4. LEVELS OF CDD, CDF, AND PCB CONGENERS IN ENVIRONMENTAL

MEDIA AND FOOD 4-1

4.1. INTRODUCTION 4-1

4.2. CONCENTRATIONS IN SOIL 4-2

4.2.1. North American Data 4-2

4.2.2. European Data 4-4

4.2.3. Soil Summary 4-5

4.3. CONCENTRATIONS IN WATER 4-6

4.3.1. North American Data 4-6

4.3.2. European Data 4-7

4.3.3. Water Summary 4-7

4.4. CONCENTRATIONS IN SEDIMENT 4-8

4.4.1. North American Data 4-8

4.4.2. European Data 4-11

4.4.3. Sediment Summary 4-13

4.5. CONCENTRATIONS IN FISH AND SHELLFISH 4-13

4.5.1. North American Data 4-14

4.5.2. European Data 4-18

4.5.3. Fish Summary 4-20

4.6. CONCENTRATIONS IN FOOD PRODUCTS 4-22

4.6.1. Migration of CDD/CDF from Paper Packaging Into Food 4-23

4.6.2. U.S. Food 4-25

4.6.3 European Food 4-38

4.6.4 Canadian Food 4-41

4.7. CONCENTRATIONS IN AIR 4-41

4.7.1. U.S. Data 4-44

4.7.2. European Data 4-45

4.7.3. Air Summary 4-47

4.8. TEMPORAL TRENDS 4-48

4.9. SUMMARY OF CDD/CDF LEVELS IN ENVIRONMENTAL MEDIA AND FOOD 4-50

4.10. MECHANISMS FOR ENTRY OF CDD/CDFS INTO THE FOOD CHAIN 4-57

Polychlorinated dibenzo-p-dioxins (CDDs), polychlorinated dibenzofurans (CDFs), and polychlorinated biphenyls (PCBs) have been found throughout the world in practically all media including air, soil, water, sediment, fish and shellfish, and other food products such as meat and dairy products. Also, not unexpectedly, considering the recalcitrant nature of these compounds and their physical/chemical properties (i.e., low water solubilities, low vapor pressures, and high K_ows and K_ocs), the highest levels of these compounds are found in soils, sediments, and biota (ppt and higher); very low levels are found in water and air (ppq and lower). The widespread occurrence observed is not unexpected considering the numerous sources that emit these compounds into the atmosphere (discussed in Chapter 3), and the overall resistance of these compounds to biotic and abiotic transformation. (See Chapter 2.) Part-per-trillion levels of CDDs/CDFs have been found in everyday materials that are contaminated with dust--clothes dryer lint (2.4 to 6.0 ng TEQ/kg; vacuum cleaner dust (8.3 to 12 ng TEQ/kg); room air filters (27 to 29 ng TEQ/kg); and house furnace filter dust (170 ng TEQ/kg) (Berry et al., 1993). Although Berry et al. (1993) only analyzed one or two samples of these materials, the findings suggest that these compounds may be ubiquitous.

The purpose of this chapter is to provide an overview of the concentrations at which these compounds have been found in the environment and food based on data presented in the published literature. This literature summary is not all inclusive, but is meant to present the reader with a general overview of values reported in the recent literature. Only data from Government-sponsored monitoring studies and studies reported in the peer-reviewed literature are discussed in this chapter. Data are presented as they were presented in the original studies/reports. No attempt has been made to verify or assess the adequacy of the quality assurance/quality control measures employed in these studies beyond those described in the published reports. The final section of this chapter discusses the mechanisms by which these compounds could enter the food chain.

Tables B-1 and B-2 (Appendix B) contain summaries of data from several of the numerous studies in the published peer-reviewed literature regarding concentrations of CDDs and CDFs in soil. Data on coplanar PCB congener soil concentrations were not found in the literature; the PCB soil concentration data found in the literature were reported as either total PCB concentrations or concentrations of Aroclor PCB mixtures. Descriptions of several of the studies summarized in Appendix B are presented below.

Soil sampled in 1987 from the vicinity of a sewage sludge incinerator was compared with soil from rural and urban sites in Ontario, Canada, by Pearson et al. (1990). Soil in the vicinity of the incinerator showed a general increase in CDD concentration with increasing degree of chlorination. Of the CDFs, only OCDF was detected (mean concentration 43 ppt). Rural woodlot soil samples contained only OCDD (mean concentration of 30 ppt). Soil samples from undisturbed urban parkland settings revealed only HpCDDs and OCDD, but all CDF congener groups (Cl₄ to Cl₈) were present. Those samples showed an increase in concentration from the HpCDDs to OCDD and PeCDFs to OCDD. The TCDFs had the highest mean value (29 ppt) of all the CDF congener groups. Resampling of one urban site in 1988, however, showed high variability in the concentrations of CDDs and CDFs.

Data were collected on CDD and CDF levels in soil samples from industrial, urban, and rural sites in Ontario and some U.S. Midwestern States (Birmingham, 1990). The levels of CDD/CDF in rural soils were primarily nondetected (ND), although the HpCDDs and OCDD were found in a few samples. In urban soils, the tetra- through octa-congener groups were measured for both CDDs and CDFs. The HpCDDs and OCDD dominated the homologue profile and were two orders of magnitude greater than in the rural soils. These soils also contained measurable quantities of the TCDDs and PeCDDs. Industrial soils did not contain any TCDDs or PeCDDs, but they contained the highest levels of the TCDFs, HpCDFs, and OCDF. In another study, soils from industrialized areas of a group of cities from Midwestern and Mid-Atlantic States (Michigan, Illinois, Ohio, Tennessee, Pennsylvania, New York, West Virginia, Virginia) were analyzed for levels of 2,3,7,8-TCDD (Nestrick et al., 1986). Many of the samples were taken within 1 mile of major steel, automotive or chemical manufacturing facilities, or municipal solid waste incinerators. Concentrations of 2,3,7,8-TCDD measured in this study ranged from ND to 9.4 ppt.

EPA conducted a 2-year nationwide study to investigate the national extent of 2,3,7,8-TCDD contamination (U.S. EPA, 1987). The results of this large study were summarized broadly in the primary reference (i.e., the number and types of samples per site and range of detection). The method used to analyze samples for five of the seven "tiers" of the study had a detection limit in soil, sediment, and water of 1 part per billion (ppb). [Each "tier" of sites is a grouping of sites with a common past or present use (e.g., industrialized, pristine, etc.)]. Only Tier 5 (sites where pesticides derived from 2,4,5-trichlorophenol (TCP) have been or are being used for commercial purposes), and Tier 7 (ambient sampling for fish and soil) had detection limits of 1 ppt. Consequently, the data from this study are not included in the tables, but some observations from this study with regard to soil contamination are discussed below.

Soil concentrations found in most of the 100 Tier 1 and 2 sites (i.e., sites already on or expected to be on the NPL list) were in the ppb range; although in a few sites where concentrated 2,4,5-TCP production wastes were stored or disposed, concentrations were as high as 2,000 parts per million (ppm). Offsite soil contamination of concern was confirmed in 7 of the 100 Tier 1 and 2 sites, with soil concentrations in the ppb range. Eleven of 64 Tier 3 sites (facilities and associated disposal sites where 2,4,5-TCP and its derivatives were formulated into pesticide products) were found to have soil concentrations exceeding 1 ppb, and in 7 of 11 sites where contamination was found, only one or two soil samples were above 1 ppb. Fifteen of 26 Tier 5 sites (areas where 2,4,5-TCP and pesticide derivatives had been or were being used) had concentrations above 1 ppt, and one of those had a single detection of 6 ppb. Two-thirds of all detections at the Tier 5 sites were below 5 ppt. Three of 18 Tier 6 sites (organic chemical and pesticide manufacturing facilities where improper quality control on production processes could have resulted in 2,3,7,8-TCDD being introduced into the wastestreams) had soil concentrations that exceeded the detection limit of 1 ppb, although these levels were limited to one or two samples per site. Seventeen of the 221 urban soil sites and 1 of the 138 rural sites from Tier 7 (background sites not expected to have contamination) had soil concentrations exceeding 1 ppt. The highest concentration detected (11.2 ppt) was found in an urban sample. The results from Tier 7 are consistent with the other studies discussed above regarding soil concentrations of 2,3,7,8-TCDD in nonindustrial settings.

Soil samples from rural and semi-urban sites in England, Wales, and lowland Scotland showed a general increase in concentration from the TCDDs to OCDD, whereas the CDF levels showed very little variation between the congener groups (Creaser et al., 1989). Concentrations of 2,3,7,8-TCDD at those sites ranged from <0.5 to 2.1 parts per trillion (ppt). The median values for the TCDDs to OCDD were 6.0, 4.6, 31, 55, and 143 ppt, respectively. The median values for the TCDFs to OCDF were 16, 17, 32, 15, and 15 ppt. Evaluation of soil data from urban sites in the same geographical area showed that the mean levels for the CDD and CDF congeners were significantly greater (p<0.01) than those for rural and semi-urban background soils (Creaser et al., 1990). Concentrations of 2,3,7,8-TCDD at the urban sites ranged from <0.5 to 4.2 ppt. The median values for the TCDDs to OCDD were 40, 63, 141, 256, and 469 ppt, respectively. The median values for the TCDFs to OCDF were 140, 103, 103, 81, and 40 ppt. The significantly elevated levels of the lower congeners, together with higher overall CDD/CDF concentrations, are indicative that local sources and short-range transport mechanisms are major contributors of CDDs and CDFs to urban soils.

Analysis of four sites in Hamburg, Germany, contaminated by an organochlorine pesticide manufacturing company showed patterns of CDD and CDF distribution that are similar to the urban and industrial sites examined in England, Wales, and Scotland (Sievers and Friesel, 1989). The study indicated that CDDs and CDFs showed a regular increase in concentration with increasing degree of chlorination (although individual data points were not presented). The maximum concentrations of 2,3,7,8-TCDD ranged from 900 ppt to 874,000 ppt. The very high concentrations of 2,3,7,8-TCDD at the sites were attributed to an admixture of wastes from 2,4,5-T production.

A soil sampling survey in Salzburg, Austria, also showed that the concentrations of CDD/CDFs were higher in urban and industrial sites than in rural sites (Boos et al., 1992). The total CDD content of the soils ranged from 33.7 to 1236.7 ppt for urban sites; 92.2 to 455 ppt for industrial sites; and 7.1 to 183.6 ppt for rural sites. The total CDF content of the soils ranged from 45.6 to 260.8 ppt for urban sites; 53.0 to 355.3 ppt for industrial sites; and 12.0 to 77.7 ppt for rural sites.

Rotard et al. (1993) measured CDD/Fs in soil samples collected from forest, grassland. and plowland sites in western Germany. The highest mean concentration of CDD/Fs were found in the sursoil and topsoil layers of deciduous (38.0 ng TEQ/kg dry matter) and coniferous forests (36.9 ng TEQ/kg dry matter). Grassland and plowland sites had mean concentrations of 2.3 ng TEQ/kg dry matter and 1.7 ng TEQ/kg dry matter, respectively.

Some general observations for CDD and CDF levels in soils are possible from the data presented in the various soil studies discussed above:

• Generalizations about the prevalence of specific congeners within a congener group are not possible.

• As the degree of chlorination increases, the concentrations increase. Concentrations of the hepta- and octa-chlorinated congeners are generally higher than the tetra-, penta-, and hexa-chlorinated congeners.

• Concentrations associated with industrial sites clearly are the highest, with concentrations in the hundreds to thousands of parts per trillion.

• Concentrations in settings identified as urban are higher than those in areas identified as rural.

Based on the above studies, 95 samples were selected as representing background conditions in the United States. The mean TEQ level was estimated to be 8 ppt assuming that nondetects equal half the detection limit. Similarly, 133 background samples were selected from the European studies and were estimated to have a mean of 9 ppt of TEQ.

Tables B-3 and B-4 (Appendix B) contain summaries of data from the limited number of published studies regarding concentrations of CDDs and CDFs in water. Data on coplanar PCB congener water concentrations were not found in the literature. Several of these studies are discussed below.

