shown to accumulate CDDs/CDFs when exposed to contaminated sediments and also to bioconcentrate CDDs/CDFs dissolved in water. However, since most of the CDDs/CDFs in the water column and sediment are associated with particulate matter and dissolved organic matter, the accumulation observed in the environment may be primarily food chain-based starting with uptake by benthic organisms (e.g., mussels, chironomids) directly from sediment pore waters and/or by ingestion or filtering of contaminated particles. Those organisms consuming benthic organisms (e.g., crayfish, suckers) would then pass the contaminants up the food chain (Muir et al., 1992).

2.6.1.3.1. Photodegradation. Photodegradation appears to be the most environmentally significant degradation mechanism for CDDs/CDFs in water, air, and soil. CDDs/CDFs absorb electromagnetic radiation at wavelengths greater than 290 nm (i.e., the lower bound of the sun's radiation reaching the earth's surface) and, therefore, can be expected to be subject to photolysis by sunlight (Koester and Hites, 1992b). The photochemistry of CDDs has been reviewed by Miller and Zepp (1987), Choudry and Webster (1987), and Esposito et al. (1980). This section summarizes the key findings of those reviews and the results of recent environmentally significant studies.

Laboratory studies have demonstrated that CDDs/CDFs undergo photolysis, typically following first order kinetics, in the presence of a suitable hydrogen donor such as oil or an organic solvent. Study results, when extrapolated to environmental conditions, indicate half-lives ranging from hours to days. The major products of photolysis are lower chlorinated CDDs/CDFs. In general, the rate of photolysis increases as the degree of chlorination decreases and, within a congener group, as the degree of ortho substitution decreases.

Most studies performed to date have been in a laboratory setting using laboratory lighting, pure compounds, and solvent solutions or clean solid surfaces as the reaction substrate. Because of the difficulties inherent in controlling experimental variables, few studies have been performed with gaseous-phase CDDs/CDFs or with surfaces or solutions that may more accurately simulate real world matrices. Thus, although photolysis of CDDs/CDFs at environmentally significant rates has been observed in laboratory studies, the results of these studies may not be representative of photolysis rates that occur under actual environmental conditions. The following paragraphs summarize some of the key studies to date regarding photolysis of CDDs/CDFs in the environment and the relevance of their findings.

Photodegradation in Water. Numerous studies have demonstrated that CDDs/CDFs will undergo photolysis following first order kinetics in solution. Photolysis is slow in water but increases dramatically when solvents serving as hydrogen donors are present such as hexane, benzene, methanol, acetonitrile, isooctane, and acetonitrile/water (Dobbs and Grant, 1979; Crosby et al., 1978; Dulin et al., 1986; Choudry and Webster, 1989; Friesen et al., 1990a; Hutzinger, 1973; Buser, 1988; Koester and Hites, 1992; and others). As noted above, the photolytic behavior of CDDs/CDFs in an organic solvent or a water-organic solvent may not accurately reflect the photolytic behavior of these compounds in natural waters. Natural waters have differing quantities and types of suspended particulates and dissolved organic material that could either retard or enhance the photolysis of CDDs/CDFs. For example, Choudry and Webster (1989) reported that photolysis of 1,3,6,8-TCDD was slower in a pond water matrix than was predicted from a laboratory solution. Conversely, Friesen et al. (1990a) and Friesen et al. (1993) reported that photolysis of PeCDD, HpCDD, TCDF, and PeCDF proceeds much faster in a pond or lake water matrix than was predicted from or measured in a laboratory solution.

Dobbs and Grant (1979) investigated the photolysis of a series of hexa-, hepta-, and octa-CDDs in hexane. Photolysis half-lives ranged from 0.4 days to 2 days. Meta- and para-substituted congeners were degraded more rapidly than ortho-substituted congeners. Dulin et al. (1986) studied the photolysis of 2,3,7,8-TCDD in various solutions under sunlight and artificial light. Using the results obtained in a water:acetonitrile solution (1:1, v/v) under sunlight conditions, Dulin et al. (1986) calculated the half-life of 2,3,7,8-TCDD in surface water in summer at 40 degrees north latitude to be 4.6 days. The quantum yield for photodegradation of 2,3,7,8-TCDD in water was three times greater under artificial light at 313 nm than under sunlight, and the artificial light photolysis quantum yield for hexane, a good hydrogen donor, was 20 times greater than for the water:acetonitrile solution, a poor hydrogen donor.

