Volume II Chapter 2.0 Pages 2 of 3 page next page 3

2.4.5. Organic Carbon Partition Coefficient 2-18

2.4.6. Photo Quantum Yields 2-19



2.6.1. Environmental Fate of Chlorinated Dibenzo-p-dioxins (CDDs) and Chlorinated Dibenzofurans (CDFs) 2-21 Summary 2-21 Transport Mechanisms 2-21 Transport Mechanisms in Air 2-21 Transport Mechanisms in Soil 2-26 Transport Mechanisms in Water. 2-28 Transformation Processes 2-30 Photodegradation 2-30

2.4.5. Organic Carbon Partition Coefficient

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

Measured log Koc 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 Koc of 2,3,7,8-TCDD. Log Koc 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 Koc.

The average log Koc 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 Koc value. The soil was stirred with water in 2-liter flasks. The log Koc 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 Koc of 2,3,7,8-TCDD with several other co-contaminants such as crankcase oils and surfactants. An average log Koc 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 Koc of 2,3,7,8-TCDD. The study resulted in a log Koc value of 6.6.

Four studies for log Koc 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 Koc values in the ranges reported for these chemicals are expected to strongly sorb to particulate matter.

2.4.6. Photo Quantum Yields

Diagram 2-5

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.


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

table Table 2-4. Vapor-to-Particle-Bound Ratio (V/P) for CDDs and CDFs in Ambient Air: Monitoring Results and Modeling Estimates .
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.
expand table table 2-4

... 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.

table Table 2-5. Rain Scavenging Ratios (W) and Percent Washout Due to Particulates (%P) for CDDs and CDFs in Bloomington and Idianapolis Ambient Air .
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

expand table table 2-5
... to be the dominant process for Indianapolis. The difference was attributed to the higher total suspended particulate matter in Indianapolis air. 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 Kocs 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). 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). Transformation Processes 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/l of 2,3,7,8-TCDF, 0.1 days for a solution containing 2 ng/l of 1,2,3,4-TCDF, and 0.4 days for a solution containing 0.3 ng/l 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 (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 g/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.

Orth et al. (1989) conducted photolysis experiments with vapor-phase 2,3,7,8-TCDD under illumination with a light source and filters to achieve radiation in the UV region from 250 nm to 340 nm. Carrier gases included air, helium, and an isobutane/helium mixture. The rate constants in helium and air were very similar, 5.4 x 10-3 sec-1 and 5.9 x 10-3 sec-1, which corresponds to a quantum yield in air of 0.033 + 0.046.

No products could be observed in the mass spectrometer, so Orth et al. (1989) postulated that the product might be sorbing to the surface of the photolysis cell and being lost from potential analysis. Further studies were suggested to study product sorption to surfaces and to determine any wave length dependence of the photoinduced loss across the absorption band studied.

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.