A survey of 49 drinking water supplies in Ontario, including supplies in the vicinity of chemical industries and pulp and paper mills, was initiated in 1983 (Jobb et al., 1990). As of February 1989, 4,347 congener analyses had been performed on 399 raw and treated water samples. OCDD was detected in 36 of 37 positive results and ranged from 9 to 175 ppq in raw samples (33 positive samples) and 19 to 46 ppq in treated samples (4 positive samples). These low concentrations were found primarily in water obtained downstream of industrialized areas in the St. Clair/Detroit River system. Concentrations of 2,3,7,8-TCDD were not detected in any sample. Because CDDs and CDFs are hydrophobic concentrations of compounds and consequently have a tendency to sorb onto particulate matter in water, conventional water treatment processes are expected to be effective in removing the contaminants along with the particulates. This is substantiated by the fact that 33 of the 37 positive results were raw water samples. Because of the relatively low levels of CDDs detected in the samples, it is difficult to ascertain whether the CDDs were particulate-associated or dissolved.

A survey of 20 community water systems throughout New York State was conducted in 1986 (Meyer et al., 1989). The sampling sites were representative of the major surface source waters in New York. They included sources receiving industrial discharges or known to contain dioxin-contaminated fish, as well as waters in more remote areas. TCDFs were detected in the finished water at the Lockport facility (duplicate samples had concentrations of 2.1 and 2.6 ppq). Except for a trace of OCDF detected at one location, no other CDDs/CDFs were detected in finished water at any of the other 19 community water systems surveyed. Raw water sampled at the Lockport facility contained concentrations of TCDDs (1.7 ppq) as well as TCDFs to OCDF (18, 27, 85, 210, and 230 ppq, respectively). As can be seen from the data, the CDF congener group concentrations increased with increasing chlorine number.

CDDs in surface water samples collected from the Eman River in southern Sweden generally increased in concentration from the TCDDs to OCDD; whereas the CDF levels showed very little variation between the congener groups (Rappe et al., 1989b). In general, however, the levels of CDFs were higher than the levels of CDDs. Concentrations of 2,3,7,8-TCDF were 0.022 parts per quadrillion (ppq) in Jarnsjon and 0.026 ppq in Fliseryd. The filtered water, before chlorination and distribution as drinking water, had no detectable tetra-, penta-, or hexa-chlorinated congeners of CDDs or CDFs, but the HpCDDs and OCDD were detected at 120 and 170 ppq, respectively.

Some general observations for CDD and CDF levels are possible from the limited data available from the various water studies above:

• CDDs/CDFs are seldom detected in drinking water at ppq levels or higher.

• Raw water samples generally have higher concentrations of CDDs/CDFs than finished water samples.

• The concentration of CDDs and CDFs in surface water generally increases from the tetra-chlorinated to the octa-chlorinated congener groups.

Based on the above studies, a total of 214 samples were selected as representing background conditions in North America. The mean TEQ level was computed as 0.0056 ppg, assuming that nondetects equal half the detection limit. It should be noted, however, that OCDD and OCDF were the only congeners for which background data were available. Of the 214 samples analyzed for OCDD, 4 were positive, and 2 out of 22 samples analyzed for OCDF were positive. No appropriate data could be found for Europe.

Tables B-5 through B-7 (Appendix B) contain summaries of data from several of the numerous studies in the published literature regarding concentrations of CDDs, CDFs, and coplanar PCB congeners in sediment. Several of these studies are discussed in the following paragraphs.

In sediment samples collected from estuaries adjacent to an industrial site in Newark, New Jersey, where chlorinated phenols had been produced, the level of OCDD was many times higher than that of 2,3,7,8-TCDD (Bopp et al., 1991). The study indicated that there probably is a significant regional source (i.e., combustion and/or use of a common wood preservative, pentachlorophenol) for OCDD depleted in 2,3,7,8-TCDD relative to the local industrial source. A high correlation was found between 2,3,7,8-TCDD and 2,3,7,8-TCDF (R²=0.87), which suggests that the industrial site was a major source of 2,3,7,8-TCDF to the natural waters of the study area. An interesting note is that the bottom section (108-111 cm) of one sediment core contained 2,3,7,8-TCDD at a concentration of 21,000 ppt, the highest concentration measured in the study. This value was consistent with deposition of that sample during the mid to later stages of active 2,4,5-T production at the site from the late 1950s to early 1960s. CDD and CDF in Hudson River sediment samples contained primarily the higher chlorinated (Cl₆ to Cl₈) CDD and CDF congeners (Petty et al., 1982). Concentrations of the HpCDDs and OCDD homologues ranged from 5 to 15 ppb, and the OCDD homologue in most instances accounted for more than half of the total CDD residue. Likewise, the HpCDFs and OCDF occurred at the highest levels (ca. 1 ppb).

Surface sediment samples were collected from several estuaries in the United States (Norwood et al., 1989). The sampling sites included Black Rock Harbor in Bridgeport, Connecticut, (an industrialized urban estuary); central Long Island Sound (a relatively clean reference site); Narragansett Bay, Rhode Island, (where chemical industries may have contributed to the input); New Bedford Harbor, Massachusetts, (a section of which is a National Superfund Site because of PCB contamination); and Eagle Harbor, Washington, (the site of a creosote wood treatment facility). The sediments in New Bedford Harbor were reported to be more heavily contaminated with CDFs, especially with regard to the HxCDF congeners that were greater by a factor of 40 (although individual data points were not presented). In contrast, sediments from Eagle Harbor were practically devoid of CDFs and showed a large increase in the HpCDD and OCDD congeners closer to the treatment facility. Narragansett Bay and Black Rock Harbor were similar in both concentration and distribution of CDDs and CDFs, and Black Rock Harbor contained slightly higher levels of the tetra- to hepta-CDD and CDF congeners. Sediment from Long Island Sound was cleaner and had a distribution of CDFs between that of Narragansett Bay and Black Rock Harbor. Sediment with the least contamination was collected in New Bedford Harbor, up-river from the PCB facilities; the highest OCDD concentration (1400 ppt) was detected in Eagle Harbor.

Sediment samples from Siskiwit Lake, on Isle Royale, Lake Superior, were examined to evaluate the atmospheric input of CDDs and CDFs to the lake (Czuczwa et al., 1984). The water level in Siskiwit Lake is 17 meters higher than that in Lake Superior, and in addition, there are no anthropogenic inputs in the drainage basin of Siskiwit Lake. Consequently, the atmosphere is the only source of anthropogenic chemicals in that lake. OCDD was most predominant, and the HpCDD and HpCDF congeners also were abundant. The study indicated that the considerable decrease in concentration of all CDD and CDF between 6 and 8 cm of the sediment core depth (i.e., sediment believed to have been deposited about 1940).

Surficial sediments collected from Jackfish Bay on the north shore of Lake Superior contained moderate concentrations of the TCDF and OCDD congeners, with trace concentrations of other congeners (Sherman et al., 1990). The concentration of OCDD was similar to that found in the Siskiwit Lake sediment samples. The OCDF and OCDD profile for a sediment core collected from Moberly Bay was similar to the surficial sediment pattern. These congener groups predominated at all depths where detectable concentrations occurred. In addition, low concentrations of the HpCDD and PeCDF and HpCDF congeners were detected. The concentration profile of the HpCDF congener group showed a relatively high value that dropped abruptly to nondetectable (<60 ppt) below a depth of 10 cm. This abrupt change corresponded to a section date 1973 that reflects an operational change at the pulp mill.

A survey of surficial harbor sediments collected near a wood preserving plant in Thunder Bay, Ontario, Canada, on the north shore of Lake Superior, found CDDs and CDFs; the highest concentrations of which occurred at stations closest to the plant dock, and lower concentrations at locations further from the source (McKee et al., 1990). No CDDs or CDFs were detected below the surficial layer. TCDD and PeCDD congeners were below analytical detection limits in all samples. However, the concentrations of the HxCDDs to OCDD congeners increased with the degree of chlorination. The maximum concentrations of the HxCDDs to OCDD ranged from 5,700 ppt for the HxCDDs to 980,000 ppt for the OCDD. As with the CDD distribution profile, the HxCDFs to OCDF increased with the degree of chlorination.

Bottom surficial sediments (0-3 cm) were collected from the sedimentation basins of Lake Ontario to assess the levels of the various PCB congeners (Oliver and Niimi, 1988). Concentrations of 2,3',4,4',5-PeCB; 2,3,3',4,4'-PeCB; and 2,3,3',4,4',5-HxCB in the sediment were 15, 10, and 2.1 ppb, respectively. A baseline assessment of CDDs and CDFs was performed on the Elk River, a semi-rural area located about 25 miles northwest of Minneapolis-St. Paul, Minnesota (Reed et al., 1990). Sediment samples were collected from Lake Orono, a reservoir on the Elk River, and from an abandoned gravel pit. Although none of the sediment samples contained 2,3,7,8-TCDD, the gravel pit sediments contained measurable concentrations of TCDFs. Only one Lake Orono sample contained measurable concentrations of 2,3,7,8-TCDF (0.31 ppt) and total TCDF (0.54 ppt). The gravel pit samples also contained HpCDDs to OCDD and PeCDFs to OCDF. Lake Orono samples contained HpCDDs, OCDD, and HpCDF congeners. The HpCDDs ranged from 7.3 ppt in the lake inlet to 110 ppt in the gravel pit and the lake, near the dam. OCDD concentration ranged from 450 ppt in the gravel pit to 600 ppt in the middle of Lake Orono. The sediment profiles reflected combustion source influences.

The Sheboygan River, a Wisconsin tributary to Lake Michigan, is polluted with PCBs from the mouth to about 14 miles upstream (Sonzogni et al., 1991). That portion of the river is a Superfund site as well as one of the Great Lakes "Areas of Concern." Sediment cores were collected at Rochester Park, near the original source of the PCBs, about 14 miles upstream from the mouth. The PCB congeners 2,3',4,4',5-PeCB; 2,3,3',4,4'-PeCB; and 3,3',4,4'-TCB were detected in all samples and ranged from about 5 to 1500 ppb. The remaining coplanar PCB congeners were detected less frequently and ranged from nondetectable to slightly over 100 ppb. The PCB congener 2,3',4,4',5-PeCB appears to be the most common coplanar PCB in environmental samples and was found in the Sheboygan River sediments in the highest weight percent. The eight toxic PCBs detected in this study were present in relatively low concentrations compared to total PCBs or other more abundant congeners.

Sediments collected from Waukegan Harbor in Lake Michigan contained the coplanar PCB congeners 3,3',4,4'-TCB and 2,3,3',4,4'-PeCB (Huckins et al., 1988). The percentage of 3,3',4,4'-TCB in the samples averaged (0.16 percent ± 0.15) varied by 1.4 orders of magnitude, with concentrations ranging from 13 to 27,500 ppb. The percentage of 2,3,3',4,4'-PeCB averaged 0.66 percent ± 0.37, with concentrations ranging from 102 to 131,000 ppb. In another Lake Michigan study, sediment samples collected from Green Bay contained concentrations of all 11 coplanar PCB congeners (Smith et al., 1990). The dominant congeners were 2,3,4,4',5-PeCB and 2,3,3',4,4'-PeCB with concentrations of 11 and 5.8 ppb, respectively.

Sediment samples from the vicinity of a magnesium production plant in Norway were analyzed for CDDs and CDFs (Oehme et al., 1989). The concentration distribution of CDD and CDF congeners was rather homogeneous except for a slight decrease at a sampling station further downstream of the plant. However, the deeper sediments (4-6 and 11-13 cm depth) at that site had somewhat higher levels. Another sampling station even further downstream had concentrations that were a factor of 4 to 10 lower, thereby indicating substantial transport of CDDs and CDFs. The TCDF congener profiles were the same as those for magnesium production. In addition, the PeCDF congener profiles were very similar to those found in the wastewater.

Trapped sediments from the archipelago of Stockholm, Sweden, displayed CDD and CDF congener distribution patterns that were very similar to those exhibited in total air and air particulates (Rappe and Kjeller, 1987). The HpCDDs, OCDD, and HpCDF were the dominant congener groups in the sediment.