Podoll et al. (1986) used the Dulin et al. (1986) quantum yield data for the water:acetonitrile solution to calculate seasonal half-life values for dissolved 2,3,7,8-TCDD at 40 degrees north latitude in clear near-surface water. The seasonal values for half-lives were calculated to be 0.9 days in summer, 2.1 days in fall, 4.9 days in winter, and 1.1 days in spring.

Choudry and Webster (1989) studied the photolytic behavior under 313 nm light of a series of CDDs in a water:acetonitrile solution (2:3, v/v). Assuming that the quantum yields observed in these studies are the same as would be observed in natural waters, Choudry and Webster (1989) estimated the mid-summer half-life values at 40 degrees north latitude in clear near-surface water to be as follows: 1,2,3,7-TCDD (1.8 days); 1,3,6,8-TCDD (0.3 days); 1,2,3,4,7-PeCDD (15 days); 1,2,3,4,7,8-HxCDD (6.3 days); 1,2,3,4,6,7,8-HpCDD (47 days); and OCDD (18 days). In addition, the authors also experimentally determined the sunlight photolysis half-life of 1,3,6,8-TCDD in pond water to be 3.5 days (i.e., ten times greater than the half-life predicted by laboratory experiments).

A recent study by Friesen et al. (1990a) examined the photolytic behavior of 1,2,3,4,7-PeCDD and 1,2,3,4,6,7,8-HpCDD in water:acetonitrile (2:3, v/v) and in pond water under sunlight conditions at 50 degrees north latitude. The observed half-lives of these two compounds in the acetonitrile solution were 12 and 37 days, respectively, and 0.94 and 2.5 days in pond water, respectively.

Crosby et al. (1973) reported that polychlorinated dibenzofurans undergo photolytic dechlorination in the presence of a hydrogen donor, with more highly chlorinated congeners being more stable. In contrast, Hutzinger (1973) and Buser (1976) reported that the more highly chlorinated congeners undergo photodegradation at a rate similar to that of lower chlorinated CDFs. Hutzinger (1973) found that both 2,8-DCDF and OCDF photolyze rapidly in methanol and hexane.

Buser (1988) studied the photolytic decomposition rates of 2,3,7,8-TCDF, 1,2,3,4-TCDF, and 1,2,7,8-TCDF in dilute isooctane solutions under sunlight and artificial laboratory illumination (fluorescent lights). When the solutions were illuminated with sunlight, the estimated half-lives were 0.2 days for a solution containing 3 ng/µl of 2,3,7,8-TCDF, 0.1 days for a solution containing 2 ng/µl of 1,2,3,4-TCDF, and 0.4 days for a solution containing 0.3 ng/µl of 1,2,7,8-TCDF. For the same solutions illuminated with artificial light, the half-lives were greater than 28 days.

Friesen et al. (1993) studied the photodegradation of 2,3,7,8-TCDF and 2,3,4,7,8-PeCDF using water: acetonitrile (2:3, v/v) and lake water. The observed half-lives of the TCDF and PeCDF in the acetonitrile solution were 6.5 and 46 days, respectively, and 1.2 and 0.19 days in lake water, respectively.

Photodegradation in Soil. As discussed in Section 2.6.1.2.2 (Transport Mechanisms in Soil), photodegradation of CDDs/CDFs is limited only to the soil surface. Below the top few millimeters of soil, photodegradation is not a significant process (Puri et al., 1989; Yanders et al., 1989). Substantial research on the environmental persistence of 2,3,7,8-TCDD has been performed as part of the decontamination of the area around the ICMESA chemical plant in Seveso, Italy. That area was contaminated when a trichlorophenol reaction vessel overheated in 1976, blowing out the safety devices and spraying dioxin-contaminated material into the environment. The levels of dioxin in the soil decreased substantially during the first 6 months following the accident, reaching a steady state of 1/5 to 1/11 of the initial levels (DiDomenico et al., 1982). An experiment was conducted at this site to determine the effectiveness of photolysis in decontaminating surface deposits on foliage. Test plots were sprayed with olive oil to act as a hydrogen donor, and the levels of dioxin on grass were found to be reduced by over 80 percent within 9 days (Crosby, 1981). The 2,3,7,8-TCDD in contaminated soil was also found to be photolabile in sunlight when the soil was suspended in an aqueous solution of a surfactant. The destruction of 8 µg/ml of 2,3,7,8-TCDD in 0.02 M hexadecylpyridinium chloride could be accomplished in 4 hours (Botre et al., 1978).