Bottom surface sediment samples collected from the Baltic Sea showed interesting CDD and CDF distribution patterns (Rappe et al.,1989a). The background samples, one between the Swedish and Soviet coasts and the other between the Swedish and Finnish coasts, contained similar levels and distribution profiles. The study indicated that the pattern of the TCDF congeners at these sites was typical of the "incineration pattern" (i.e., patterns resulting from MSW incineration, car exhausts, steel mills, etc.) which also had been found in samples of air and air particulates. However, sediment samples collected at a distance of 4 to 30 km from a pulp mill revealed a congener distribution pattern typical of bleaching mills. The TCDFs found in the sediment 4 km from the pulp mill contained only two major congeners. The sediment collected 30 km from the mill displayed the same pattern.

Surface sediments collected from 18 lake areas in central Finland were analyzed for CDDs, CDFs, and PCBs. Although 2,3,7,8-TCDD was not detected in any samples, two other TCDD congeners that are common and abundant in pulp mill effluents were detected. In addition, the HxCDD to OCDD congener groups as well as the HpCDF and OCDF congener groups not normally linked to pulp mills, also were detected. The study suggested that these may have resulted from combustion operations in the more densely populated and industrialized areas in south Finland. Coplanar PCBs were detected at low background levels. The most toxic PCB congener, 3,3',4,4',5-PeCB (IUPAC No. 126), was found only in one area located nearest to a PCB leakage point. The concentration of PCB 126 at that site was 110 ppt.

Evaluation of sediments in Hamburg Harbor in Germany revealed high concentrations of the TCDDs through OCDD (mean concentrations of 564, 1112, 2744, 4040, and 7560 ppt, respectively) and the TCDFs through OCDF (mean concentrations of 526, 2980, 4106, 2358, and 2712 ppt) (Gotz et al., 1990). The average concentration of 2,3,7,8-TCDD was 375.3 ppt. The high concentrations of 2,3,7,8-TCDD, especially in the Moorfleeter Canal and the Auserer Vering Canal, were attributed to discharges from an organochlorine pesticide manufacturing plant. The patterns of 2,3,7,8-TCDD and the other HpCDD congeners are characteristic of the patterns resulting from the production of 2,4,5-T and 2,4,6-trichlorophenol. In addition, the pattern of the HpCDF congeners can be linked to emissions from thermal processes employed by chemical industries in the production of chlorinated organic chemicals. The high concentrations of hepta- and octa-CDDs/CDFs may also be the result of other industrial combustion processes in the Hamburg area.

Some general observations for CDD and CDF levels are possible from the data presented in the various sediment studies above:

• The CDD and CDF congener distribution patterns in sediment generally follow those exhibited by the contaminant source.

• The concentration of hexa- to octa-chlorinated CDD and CDF congeners in sediment is usually the result of industrial processes and generally increases with the degree of chlorination, but decreases uniformly with distance from the source.

Based on the above studies, seven samples were selected as representing background conditions in the United States. The mean TEQ level was computed as 3.9 ppt assuming that nondetects equal half the detection limit. Similarly, 20 background samples were selected from the European data with a mean of 34.9.

Tables B-8 through B-10 (Appendix B) contain summaries of data from the numerous studies in the published literature regarding concentrations of CDDs, CDFs, and coplanar PCB congeners in fish and shellfish. PCB congener data were found only for North American species. It should be noted that some studies reported fish concentrations on a whole weight basis and others reported concentrations for fish fillets. Whole weight concentrations were converted to fillet concentrations assuming that the fillet contained one-half the concentration of the whole fish (USEPA 1990; Branson et al. 1985). This was necessary for estimating human exposures because it is assumed that fish fillets, and not whole fish, are ingested by humans.

A large quantity of fish data were collected as part of EPA's National Study of Chemical Residues of Fish (NSCRF), more commonly referred to as the National Bioaccumulation Study, during the period of 1986 to 1989 (U.S. EPA, 1992). Based on these data, several summaries were prepared and are presented here. Tables B-8 and B-9 include the dioxin and furan data collected as part of the National Bioaccumulation Study. Samples were collected from a wide variety of sites across the United States, including 314 sites thought to be influenced by point or nonpoint sources and 35 sites identified as relatively free of influence from point and nonpoint sources. This latter group of sites can be characterized as background per the definition used in this document. Background data are presented in Table 4-1. Table B-10 includes similar data for the various PCB congener groups from 362 National Bioaccumulation Study sites. Because the specific PCB congeners could not be identified, it is not known what percentage of these concentrations represent the PCBs identified as dioxin-like. Twenty of these sites were identified as background sites. The total PCB, all 209 congeners, mean concentration for these background sites was 46,900 ppt. Because the dioxin-like PCBs consist of only 11 of the 209 possible PCB congeners, it may be that they are a small percentage of the total. However, only congener specific analysis can ultimately confirm this. As discussed at the end of this section (in 4.5.3), this study was selected as the best basis for estimating background fish levels in the United States.

Samples of striped bass, blue crabs, and lobsters collected from Newark Bay and the New York Bight all contained high levels (up to 6,200 ppt) of 2,3,7,8-chlorine substituted tetra- and penta-CDDs and CDFs (Rappe et al., 1991). The levels of 2,3,7,8-TCDD were higher than any other New Jersey samples, and the highest sample in this study may be the highest level of 2,3,7,8-TCDD ever reported for aquatic animals. The crustaceans resembled one another in congener pattern. Specifically, they all contained both a large number and large amounts of CDD and CDF congeners in addition to the 2,3,7,8-chlorine substituted compounds. The striped bass samples, on the other hand, contained primarily the 2,3,7,8-chlorine substituted congeners.

Carp, catfish, striped bass, large mouth bass, and lake trout were collected from sites in the Hudson River and the Great Lakes Basin that were contaminated with industrial chemicals or contained known or suspected levels of PCBs (Gardner and White, 1990). The congener 2,3,7,8-TCDF was detected in 12 fish at levels that ranged from 3 to 93 ppt. A 2,3,7,8-chlorine substituted PeCDF was detected in 14 fish at levels ranging from 4 to 113 ppt. An interesting observation in this study was that 2,4,6-chlorine substituted CDFs were detected in four fish samples, suggesting that those fish may have been exposed to chlorinated phenols. The study indicated that the 2,4,6-

chlorine substituted CDFs occurred in the fish at levels similar to those of the 2,3,7,8-chlorine substituted CDFs, but with less frequency.

Samples of lake trout or walleye collected from each of the Great Lakes and Lake St. Clair were analyzed for CDDs and CDFs (De Vault et al., 1989). CDF and CDD concentrations in lake trout were substantially different for each lake and between sites in Lake Michigan, probably reflecting differences in types and amounts of loadings to the lakes. In all of the sampling sites except Lake Ontario, 2,3,7,8-TCDF was the dominant CDF congener in lake trout and ranged from 14.8 ppt in Lake Superior to 42.3 ppt in Lake Michigan. In Lake Ontario, the dominant congener in lake trout was a 2,3,7,8-chlorine substituted PeCDF. The distribution of CDF congeners in the Lake Erie walleye was very similar to that of the lake trout from Lake Superior. With regard to CDDs, the concentrations of 2,3,7,8-TCDD ranged from 1 ppt in Lake Superior to 48.9 ppt in Lake Ontario. With the exception of Lake Ontario, the dominant CDD congener was a 2,3,7,8-chlorine substituted PeCDD. A 2,3,7,8-chlorine substituted HxCDD also contributed significantly to the total CDD concentrations. As with CDFs, the distribution of CDD congeners in the Lake Erie walleye was very similar to that of the lake trout from Lake Superior.

In another study, CDDs and CDFs were measured in four species of salmonids (lake trout, coho salmon, rainbow trout, and brown trout) collected from Lake Ontario (Niimi and Oliver, 1989a). Levels of 2,3,7,8-TCDD in whole fish ranged from 6 to 20 ppt, and the HxCDD congener group was most dominant in all fish. High levels of OCDD also were detected in lake trout and coho salmon, but not in rainbow trout or brown trout. Although the total CDF levels were about 25 percent lower than the total CDD concentrations, the levels of 2,3,7,8-TCDF (which was the dominant component of the TCDF congener group) were the same range as 2,3,7,8-TCDD (6 to 20 ppt). However, the study suggested that, although collection sites can influence chemical levels and congener composition, comparisons of chemical levels and congener frequencies may not be suitable because of differences resulting from localized factors. The study also indicated that the importance of the various CDD and CDF congeners can differ with location (i.e., the same species of fish collected at different locations in a study area may reveal that the most common congener is different at each site).

Travis and Hattemer-Frey (1991) evaluated data generated as part of the National Dioxin Study regarding 2,3,7,8-TCDD concentrations in fish. The fish were collected from 304 urban sites in the vicinity of population centers or areas with known commercial fishing activity, including sites from the Great Lakes Region. Data from that study indicated that concentrations of 2,3,7,8-TCDD in whole fish from urban sites ranged from nondetectable to 85 ppt. In addition, only 29 percent of the fillets from urban sites had detectable levels of 2,3,7,8-TCDD, with a geometric mean concentration of 0.3 ppt. Whole fish samples from the Great Lakes Region had higher 2,3,7,8-TCDD levels than fish from urban areas (e.g., 80 percent vs 35 percent detectable levels). In the Great Lakes Region, 2,3,7,8-TCDD concentrations in whole fish samples ranged from nondetectable to 24 ppt, with a geometric mean of 3.8 ppt. These levels were 10 times higher than the concentration in whole fish from urban areas. Likewise, the mean concentration of 2,3,7,8-TCDD in Great Lakes Region fish fillets (2.3 ppt) was about seven times higher than the levels in the fillets from urban areas (0.3 ppt). As with the whole fish samples, fish fillet samples from the Great Lakes Region had higher 2,3,7,8-TCDD levels than fillets from background urban areas (e.g., 67 percent vs 29 percent detectable levels). Comparable levels of 2,3,7,8-TCDD were detected in whole bottom feeders and predators from the Great Lakes Region.

Samples from all trophic levels in the Lake Ontario ecosystem were analyzed for PCB congeners (Oliver and Niimi, 1988). Analysis revealed that the PCB concentration increased from water to lower organisms to small fish to salmonids, demonstrating the classical biomagnification process. In addition, the chlorine content of the PCBs increased at the higher trophic levels. PCBs with the highest chlorine content (57 percent) were found in sculpin, small bottom-living fish that feed on benthic invertebrates. The TrCBs and TCBs comprised a much higher percentage of the PCBs in the lower trophic levels than in salmonids and small fish. The percentage of PeCBs and OCPB in all samples was fairly uniform, but the HxCBs and HpCBs comprised a much larger fraction of the PCBs in the small fish and salmonids than in the lower trophic levels.

A study regarding the distribution of PCBs in Lake Ontario salmonids (brown trout, lake trout, rainbow trout, and coho salmon) showed that the PeCBs and HxCBs were dominant in all species (Niimi and Oliver, 1989b). The 10 most common PCB congeners represented about 52 percent of the total content and did not appear to be influenced by species or total concentration. The homologues observed averaged approximately 56 percent chlorine by weight in whole fish and muscle. The analysis of the chlorine content suggested that the more persistent congeners tend to behave as a homogeneous mixture instead of as individual congeners.

Evaluation of fish in the Baltic Sea (Gulf of Bothnia) and northern Atlantic Ocean in the vicinity of Sweden revealed that concentrations of CDDs and CDFs in herring from the Atlantic Ocean were lower than those in the Gulf of Bothnia (Rappe et al., 1989b). Detectable levels of 2,3,7,8-TCDD in salmon were found in both wild homing (4.6 to 19 ppt) and hatchery-reared (0.2 to 0.3 ppt) varieties in the Gulf of Bothnia. In addition, concentrations of the same representative congeners of the Cl₅ to Cl₈ CDD and CDF congener groups found in herring were found in both varieties of salmon. Levels of those congeners in the wild salmon, however, were five to ten times higher than the herring levels, while the levels in the hatched salmon essentially were the same as in the herring samples. Perch collected at a distance of 1-6 km from a pulp mill in the southern part of the Gulf of Bothnia contained 2,3,7,8-TCDD and 2,3,7,8-TCDF; the levels were higher in the samples collected closer to the pulp mill. These two compounds have been identified in bleaching effluents from pulp mills as well as in bleached pulp. Arctic char collected from Lake Vattern, a popular fishing lake in southern Sweden, contained levels of 2,3,7,8-TCDD (6.5 to 25 ppt), 2,3,7,8-TCDF (21 to 75 ppt), and representative congeners of the PeCDD and PeCDF homologues. There was a good correlation between the weight of the fish and the levels of CDDs and CDFs. The main general pollution sources of the long, deep, narrow lake are two pulp mills.

Fish (cod, haddock, pole flounder, plaice, flounder, and eel), mussels, and edible shrimps from a fjord area contaminated by wastewater from a magnesium factory in Norway were analyzed for CDDs and CDFs (Oehme et al., 1989). Certain magnesium production processes can result in the formation of substantial amounts of CDDs and CDFs as byproducts. The congener pattern of Cl₄ and Cl₅ CDDs and CDFs released in wastewater during the magnesium production process is very characteristic and is dominated by congeners with chlorine in the positions 1,2,3,7 and/or 8. Fish and shellfish differ in their ability to bioconcentrate CDD and CDF congeners. For example, fish generally only concentrate the most toxic 2,3,7,8-substituted congeners, whereas shellfish can usually concentrate most of the congeners. Nearly all congeners were present in the shrimp and mussel samples. Although these organisms displayed the very characteristic PeCDF congener pattern of the magnesium production process, some deviations were found in the TCDF congener distribution within those species. For fish, the concentrations of CDDs and CDFs are dependent on the exposure level, fat content, living habit, and the species degree of movement. The highest CDD and CDF levels were found in comparatively high fat-content bottom fish collected close to the source. Cod and haddock, lower fat-content nonstationary fish, had much lower concentrations, even in the vicinity of the magnesium production factory. An interesting note is that the main stream of the fjord follows the west coast; subsequently, cod and eel samples collected along the west coast of the fjord had considerably higher levels of CDDs and CDFs than those collected from the eastern fjord entrance. Similarly, the level of 2,3,7,8-TCDD in mussels decreased by one order of magnitude from the vicinity of the magnesium production factory to the outer region of the fjord system.

Brown trout, grayling, barbel, carp, and chub collected in the Neckar River in southwest Germany contained much higher levels of 2,3,7,8-TCDF than in eels collected from the same river and the Rhine River (Frommberger, 1991). In addition, eels from both rivers showed very similar patterns for CDD and CDF congener distribution, whereas the patterns of CDD and CDF distribution generally showed some degree of difference among the other fish collected from the Neckar River. Perch and bream collected from various locations in the vicinity of Hamburg Harbor, however, showed similar patterns in the distribution of the Cl₄ to Cl₈ CDD and CDF congener groups (Gotz et al., 1990). In general, the levels of CDFs were higher than the level of CDDs in these fish, especially with regard to the TCDFs to HxCDFs. Pooled samples of eels collected at six different localities in the Netherlands contained low levels of CDDs and CDFs, the major congeners of which were 2,3,7,8-chlorine substituted (Van den Berg et al., 1987). The concentrations of the various congeners identified in the eel samples ranged from 0.1 to 9.1 ppt. The sample with the highest concentration of 2,3,7,8-TCDD (9.1 ppt) was collected from Broekervaart in a location that was not far from a chemical waste dump that contained high concentrations of the same congener.

Some general observations for CDD and CDF levels are possible from the data presented in the various fish and shellfish studies above:

• Fish and shellfish differ in their ability to bioconcentrate CDD and CDF congeners. Fish generally concentrate the most toxic 2,3,7,8-substituted congeners, but shellfish can usually concentrate most congeners.

• For fish, the concentrations of CDDs and CDFs are dependent on the exposure level, fat content, living habits, and the degree of movement of the species. Comparatively high fat-content bottom fish collected close to the contaminant source generally have the highest CDD/CDF levels, whereas lower fat content, nonstationary fish have much lower concentrations, even in the vicinity of the contaminant source.

• The National Dioxin Study indicated that the levels of 2,3,7,8-TCDD in fish from the Great Lakes Region were higher than those from urban areas. Comparable levels were detected in whole bottom feeders and predators from the Great Lakes Region.

• With regard to PCBs, concentrations increase from water to lower organisms to small fish to salmonids, and the chlorine content of the PCBs increase at the higher trophic levels.

The background fish data collected as part of EPA's National Bioaccumulation Study (EPA, 1992) were selected as the best basis for identifying background levels in U.S. fish. Sixty fish samples were collected from fresh and estuarine water at a total of 34 sites where no obvious industrial sources were present. The average TEQ (assuming zero for the nondetects) was 0.59 ppt, and 1.2 ppt (assuming half the detection limit for the nondetects). In the original study, some of the samples were analyzed on a whole body basis and others on a fillet basis. However, for purposes of this document, whole body data were converted to a fillet basis. All concentrations were expressed on a wet weight basis. This information is based on the report and clarifications provided via a personal communication from Annette Huber, Office of Water, to John Schaum, Office of Research and Development, May 17, 1993. Several points should be considered in using these estimates to assess human exposure:

• The National Bioaccumulation Study data were derived from fresh and estuarine water fish and, therefore, are not representative of open ocean fish. Some types of open ocean fish such as tuna and sword fish are commonly eaten. The cod and haddock data from Schecter et al. (1993) could be representative of these species. (See discussion in Section 4.6.2.) These data showed a range of 0.023 to 0.13 ppt. Presumably, the contaminant levels in open ocean fish are lower than fresh water and estuarine fish because they live in waters farther from dioxin sources.

• Whole fish contaminant levels are normally twice as high as fillets, which is generally considered the edible portion. However, some small fish and shell fish, such as clams, are typically eaten whole.

• These "background" samples may not be representative of what many individuals consume. EPA (1992) found an average of 11 ppt of TEQ across all 314 locations sampled. Even though these other locations were near industrial point sources, recreational or subsistence fishermen from local populations may consume fish from these waters.

• Market basket surveys would probably provide the best information on dioxin levels in fish commonly consumed by the general population. Data of this type were provided by Schecter et al. (1993). As discussed in Section 4.6.2., this study analyzed five fish collected from a supermarket and found an average of 0.05 ppt of TEQ. These data may mean that fish exposure levels are lower than the value selected here as representative of background levels (i.e., 1.2 ppt). However, only a limited number of samples were analyzed. Also the fish samples represent ocean species, whereas the National Bioaccumulation Study sampled freshwater and estuarine fish.

Dietary intake is generally recognized as the primary source of human exposure to CDD/Fs (Rappe, 1992). Several studies have estimated that over 90 percent of the average daily exposure to CDD/Fs are derived from foods (Rappe, 1992; Henry et al., 1992; Fürst et al., 1991). CDD/Fs in fatty foods such as dairy, fish, and meat products are believed to be the major contributors to dietary exposures (Rappe, 1992; Henry et al., 1992). Travis and Hattemer-Frey (1991), using a fugacity model, estimated that the food chain, especially meat and dairy products, accounts for 99 percent of human exposure to 2,3,7,8-TCDD.

Analysis of trace levels of CDD and CDF congeners in food has in the past been hindered by lack of sensitive analytical detection methods, extraction difficulties from the high-lipid content food products in which these chemicals are most often found, and the presence of other potentially interfering organochlorine compounds. As the analytical difficulties associated with detecting CDD and CDF congeners at ppt levels or lower are overcome (Firestone, 1991), more food data should be generated.

Tables B-11 and B-12 (Appendix B) contain summaries of data from the recent published literature regarding concentrations of CDDs and CDFs in food products. Most of the selected studies investigated "background" levels of CDDs and CDFs rather than studies targeted at areas of known contamination. Table B-13 contains a summary of PCB congener concentrations in food products.

The studies summarized in Tables B-11 and B-12 primarily examined CDD and CDF levels in products of animal origin (i.e., fish, meat, eggs, and dairy products). Because of their lipophilic nature, CDDs and CDFs are expected to accumulate in these food groups. The data in the tables indicate that CDDs and CDFs are found at levels ranging from the intermediate ppq up to the low ppt range. As expected, the highest levels reported are those measured in foods with high animal fat content. The highest reported congener concentrations are for the HpCDDs and OCDD. In general, for the less-chlorinated congener groups (i.e., Cl₄ - Cl₆), the CDF levels measured were larger than the CDD levels but were still within an order of magnitude. The situation is reversed for the Cl₇ and Cl₈ congener groups.

In the past, low levels of CDDs and CDFs have been detected in bleached paper. (See discussion in Chapter 3.) Because bleached paper is sometimes used for food packaging, concern has been expressed that CDD/Fs may migrate from the paper into the food.

Using refined and highly sensitive analytical methods, LaFleur et al. (1990) observed the migration of 2,3,7,8-TCDD; 2,3,7,8-TCDF; and 1,2,7,8-TCDF from bleached paper milk cartons into whole milk. After 12 days of exposure, 6.7 percent of the 2,3,7,8-TCDD; 18 percent of the 2,3,7,8-TCDF; and 13 percent of the 1,2,7,8-TCDF in the milk carton leached into the milk. The concentrations of the three congeners in milk were 8.5, 110, and 49 pg/kg for 2,3,7,8-TCDD; 2,3,7,8-TCDF; and 1,2,7,8-TCDF, respectively. [Note: These data are not reported in Appendix B; only data for raw milk are reported.]

The study results reported by LaFleur et al. (1990) were performed by the National Council of the Paper Industry for Air and Stream Improvement (NCASI) at the request of the U.S. Food and Drug Administration (FDA) as part of a cooperative Federal agency effort to assess the risks posed by dioxin contamination of paper products (i.e., the Federal Interagency Working Group on Dioxin-in-Paper). In addition to assessing the migration of CDDs and CDFs from milk cartons, studies were also conducted to assess the extent of CDD/CDF migration into food from coffee filters, cream cartons, orange juice cartons, paper cups for hot beverages, paper cups for soup, paper plates for hot foods, dual ovenable trays, and microwave popcorn bags. Migration of CDD/Fs from the paper into food was observed in all studies.

The FDA report presented data on direct measurements in these paper articles, showing TCDD and TCDF levels in the 1 - 13 ppt range. These levels were similar to the levels measured in bleached wood pulp which averaged about 8 ppt at the time of the study. As discussed in Section 3.2, the paper industry has made process changes that they expect have generally reduced dioxin levels in bleached paper pulp to less than 2 ppt of TEQ. Similar or lower levels could be expected in final paper products. NCASI reports that essentially no detectable migration of dioxin to milk occurs from cartons at these levels.

The results of these migration studies and an assessment of the risks to the general population posed by migration from paper are addressed in detail in U.S. EPA (1990a). The CDD/CDF levels currently found in food due to any leaching of dioxin-like compounds from paperboard containers are expected to be significantly lower than those reported in U.S. EPA (1990a) because of process changes implemented by the pulp and paper industry to reduce formation of CDDs and CDFs.

The published data on measured levels of CDDs, CDFs, and dioxin-like compounds in U.S. food products have generally come from studies of a specific food product(s) in a specific location(s) rather than from large survey studies designed to allow estimation of daily intake of the chemicals for a population. For example, CDD/Fs are not routinely monitored in the U.S. Food and Drug Administration's (FDA) Surveillance Monitoring Program for domestic and imported foods (conversation between Dr. S. Page, FDA, and G. Huse, Versar, Inc., February 8, 1993) nor are they routinely monitored by the U.S. Department of Agriculture (USDA) in the National Meat and Poultry Residue Monitoring Program (conversation between Dr. E. A. Brown, USDA-FSIS, and G. Schweer, Versar, Inc., February 8, 1993).