Buser (1988) studied the photolytic decomposition rates of 2,3,7,8-TCDF, 1,2,3,4-TCDF, and 1,2,7,8-TCDF dried as thin films on quartz vials. When exposed to sunlight, the substances slowly degraded with reported half-lives of 5 days, 4 days, and 1.5 days, respectively.

Koester and Hites (1992b) studied the photodegradation of a series of tetra- through octa-chlorinated CDDs and CDFs on silica gel. In general, the CDFs degraded much more rapidly than the CDDs, and half-lives increased with increasing level of chlorination (1,2,7,8-TCDF excluded). The half-lives for CDDs ranged from 3.7 days for 1,2,3,4-TCDD to 11.2 days for OCDD. The half-lives for CDFs ranged from 0.1 day for 1,2,3,8,9-PeCDF to 0.4 days for OCDF.

Photodegradation in Air. Photolysis of CDDs/CDFs in the atmosphere has not been well-characterized. Based on the data generated to date, however, photolysis appears to be the most significant mechanism for degradation of those CDDs/CDFs present in the atmosphere in the gas phase. For airborne CDDs/CDFs sorbed to particulates, photolysis appears to proceed very slowly, if at all. Because of the low volatility of CDDs/CDFs, few studies have been attempted to measure actual rates of photodegradation of gaseous-phase CDD/CDF, and only recently have studies been undertaken to examine the importance of photolysis to particulate-bound CDDs/CDFs.

Podoll et al. (1986) estimated the photolysis rate of 2,3,7,8-TCDD vapors in the atmosphere. Based on the quantum yield for photolysis in hexane, the half-life in summer sunlight at 40 degrees north latitude was calculated to be 1 hour, but Podoll et al. (1986) stated this estimate is an upper limit.

Mill et al. (1987) reported preliminary photolysis experiments with vapor phase 2,3,7,8-TCDD. The half-life for vapor phase 2,3,7,8-TCDD in simulated sun was several hours. The photolysis of 2,3,7,8-TCDD sorbed onto small diameter fly ash particulates suspended in air was also measured. The results indicated that fly ash appears to confer photostability on 2,3,7,8-TCDD. There was little (8 percent) to no loss observed on the two fly ash samples after 40 hours of illumination.

Koester and Hites (1992b) recently studied the photodegradation of CDDs/CDFs naturally adsorbed to five fly ashes (one from a hospital incinerator, two from municipal incinerators, and two from coal-fired power plants). Although they found that CDDs/CDFs underwent photolysis in solution and when spiked onto silica gel, no significant degradation was observed in 11 photodegradation experiments performed for periods ranging from 2 to 6 days. Three additional experiments were performed to determine what factors may have been inhibiting photolysis. From the results of these additional experiments, Koester and Hites (1992b) concluded that: 1) the absence of photodegradation was not due to the absence of a hydrogen-donor organic substance; 2) other molecules or the ash, as determined by a photolysis experiment with an ash extract, inhibit photodegradation either by absorbing light and dissipating energy or by quenching the excited states of the CDDs/CDFs; and 3) the surface of the ash itself may hinder photolysis by shielding the CDDs/CDFs from light.

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.

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.

Based on their low vapor pressures, low water solubilities, and high K_oc 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).

2.6.2.3.1. 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. The predicted half-lives for the congener groups containing the 11 coplanar PCBs are as follows:

Congener Group Half-Life in Air (days)

TeCBs 11 to 20

PeCBs 12 to 31

HxCBs 32 to 62

HpCBs 94+

Sedlak 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.

Water System Half-Life in Water (days)

Fresh surface water 4 to 11

Marine surface water 1,000 to 10,000

Cloud water 0.1 to 10

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 may be 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, S_N1 and S_N2 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.

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.

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.

2.7.3.1. 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:

Half-Life Half-Life

Compound in Water (hrs) in Air (min)

1,2,3,4-TBDD 7 <1

2,3,7,8-TBDD 2 0.3

2,3,7,8-TBDF 1.7 0.2

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.

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.

PBDD Half-Life PBDF Half-Life

No. of Br in Air (hrs) in Air (hrs)

4 50 420

5 50 430

6 100 960

7 200 1900

8 770 3800

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

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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