However, USDA has developed some site-specific, though dated (late 1970s), CDD monitoring data. These efforts were in response to a decline in general health noted by inspectors in several cattle herds in Michigan. Wood products in the local barns and other cattle holding premises, presumed to be treated with pentachlorophenol (PCP), were suspected as the cause of this health decline (Buttrill et al., no date; Tiernan et al., 1978). PCP was suspected to contain trace CDD and CDF levels as manufacturing contaminants at that time. In response to this incident, two national investigations were performed by USDA. The first study involved the analysis of peritoneal adipose and liver samples collected from beef cattle in 23 States (Tiernan et al., 1978), while the second study involved the analysis of adipose tissue samples (body region not specified) collected from dairy cattle in 30 States--neither study specified the cattle breeds for any sample. HxCDD, HpCDD, and OCDD were screened for in the analyses of samples from each study. In the beef cattle study (Tiernan et al., 1978), 220 samples were analyzed: 189 peritoneal adipose samples and 31 liver samples. No residues were detected in any liver samples. A total of 19 (i.e., 10 percent) of the 189 adipose samples were found to positively contain HxCDD, HpCDD, or OCDD levels >0.10 ppb, while 56 (i.e., 30 percent) contained levels <0.10 ppb that were detectable based on the signal-to-noise ratio of the analytical instrumentation. OCDD accounted for the majority of the samples that positively contained CDDs (i.e., 17 or 9.0 percent) while only 3 samples contained HxCDD and 2 samples contained HpCDD residues, respectively. A total of 358 adipose samples were analyzed in the dairy cattle study (Buttrill et al., no date). Nine samples (i.e., 2.5 percent) positively contained CDD levels >0.19 ppb or the "level of reliable measurement", while another 30 samples (i.e., 8.4 percent) contained CDD levels that were identifiable yet below the "level of reliable measurement" (i.e., not positively identified due to low concentration levels). As with the beef cattle study results, OCDD accounted for the majority (eight) of positive samples. HpCDD was identified in only a single sample that also contained OCDD. HxCDD was identified as well in only a single sample. The data from the USDA studies are not useful for estimating CDD/F exposure for two reasons. First, the samples were analyzed for only 3 of the 17 CDD/F congeners with dioxin-like toxicity, and these were reported on a homolog basis rather than a congener-specific basis. Second, the limit of detection was at or above 0.1 ppb or 100 ppt. Background levels for individual congeners appear to be much less than 100 ppt. For example, the highest congener levels in beef fat analyzed by Fürst et al. (1990) were 5.4 ppt for OCDD.

FDA has also conducted some limited analyses for the higher-chlorinated dioxins in market basket samples collected under FDA's Total Diet Program (Firestone et al. 1986). Food samples found to contain PCP residues >0.05 m g/g were analyzed for 1,2,3,4,6,7,8-HpCDD and OCDD. Also, selected samples of ground beef, chicken, pork, and eggs from the market basket survey were analyzed for these dioxin congeners, regardless of the results of PCP residue analysis. A total of 16 ground beef samples, 18 pork samples, 16 chicken samples, and 17 eggs samples with no PCP contamination were collected between 1979 and 1984 at various locations throughout the United States and analyzed for 1,2,3,4,6,7,8-HpCDD and OCDD. No dioxin residues were detected in any of the ground beef or egg samples. OCDD was observed at detectable concentrations in only 2 of the 18 pork samples (27 ppt 53 ppt) and 2 of the 16 chicken samples (29 ppt, 76 ppt). One chicken sample with PCP residues >0.05 m g/g had detectable residues of both 1,2,3,4,6,7,8-HpCDD (28 ppt) and OCDD (252 ppt). Egg samples from Houston, Texas and Mesa, Arizona with PCP residues >0.05 µg/g had detectable 1,2,3,4,6,7,8-HpCDD levels ranging from 21 ppt to 588 ppt, and OCDD levels ranging 80 ppt to 1610 ppt. These levels were attributed to local PCP contamination (Firestone et al., 1986). Milk samples contaminated with PCP at levels ranging from 0.01 m g/g to 0.05 m g/g PCP contained no detectable dioxins. It should be noted that these food residue data were not used in this assessment of dioxin exposures in the United States because the reported limits of detection (10 to 40 ppt) for the FDA analyses were considerably higher than the levels of dioxins observed in foods from more recent studies. Also, the study only analyzed for residues of 2 of the 17 toxic CDD/CDF congeners. Finally, the study focussed on samples with PCP contamination and, therefore, was not generally representative of background exposures.

The primary sources of information on background levels of CDD/Fs in U.S. foods are studies conducted by the California Air Resources Board (CARB), the results of background analysis from the NCASI study (Stanley and Bauer, 1989; LaFleur et al., 1990) and Schecter et al. (1993). Each of these three studies is summarized below.

CARB collected multiple samples of seven types of foods from commercial food sources in two urban areas of California (Stanley and Bauer, 1989). Foods were collected randomly, but an emphasis was placed on food stuffs of California origin (Stanley and Bauer, 1989). The types of food stuffs included saltwater fish, freshwater fish, beef, chicken, pork, milk, and eggs. A total of 210 samples were collected in Los Angeles (30 individual samples of each of the 7 types of foods), and 140 samples were collected in San Francisco (20 individual samples of each of the 7 types of foods). Food items were composited before chemical analysis to obtain a sample that was representative of average levels of PCDDs and PCDFs in the food stuffs, increase the probability of detection, and reduce the cost of chemical analysis. Samples were composited separately for each type of food stuff, within each geographical area. Each composite sample contained 6 to 10 individual food samples, and 5 to 8 composite samples were analyzed for each food type. Beef (ground beef), pork (bacon), and chicken samples were analyzed on a lipid weight basis, but were subsequently converted to a wet weight basis, for the purposes of this report, by multiplying the lipid weight concentration of CDD/CDFs by the fraction of fat contained in the food product of interest. Milk and fish samples were also analyzed on a lipid weight basis. Egg samples were analyzed for CDD/CDFs on a wet weight basis. The CARB data are summarized in Table 4-2.

The NCASI study (as described by LaFleur et al., 1990; and Henry et al., 1992) collected random food samples directly from the shelves of grocery stores located in the southern, midwestern and northwestern regions of the United States. The samples were analyzed for 2,3,7,8-TCDD and 2,3,7,8-TCDF. These data are summarized in Table 4-3.

Schecter et al. (1993) conducted a complete congener analyses of 18 food samples collected directly from a supermarket in Binghamton, New York in early 1990. The samples included five fish, three types of beef (ground beef, beef sirloin tip, and beef rib steak), one chicken drumstick, one porkchop, one lamb, one ham, one bologna, one heavy cream, and four types of cheese. The following ranges of TEQ levels on a whole weight basis were found: fish: 0.01 - 0.13 ppt; meat: 0.03 - 1.5 ppt; and dairy products: 0.04 - 0.7 ppt. These data are summarized in Table 4-4.

Background TEQ concentrations of CDD/Fs in beef/veal and pork were estimated using data from eight CARB samples (Stanley and Bauer, 1989), three NCASI background samples (LaFleur et al., 1990), and three samples from Schecter et al. (1993). The CARB and Schecter et al. samples were analyzed for 16 2,3,7,8-substituted CDD/F congeners; the NCASI samples were analyzed for 2,3,7,8-TCDD and 2,3,7,8-TCDF only. At least one congener was detected in 13 of the 14 composite beef samples. One sample had no detectable congeners. The congeners most frequently detected in beef/veal were 1,2,3,4,6,7,8-HpCDD and OCDD, and only one congener was not detected in any of the samples. For the purposes of this report, the total whole weight TEQ for beef was calculated by assuming that the lipid content of beef was 19 percent and by using one-half the detection limits to represent the concentration of nondetectable CDD/F congeners in the samples. Using this methodology, the total background TEQ was estimated to be 0.48 ppt for beef on a wet weight basis. If nondetectable concentrations are assumed to be zero; the estimated total TEQ for beef is estimated to be 0.29 ppt. All of the pork samples analyzed by CARB, NCASI, and Schecter et al. (1993) had at least one 2,3,7,8-substituted CDD/F at detectable concentrations. The hepta- and octa-chlorinated dioxins and furans were detected most frequently, and only one congener was not detectable in any of the samples analyzed. Using an assumed lipid content of 15 percent and one-half the detection limit to represent nondetectable concentrations, the estimated total whole weight TEQ for pork is 0.26 ppt. If nondetectable concentrations are assumed to be zero, the estimated TEQ for pork is 0.10 ppt. Therefore, the TEQ concentration for pork is expected to be between 0.10 ppt and 0.26 ppt.

Background TEQ concentrations for chicken are based on data from CARB (Stanley and Bauer, 1989) and Schecter et al. (1993). Nine composite chicken samples were analyzed for 16 2,3,7,8-substituted CDD/F congeners. All chicken samples contained detectable concentrations of at least one CDD/F congener. The hepta-chlorinated CDD/Fs and OCDD were detected most frequently, and 5 of the 16 congeners were not detected in any of the 9 chicken samples. The total background whole weight TEQ for chicken is estimated to be 0.19 ppt using an assumed lipid content of 15 percent and one-half the detection limit to represent the concentration of nondetectable congeners. Using zeros to represent nondetectable concentrations, the estimated total whole weight TEQ for chicken is 0.07 ppt. Therefore, the TEQ for chicken is expected to be between 0.07 ppt and 0.19 ppt. Background TEQs for eggs are based on data from CARB (Stanley and Bauer, 1989). For eggs, seven out of eight composite samples had no detectable concentrations of 2,3,7,8-substituted CDD/F congeners. Only one sample contained detectable concentrations of OCDD, 2,3,7,8-TCDF, and 1,2,3,4,6,7,8-HpCDF. Using one-half the detection limit for nondetectable concentrations, the estimated total TEQ for eggs is 0.14 ppt; but using zero for the nondetectable concentrations, the estimated TEQ for eggs is only 0.0004.

Background levels of CDD/Fs in U.S. milk are based on a very limited data set. CARB (Stanley and Bauer, 1989) analyzed eight packaged milk samples and found an average of 0.06 ppt TEQ. Becuase these samples may have been impacted by leaching of CDD/Fs from packaging materials, they were eliminated from the analysis of background concentrations conducted for this report.

LaFleur et al. (1990) analyzed a single background milk sample for 2,3,7,8-TCDD and 2,3,7,8-TCDF. The sample contained 2,3,7,8-TCDD at a concentration of 0.0018 ppt and nondetectable concentrations of 2,3,7,8-TCDF. Based on the LaFleur et al. (1990) data, the TEQ for these two congeners is estimated to be 0.0018 ppt whether one-half the detection limit or zero is used to represent the nondetectable concentration of 2,3,7,8-TCDF.

EPA (1991b) collected milk samples from several sites in the vicinity of a municipal waste incinerator in Rutland, Vermont, and two background samples from a dairy farm 123 kilometers from the incinerator where no obvious industrial sources of CDD/F were present. All samples were taken from bulk storage tanks at the farms. The report indicated that facility emissions could not be correlated with the levels of CDD/F and other contaminants measured in various environmental media. For all milk samples, the majority of the congeners were not detected. It was reported that only OCDD was consistently detected at levels from 0.2 to 2.4 pg/g in the farms near the facility. The levels in milk from the three farms near the facility ranged from about 0.2 to 0.4 pg of TEQ/g whole milk, and the TEQ for the background samples collected from the distant farm was 0.12 pg/g. The TEQs were calculated by EPA (1991b) by setting the nondetects equal to the detection limit. The 0.12 ppt TEQ background value estimated by EPA is nearly 2 orders of magnitude higher than the TEQ for milk based on the NCASI data. (This is probably due largely to the incomplete congener analysis conducted by LaFleur et al.) Examination of the raw data supporting this study indicated that all of the CDD/F congeners in the background sample were nondetectable. Consequently, if nondetects are set to zero, the total background TEQ for milk would be zero. If half the detection limits are used to calculate the total TEQ level, the estimated value is 0.07. Therefore, the total background TEQ level for milk is expected to be between zero and 0.07 pg/g.

Some idea of the total TEQ level of CDD/F in milk samples can be gained by assuming that levels in beef fat are similar to levels in milk fat. This assumption implies that the differences in feeding/raising practices of dairy cattle vs. beef cattle do not cause substantial differences in CDD/F exposure. Beef contains approximately 20 percent fat, and whole milk is about 4 percent fat. Thus, on a whole food basis, CDD/F levels in beef should be about five times higher than in milk. Support for this concept can be seen in the German data presented in Table 4-2. This table shows that the TEQ level in milk fat is 1.35 ppt and in beef fat is 1.08 ppt. On this basis, the North American data for beef (0.48 ppt of TEQ) suggest that milk would be about 0.1 ppt of TEQ. This lends support to the background level reported in EPA (1991b) of 0.12 ppt, based on only two samples.

Schecter et al. (1992) reported on the analysis of 2,3,7,8-substituted CDD/Fs in U.S. dairy products. Cottage cheese, soft cream cheese, and American cheese samples were selected randomly from New York supermarkets and analyzed on a wet-weight basis. All of the dairy products sampled had at least 13 detectable congeners out of the 17 evaluated. Only one congener (1,2,3,7,8,9-HxCDF) was not detectable in any of the five dairy products. Based on these data, the total background TEQ concentration of dairy products is estimated to be 0.36 ppt if one-half the detection limit is used to represent nondetectable concentrations and 0.35 if zero is used to represent nondetectable concentrations.

Data on CDDs and CDFs in U.S. fruit and vegetable products are extremely limited. The Ministry of the Environment, Ontario, conducted a study of CDDs and CDFs in locally produced and imported fruits and vegetables, some of which originated in the United States (Ministry of the Environment, 1988; Birmingham et al., 1989). Samples of fresh apples, peaches, potatoes, tomatoes, and wheat products were analyzed. In general, the minimum detection limits for these analyses were less than 1 ppt. The report indicated that "fruit and vegetable samples were substantially free of PCDD and PCDF residues, especially the more toxic tetra, penta, and hexachlorinated forms" (Ministry of the Environment, 1988). OCDD was the only congener detected in any of the samples. One apple and one peach sample contained detectable OCDD concentrations (8 ppt and 0.6 ppt, respectively). Detectable OCDD concentrations were found at concentrations ranging from 1 to 3 ppt in potatoes and 0.6 to 0.7 ppt in wheat samples. None of the tomato samples contained detectable levels of any CDD or CDF congeners. Based on these results, Birmingham et al. (1989) estimated the TEQs for fruits, vegetables, and wheat products to be 0.004 ppt, 0.002 ppt, and 0.0007 ppt, respectively.

As discussed in Volume III, dioxin contamination of fruits and vegetables is thought to occur primarily via particle deposition or vapor adsorption onto outer layers with little penetration to inner portions. Plant uptake from the soil via the roots is generally considered negligible. However, the work of Hülster and Marschner (1993) indicates that zucchini and pumpkins were exceptions. For these plant species, it appears that root uptake occurs and leads to a uniform concentration within the fruit. The concentration of CDDs and CDFs in zucchini squash grown on "uncontaminated" soil (0.4 ppt TEQ soil concentration) ranged from 0.5 to 0.7 ppt TEQ dry weight. These reported values may be converted to whole weight TEQ concentrations by using an assumed moisture content of 93.7 percent (USDA, 1979-1984). The resulting range of whole weight concentrations for zucchini is 0.03 to 0.04 ppt TEQ. Müller et al. (1993) also evaluated CDDs and CDFs in vegetables (carrots, lettuce, and peas) grown at both contaminated plots and control plots. For the control plots, the highest levels of CDDs and CDFs were observed in carrot peels: 0.55 ppt TEQ dry weight, or 0.07 ppt TEQ whole weight, assuming a moisture content for carrots of 87.8 percent (USDA, 1979-1984). Lower concentrations were observed in samples from the cortex of the carrots, indicating that the "contamination source for the peel of carrots is the soil" (Müller et al., 1993). Lettuce concentrations ranged from 0.1 to 0.4 ppt TEQ dry weight. This is equivalent to a whole weight concentration range of 0.005 to 0.018 ppt TEQ, assuming a moisture content of 95.4 percent for lettuce (USDA, 1979-1984). Concentrations in peas from contaminated plots ranged from 0.04 to 0.12 ppt TEQ dry weight (0.004 to 0.013 ppt TEQ whole weight, assuming a moisture content of 88.9 percent). Lower concentrations in peas (i.e., close to the detection limit; exact value not given) were reported for control plots. Similar data for vegetables grown in the United States were not available.

The high fat levels in vegetable oil suggest that it may be important to consider as a source of human exposure. Vegetable oils can be made from a variety of plants including corn, olives, peanuts, sunflower seeds, safflower seeds, linseed, and cotton seed. Many of these items are protected from atmospheric deposition, which implies that their CDD/F levels would be low. However, Theleen (1991) estimated that vegetable oil could contribute about 10 percent of a person's total daily intake in the Netherlands (14 of 120 pg TEQ/d). This estimate was based on the Fürst et al. (1990) study that found nondetects for most congeners except some of the higher chlorinated congeners of CDD and CDF (detection limit = 0.5 ppt). Half the detection limit was used for the nondetects, and most of the congeners were not detected. Consequently, the actual value could be much lower. No data could be found on CDD/F levels in vegetable oil in North America.

The U.S. food data are summarized in Table 4-5. The background TEQ estimates are presented first assuming that nondetects equal half the detection limits and second assuming that nondetects equal zero. For food groups such as eggs, a wide range of TEQ estimates are seen indicating a high percent of nondetects among individual congeners. The upper mean TEQ estimates are generally comparable to the TEQ estimates derived from studies conducted in Germany and Canada (as discussed below). These studies did not report TEQs based on assuming nondetects equal to zero but did report many nondetects in some food groups. In summary, the limited number of U.S. food samples and the high incidence of nondetects make an uncertain basis for estimating national background levels. However, the general agreement with food level estimates reported for Canada and Germany provides some reassurance that these U.S. values are reasonable. It is clear, however, that a large survey is needed to confirm residue levels of CDD/F in the U.S. food supply. For the purposes of calculating background exposures to CDD/Fs via dietary intake, the upper-range background TEQs (i.e., those calculated using one-half the detection limit for the nondetects) were used (See Chapter 5.)

Relatively extensive multiyear surveys of the levels of CDDs and CDFs in food are being undertaken in Sweden and the United Kingdom (de Wit et al., 1990 and Startin et al., 1990). The most extensive investigations reported to date that involve testing of a variety of randomly selected food samples collected within the framework of official food control have been performed in the Federal Republic of Germany (Beck et al. 1989; Fürst et al., 1990). The detailed results of these studies are included in Appendix B. Fürst et al. (1990) analyzed 107 food samples collected in Germany. The results of this study are presented in Table 4-6. All samples, except some of the milk, were randomly collected during official food monitoring programs. The authors speculated that a source may have been near the areas where the milk samples were collected because they appeared higher than other milk tested in Germany which showed levels around 1 ppt TEQ. In a later report, Fürst et al. (1991) reported that a much larger survey of dairies in Germany had been completed. This survey analyzed 168 samples of milk and milk products collected at dairies prior to bottling. They found an arithmetic mean of 1.35 pg of TEQ/g of fat. TEQs in these studies were estimated by assuming that nondetects equalled half the detection limits. The percent detected was not reported. Fürst et al. (1991) provided a summary of the results of several European studies. The data summaries relevant to background levels in meat and dairy products from Fürst et al. (1991) are presented in Table 4-7. Fürst et al. (1991) report that information on CDD and CDF levels in vegetables and fruits is scarce and that the available data indicate a background of below 1 ppt.

Beck et al. (1989) analyzed 12 food samples collected randomly from food markets in West Berlin, Germany. Chicken, eggs, butter, pork, ocean perch, cod, herring, vegetable oil, cauliflower, lettuce, cherries, and apples were analyzed for CDD/Fs. CDD/Fs were detected in samples of animal origin in the ppq to ppt range (fat weight basis). No CDD/F congeners were detected at a detection limit of 0.01 ppt (whole weight basis) in samples of plant origin.

Theelen et al. (1993) collected food products from various locations in the Netherlands and analyzed them for 2,3,7,8-chlorine substituted dioxins, furans, and planar PCBs. Meat samples were collected from slaughter houses throughout the Netherlands. Fish, mixed meats, and cheeses were gathered at various grocery stores. Mixtures of foods in these categories were prepared based on the proportion of the average annual consumption rate that different food items in these categories represented. The food industry provided purified oils and fats. Mixtures of these items were also prepared in proportion to their annual use in the Netherlands. The concentration of CDD/Fs in these food products are presented in Table 4-8.

Birmingham et al. (1989) analyzed CDD/F residues in food collected in Ontario, Canada. Most of the food was grown in Canada, although some was from the United States. They reported analyzing 25 composite samples from 10 food groups. The precise number of samples in each food group was not reported. No TeCDD, PeCDD, HxCDD, TeCDF, or PeCDF were found at detection limits of 0.1 to 7 ppt. Low ppt levels of some of the higher chlorinated CDD/Fs were detected in some foods. TEQ levels were also estimated for the major food groups. However, as shown in Table 4-9, these data were reported on a homolog basis. It is unclear what procedure was used to convert the homolog data to TEQ. The text implies that nondetects were treated as zero for purposes of estimating TEQ. In addition to the animal food data shown in Table 4-9, measurements were also made in potatoes, apples, tomatoes, peaches, and wheat. Only OCDD was detected at levels ranging from 0.6 to 8 pg/g fresh weight. The TEQ totals for vegetables were reported as 0.004 ppt for fruit, 0.002 ppt for vegetables, and 0.0007 ppt for wheat- based products. The procedure used to develop these TEQ estimates was not clear.

Tables B-14 through B-16 (Appendix B) contain summaries of data from studies of ambient air measurements of CDDs, CDFs, and PCBs in the United States and Europe. Environmental levels of PCBs in air are based on a single source of information (Hoff et al., 1992). Relatively few studies have been conducted to measure ambient air levels of CDDs/CDFs because of the low analytical detection limits required to detect the expected low concentrations of specific CDD/CDF congeners. These detection limits in ambient air samples were not achieved until the mid 1980s. To obtain subparts-per-trillion levels of analytical detection, sampling relatively large volumes of air (e.g., 350 to 450 cubic meters of ambient air over a 24-hour period) is required. The results of several of these recent studies are summarized in the following paragraphs.

The most extensive ambient air monitoring study of CDDs/CDFs conducted to date is a multiyear monitoring effort conducted at eight sampling locations in the Southern California area by the Research Division of the California Air Resources Board from December 1987 through March 1989 (Hunt et al., 1990). The monitoring network "included a number of sites situated in primarily residential areas (San Bernadino, El Toro, and Reseda), as well as several sites in the vicinity of suspected sources of CDDs/CDFs (Cal. Trans, Commerce, North Long Beach, and West Long Beach)." The seven sites mentioned above were classified as urban locations by the definitions used in this document, while one site was classified as an industrial site (i.e., Carson--on site at manufacturer of gas cooking equipment). Additionally, four of the eight sites were part of the South Coast Air Quality Management District (SCAQMD) monitoring network. All totaled, there were nine sample collection intervals throughout this study. "Typically, five to seven stations were in contemporaneous operation during a particular session" (i.e., samples were not collected from each location at each interval). Total tetra- through octa- chlorinated CDDs and CDFs were screened for in the study as well as various 2,3,7,8-substituted CDD and CDF congeners. A total of 34 analyses were performed throughout the study for all congeners except for OCDD and OCDF, respectively, for which only 31 analyses were performed. Samples were collected over a maximum of seven intervals at each site throughout the study (i.e., Reseda and El Toro--six dates, duplicate samples on one date), while a sample was collected from the Commerce site during only a single collection interval. Sample collection intervals generally averaged 24 hours in duration.

Generally, higher substituted CDD and CDF congeners accounted for the majority of positive samples containing quantifiable CDD/CDF residues in this study (i.e., Total HxCDD/HxCDF and above). In fact, over 90 percent of the samples collected contained quantifiable levels of 1,2,3,4,6,7,8-HpCDD, Total HpCDD, and OCDD. Additionally, approximately 50 to 70 percent of the samples collected contained quantifiable levels of Total HxCDD; 2,3,7,8-TCDF; Total TCDF; Total PeCDF; Total HxCDF; 1,2,3,4,6,7,8-HpCDF; Total HpCDF; and OCDF. For all other congeners, quantifiable residues were detected in less than 25 percent of the samples collected. All CDD congener concentrations ranged from nonquantifiable levels (low limit of 0.0026 pg/m³) to an upper limit of 18.0 pg/m³. Additionally, CDF congener levels ranged from nonquantifiable levels (low limit of 0.0040 pg/m³) to an upper limit of 2.70 pg/m³.

According to Hunt et al.(1990), "The highest concentration of CDDs/CDFs congener class sums (Cl₄-Cl₈) and 2,3,7,8-substituted species were noted during a period predominated by off-shore air flows in December 1987, suggesting a regional air mass and transport phenomena. Concentrations of the CDDs/CDFs were diminished markedly in subsequent sessions where air flow patterns were primarily off-shore or of coastal origin." Hunt et al. (1990) indicated that the "CDD/CDF congener profiles (Cl₄-Cl₈) and 2,3,7,8-substituted isomeric patterns strongly suggest combustion source influences in the majority" of the samples collected.

In a long-term study of CDD/Fs in the ambient air around Bloomington, Indiana, methods were developed for measuring individual CDD/Fs at concentrations as low as 0.001 pg/m³ (Eitzer and Hites, 1989). Total CDD/F concentrations were 0.480 pg/m³ and 1.360 pg/m³ for the vapor phase and the particle-bound phase, respectively. For individual congeners, CDFs were found to decrease in concentration with increasing levels of chlorination, and CDD concentrations were found to increase with increasing levels of chlorination (Eitzer and Hites, 1989).

Clayton et al. (1993) conducted a study of CDDs and CDFs in the ambient air of three major cities and an industrial town in the United Kingdom. The annual average TEQ concentrations of CDDs and CDFs ranged from 0.04 to 0.10 pg/m³. The hepta- and octachlorinated dioxin congeners contributed the most to the total concentration of 2,3,7,8-substituted CDD/Fs, and a large number of nondetect values were reported for the tetra-, penta-, and hexachlorinated dioxins. The congeners that contributed most to the total TEQ concentrations were 2,3,7,8-TCDF; 1,2,3,4,7,8-; 1,2,3,6,7,8-; and 2,3,4,6,7,8-HxCDF. These values are relatively consistent with the concentrations in ambient German air observed by Liebl et al. (1993) and König et al. (1993a). Liebl et al. (1993) analyzed ambient air samples collected from 10 sites in Hessen, Germany, from 1990 through 1992. Concentrations ranged from 0.04 to 0.15 pg TEQ/m³. The higher concentrations were presumed to result from direct local industrial sources. König et al. (1993a) collected air samples from six sites located in Hessen, Germany. CDD/F concentrations ranged from 0.048 pg TEQ/m³ at a rural reference site to 0.146 pg TEQ/m³ at an industrial site. The results of the study also indicated that concentrations of CDDs and CDFs are typically higher in the winter than in the summer. Sugita et al. (1993) also observed higher concentrations of CDDs and CDFs in winter than in summer in an ambient air study in urban Japan. The average concentration of CDDs and CDFs was 0.788 pg TEQ/m³ in the summer and 1.464 pg TEQ/m³ in winter.

In a Swedish study, air samples were collected from a city center, suburb, remote countryside, and open coastal area (Broman et al., 1991). Analyses of the samples for dioxins and furans indicated that the concentrations of these compounds decreased with increasing distance from the city center. Total CDD/F concentrations were 1.40 pg/m³, 1.10 pg/m³, 0.40 pg/m³, and 0.22 pg/m³ for the city center, suburb, countryside, and open coastal areas, respectively. Similar patterns of decreasing concentrations with increasing distances from urban areas were also observed for individual CDD/F congeners (Broman et al., 1991). In a study of ambient air concentrations of CDDs and CDFs in Flanders, samples were collected and analyzed at rural, industrial, and urban sites (Wevers et al., 1993). Average ambient air concentrations ranged from 0.0696 pg TEQ/m³ at a rural site to 0.254 pg TEQ/m³ at a site believed to be influenced by a chemical industry and a highway.

PCBs have also been evaluated in European air samples (Halsall and Jones, 1993; König et al., 1993b). Halsall and Jones (1993) monitored urban air at two sites in the United Kingdom. The annual mean total PCB concentrations were 520 and 590 pg/m³. PCBs existed in ambient air predominantly in the vapor phase. This study also indicated that summer PCB concentrations were higher than winter concentrations. These researchers attributed the differences in seasonal patterns to volatilization from soil during summer months. Ambient air concentrations of PCBs in Hessen, Germany, ranged from 350 to 1630 pg/m³ during the period of 1990 to 1992 (König et al., 1993b). Urban areas characterized by industry and/or heavy traffic had the highest PCB concentrations in ambient air.

Based on the limited ambient air measurements that have been made in selected cities in the United States and Europe, there appears to be good agreement with respect to the magnitude of specific congeners of CDDs and CDFs in urbanized areas in the United States and Europe. Most of these measurements tend to be very close to the current analytical detection limit. This increases the probability that congeners indicated as not detected (ND) may actually be present.

A total of 84 samples from the studies summarized in Tables B-14 and B-15 was selected as representative of "background" conditions in the United States. Samples collected from pristine sites and from rural and urban locations not expected to be impacted by industrial point sources were assumed to represent "background" conditions. The mean TEQ level for these 84 samples is 0.095 pg/m3 assuming that values reported as not detected are equal to one-half the detection limit.

Based on the results of European studies, ambient air concentrations of CDDs and CDFs appear to be similar to those found in the United States. For the purposes of this study, a TEQ value of 0.10 pg/m³ was used to represent concentrations in Europe. This value represents the mean of the midpoints of the European studies for which TEQ concentrations were reported (Clayton et al. 1993; Liebl et al. 1993; König et al. 1993a; Wevers et al. 1993). Data for these European studies are not included in Tables B-14 and B-15 because individual congener data were not reported.

It is interesting to compare these values with the CDD/CDF concentrations in air recently measured by Lugar (1993) in and around McMurdo Station, Antarctica, a logistics and staging facility with a population of about 1,100. Four locations were sampled: a site upwind of the station, downwind of the station, in the center of the station, and a remote unpopulated island 30 kilometers distant from the station. CDDs/CDFs were not detected in the samples from the upwind site (congener detection limits ranged from <0.01 to 0.03 pg/m³) and the remote island sites (congener detection limits ranged from 0.001 to 0.008 pg/m³) and only sporadically at the downwind site (some congeners detected in three of five samples). CDDs/CDFs were detected in all five samples collected from the station center site (mean TEQ concentration of 0.0153 pg/m3).

Small amounts of dioxin-like compounds may be formed during natural fires suggesting that these compounds may have always been present in the environment. However, it is generally believed that much more of these compounds have been produced and released into the environment in association with man's industrial and combustion practices, and as a result, environmental levels are likely to be higher in modern times than they were in prior times. However, the trend may now be reversing (i.e., releases and environmental levels may be gradually decreasing) due to changes in industrial practices (Rappe, 1992). As discussed in Chapter 3, the potential for environmental releases of dioxin-like compounds has been reduced due to the switch to unleaded automobile fuels (and associated use of catalytic converters and reduction in halogenated scavenger fuel additives), process changes at pulp and paper mills, improved emission controls for incinerators, and reductions in the manufacture and use of chlorinated phenolic intermediates and products.

Smith et al. (1992, 1993) analyzed sediment core layers from Green Lake, located near Syracuse, New York, to determine temporal trends in the deposition of CDDs and CDFs since the beginning of the industrial era (i.e., circa 1860). This deep lake (200-foot depth) is thought to be impacted only by atmospheric deposition because no industrial inputs are present and motorboats are not allowed. Relatively constant but low concentrations of CDDs and CDFs (10 ng/kg or less) are observed in sediments deposited from 1860 to 1930. However, concentrations increase rapidly thereafter, reaching a peak in the mid-1960s when total CDD concentrations exceeded 1,300 ng/kg and total CDF concentrations exceeded 250 ng/kg. The concentrations of CDDs and CDFs have rapidly declined since the mid-1960s, and now (1986-1990) are measured at 750 ng/kg as total CDD/CDF. The authors speculate that the decline may be due to the switch to unleaded fuels for vehicles. Similar trends have been reported by Czuczwa and Hites (1984) for Great Lakes sediment.

Rappe (1991) reports testing of archived soils and plants collected in southeast England between 1846 and 1986. CDDs and CDFs were found in all samples and showed generally increasing levels of dioxins. Rappe further notes that the congener pattern is typical of those for combustion sources until about 1950 when the pattern becomes more dominated by hepta- and octa-CDDs corresponding to increases in production of chlorinated compounds. Schecter (1991) analyzed ancient liver tissues (estimated to be 100- to 400-years old) recovered from frozen bodies of Native American (Eskimo) women. He found that the dioxin levels were much lower than those commonly found in livers of people currently living in industrial areas.

Studies that may be used to assess temporal trends in human exposure to dioxins and furans are extremely limited. The use of indirect exposure assessment techniques for detecting temporal trends is difficult because large-scale, long-term, nationally- representative environmental monitoring for dioxins and furans has not been conducted. Short-term studies are generally not comparable because of differences in sampling protocols and analytical techniques used in these studies. A potentially useful study for evaluating changes in human exposure over time is EPA's National Human Adipose Tissue Survey (NHATS). The purpose of NHATS is to monitor the human body burden of selected chemicals in the general U.S. population (U.S. EPA, 1991a). NHATS uses direct measurement techniques to estimate exposures. Nationwide samples of adipose tissue are collected from surgical patients and autopsied cadavers and analyzed annually. In 1982, broadscan analysis of composited adipose tissue specimens revealed that chlorinated dioxins and furans could be detected and quantified in the U.S. population across all geographic regions and age groups (U.S. EPA, 1986). In 1987, NHATS specimens were also analyzed for dioxins and furans making temporal comparisons possible. Statistical analyses were performed to determine if significant differences existed between the concentrations of these compounds in 1982 and 1987 adipose tissue specimens.

Table 4-10 presents the estimated national average concentrations for the two time periods and the relative changes from 1982 to 1987. The estimated concentrations of 1,2,3,7,8-PeCDD; 2,3,4,7,8-PeCDF; and HxCDD in human adipose tissue were significantly lower in 1987 than in 1982 (U.S. EPA, 1991a). Similar survey designs were used in the two studies, but changes in some of the analytical methods were made in 1987 that may account for some of the differences in estimated concentrations. These changes include lower limits of detection and the use of additional internal quantitation standards that provided more accurate measurements. The levels of 2,3,7,8-TCDD; 1,2,3,4,6,7,8-HpCDD; and OCDD were also lower in 1987 than in 1982, but the differences were not found to be statistically significant. No statistical comparisons were possible for 2,3,7,8-TCDF; HxCDF; 1,2,3,4,6,7,8-HpCDF; or OCDF because one or both of the annual estimates were based on data that did not meet the minimum criteria for statistical modeling (i.e., the chemical was not detected in at least 50 percent of the composites analyzed, and/or fewer than 30 composite samples were analyzed in each year). The results of this study indicate that exposure to certain dioxins and furan congeners may have decreased over this 5-year time period. However, further studies are

needed to verify that these changes are not a result of protocol changes, but actual reductions in exposures.

This chapter has summarized data on CDD/F levels in environmental media and food with emphasis on "background levels." Data representative of background conditions in environmental media are considered to be those collected in rural, pristine, and urban (air only) areas not believed to be impacted by any local sources (e.g., incinerators and highways). Only food data from the general food supply (i.e., collected from grocery stores) were used to represent background conditions (except fish). Tables B-17 through B-30 in Appendix B present the geometric and arithmetic averages of environmental background monitoring data for CDD and CDF congeners in various media, compiled from the published literature. The geometric averages are consistently lower than the arithmetic averages. This results from the fact that the data span several orders of magnitude with the distribution skewed toward the lower end due to the large number of not detected values. To calculate a total TEQ for all CDD/CDFs for each media, the arithmetic mean background concentration for each congener was multiplied by its respective TEF value, and individual TEQs for each congener were totaled. These total TEQs are presented at the end of Tables B-17 through B-30 and summarized in Table 4-11. These total background level TEQs are used in Chapter 5 to estimate typical exposure levels in the United States. Exposure levels for Europe are based on the levels of CDD/Fs in food reported by Fürst et al. (1990), and the levels in environmental media are based on data collected from several European countries.

Standard deviations of the total mean TEQs for each media were also calculated to depict the "range" of probable CDD/CDF levels in various media. Because the total TEQs were actually a summation of mean TEQs for various congeners, the use of typical methods for calculating standard deviations was not possible. Therefore, standard deviations were based on the standard deviation of the congener that contributed most to the total TEQ. The percentage deviation from the mean for that congener was applied to the total mean TEQ for all congeners combined. The congeners selected for use in the standard deviation estimates are presented in Table 4-12. The data in this table indicate that the pentachlorinated dioxins were the highest contributors to total TEQs in most foods in the United States.

The media levels presented in Table 4-11 are shown graphically in Figure 4-1. Except for the TEQ levels in European food which are based on data reported for German food by Fürst et al. (1990) and the TEQ levels in European air which are based on data repoted for air in Germany, Belgium and the United Kingdom by König et al. (1993), Liebl et al. (1993), Wevers et al. (1993), and Clayton et al. (1993), all other TEQ levels presented in Figure 4-1 are based on the data analyzed in this study. The background TEQ levels of CDD/CDFs in water and air were found to be lower than in any of the other environmental media evaluated and were not included in Figure 4-1. For most media, the average levels appear to be similar between North America and Europe. However, differences were noted in three areas:

· Sediment - The background levels in Europe were estimated to be higher than North America. It should be noted, however, that only the 2,3,7,8-TCDD/F and OCDD/F congeners were analyzed for background sediment sites in the United States and Europe. The sediment data are quite variable and can be very high in impacted areas (i.e., 2,3,7,8-TCDD levels over 1000 ppt have been measured in industrial areas). Also, it was difficult to interpret whether some of the European data truly represent unimpacted areas. Thus, these differences may be due more to the weakness of the data base and interpretation difficulties, rather than real differences.

· Dairy Products - The dairy products data suggest that North America levels are higher than European. Dairy products include a wide variety of food items with varying amounts of fat. Thus, the CDD/F levels would vary correspondingly. Differences in the mix of dairy products used for the North America and European estimates could explain these differences.

· Pork - The pork data suggest that North America levels are higher than European levels. The low number of samples collected in Europe may mean this estimate is not representative.

In general, the differences noted above probably reflect the sparseness or inequalities in the data rather than real differences. The human tissue data (discussed in Section 5.4) suggest similar body burden levels in the North America, Europe, and other industrial countries. Thus, it seems likely the media levels would also be similar. Large-scale market basket food surveys are clearly needed to confirm these levels.

CDD/Fs can enter aquatic systems directly from sources in effluent discharges, indirectly from deposition of CDD/Fs in the atmosphere onto water bodies, and in stormwater runoff from areas where dioxin-containing material has been land-applied or atmospherically deposited. For any given water body, the dominant transport mechanism will depend on site-specific conditions. Aquatic organisms will bioaccumulate CDD/Fs and thereby enter the aquatic food chain.

Based on information currently available, the primary mechanism by which dioxin-like compounds enter the terrestrial food chain is via atmospheric deposition and sorption of vapors. Deposition can occur directly onto plant surfaces or onto soil. Deposits onto the soil can enter the food chain via direct ingestion (e.g., soil ingestion by earthworms, fur preening by burrowing animals, incidental ingestion by grazing animals, etc). CDD/Fs in soil can become available to plants and thus enter the food chain by volatilization and vapor sorption or particle resuspension and adherence to plant surfaces. Although CDD/Fs in soil can adsorb directly to underground portions of plants, uptake from soil via the roots into above ground portions of plants is thought to be insignificant (McCrady et al., 1990).

Support for this air-to-food hypothesis is provided by Hites (1991) who concluded that "background environmental levels of PCD/F are caused by PCD/F entering the environment through the atmospheric pathway." His conclusion was based on demonstrations that the congener profiles in lake sediments could be linked to congener profiles of combustion sources. Further argument supporting this hypothesis is offered below:

· Numerous studies have shown that CDD/Fs are emitted into the air from a wide variety of sources and that CDD/Fs can be commonly detected in air at low concentrations. (See Chapters 3 and 4.)

· Studies have shown that CDD/Fs can be measured in wet and dry deposition in most locations including remote areas (Koester and Hites, 1992; Rappe, 1991).

· Numerous studies have shown that CDD/Fs are commonly found in soils throughout the world. (See Chapter 4.) Atmospheric transport and deposition is the only plausible mechanism that could lead to this widespread distribution.

· Models of the air-to-plant-to-animal food chain have been constructed. Exercises with these models show that measured deposition rates and air concentrations can be used to predict food levels that are similar to levels actually measured in food (Travis and Hattemer-Frey, 1991; also Chapter 7 of Volume III).

· Alternative mechanisms of uptake into food appear less plausible:

- Uptake in food crops and livestock from water is minimal due to the hydrophobic nature of these compounds. Travis and Hattemer-Frey (1987, 1991) estimate water intake accounts for less than 0.01 percent of the total daily intake of 2,3,7,8-TCDD in cattle. Experiments by McCrady et al. (1990) show very little uptake in plants from aqueous solutions.

- Relatively little impact on the general food supply is expected from soil residues that originate from site-specific sources such as sewage sludge and other waste disposal operations. Sewage sludge application onto agricultural fields is not a widespread practice. Waste disposal operations can be the dominant source of CDD/Fs in soils at isolated locations such as Times Beach, but are not sufficiently widespread to explain the ubiquitous nature of these compounds.

- The release of CDD/Fs to the environment from the use of pesticides contaminated with CDD/Fs is beleived to have declined in recent years; however, the past and current impact of pesticide use on CDD/F levels in the food supply is uncertain. CDD/Fs have been associated with certain phenoxy herbicides most of which are no longer produced or have restricted uses. EPA has issued data call-ins requiring certain pesticide manufacturers to test their products for dioxin content. The responses, so far, indicate that levels in these products are below or near the limit of quantitation. (See Chapter 3.)

- Current CDD/F levels in food resulting from the use of bleached paper products containing CDD/Fs appears to be minimal. In the early 1980s, testing showed that CDD/Fs could migrate from paper containers into food. Current levels in paper products are now much lower than in the early 1980s. Also, testing of products such as milk and beef prior to packaging has shown detectable levels which cannot be attributed to the packaging. (See Chapter 4.)

A related issue is whether the CDD/Fs in food result more from current or past emissions. Sediment core sampling indicates that CDD/F levels in the environment began increasing around the turn of the century, but also that CDF levels have been declining since about 1980 (Smith et al. 1992). Thus, CDD/Fs have been accumulating for many years and may have created reservoirs that continue to impact the food chain. As discussed in Chapter 3, researchers in several countries have attempted to compare known emissions with deposition rates. All of these studies (including this assessment) suggest that annual atmospheric depositions exceed annual emissions by a factor of 2 to 10. One possible explanation for this discrepancy in sources may be that volatilization or particle resuspension from these reservoir sources followed by atmospheric scavenging is responsible. These mass balance studies are highly uncertain, and it remains unknown how much of the food chain impact is due to current vs past emissions.

Beck, H.; Eckart, K.; Kellert, M.; Mathar, W.; Ruhl, C.S.; Wittkowski, R. (1987) Levels of PCDFs and PCDDs in samples of human origin and food in the Federal Republic of Germany. Chemosphere 16 (8/9): 1977-1982.

Beck, H.; Eckart, K.; Mathar, W.; Wittkowski, R. (1989) PCDD and PCDF body burden from food intake in the Federal Republic of Germany. Chemosphere 18 (1-6): 417-424.

Berry, R.M.; Lutke, C.E.; Voss, R.H. (1993) Ubiquitous nature of dioxins: a comparison of the dioxins content of common everyday materials with that of pulps and papers. Environ. Sci. Technol. 27(6):1164-1168.

Birmingham, B. (1990) Analysis of PCDD and PCDF patterns in soil samples: use in the estimation of the risk of exposure. Chemosphere 20(7-9): 807-814.

Birmingham, B.; Thorpe, B.; Frank, R.; Clement, R.; Tosine, H.; Fleming, G.; Ashman, J.; Wheeler, J.; Ripley, B.D.; Ryam, J.J. (1989) Dietary intake of PCDD and PCDF from food in Ontario, Canada. Chemosphere 19:507-512.

Bopp, R.F.; Gross, M.L.; Tong, H.; Simpson, M.J.; Monson, S.J.; Deck, B.L.; Moser, F.C. (1991) A major incident of dioxin contamination: sediments of New Jersey estuaries. Environ. Sci. Technol. 25: 951-956.

Boos, R.; Himsl, A.; Wurst, F.; Prey, T.; Scheidl, K.; Sperka, G.; Glaser, O. (1992) Determination of PCDDs and PCDFs in soil samples from Salzburg, Austria. Chemosphere 25(3):283-291.

Branson, D.R.; Takahashi, I.T.; Parker, W.M.; Blau, G.E. (1985) Bioconcentration kinetics of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rainbow trout. Environ. Toxicol. Chem. 4(6):779-788.

Broman, D.; Näf, C.; Rolff, C.; Zebühr, Y. (1990) Analysis of polychlorinated dibenzo-p-dioxins (PCDD) and polychlorinated dibenzofurans (PCDF) in soil and digested sewage sludge from Stockholm, Sweden. Chemosphere 21: 1213-1220.

Broman, D.; Näf, C.; Zebühr, Y. (1991) Long-term high and low volume air sampling of polychlorinated dibenzo-p-dioxins and dibenzofurans and polycyclic aromatic hydrocarbons along a transect from urban to remote areas on the Swedish Baltic coast. Environ. Sci. Technol. 25(11):1841-1850.

Buttrill, W.H.; Malanoski, A.J.; Conrey, J.S. (no date) Dioxin: a survey of dairy cattle in the United States. Internal Report completed by the Food Safety and Quality Service-Science, Washington, DC: U.S. Department of Agriculture, Food Safety and Inspection Service.

CDEP (1988) Measurement of selected polychlorinated dibenzo-p-dioxins and

polychlorinated dibenzofurans in ambient air in the vicinity of Wallingford, Connecticut. Connecticut Department of Environmental Protection, Air Compliance Unit, Hartford, CT. Project report by ERT, Concord, MA., Project # 7265-001-004. July 8,1988.

Clayton, P.; Davis, B.; Duarte-Davidson, R.; Halsall, C.; Jones, K.C.; Jones, P. (1993) PCDDs and PCDFs in ambient UK urban air. Presented at: Dioxin '93, 13th Symposium on Dioxins and Related Compounds; Vienna, Austria; September 1993.

Creaser, C.S.; Fernandes, A.R.; Al-Haddad,A.; Harrad, S.J.; Homer, R.B.; Skett, P.W.; Cox, E.A. (1989) Survey of background levels of PCDDs and PCDFs in UK soils. Chemosphere 18(1-6): 767-776.

Creaser, C.S.; Fernandes, A.R.; Harrad, S.J.; Cox, E.A. (1990) Levels and sources of PCDDs and PCDFs in urban British soils. Chemosphere 21: 931-938.

Czuczwa, J.M.; McVeety, B.D.; Hites, R.A. (1984) Polychlorinated dibenzo-p-dioxins and dibenzofurans in sediments from Siskiwit Lake, Isle Royale. Science 226: 568-569.

DeVault, D.; Dunn, W.; Bergquist, P.A.; Wiberg, K.; Rappe, C. (1989) Polychlorinated dibenzofurans and polychlorinated dibenzo-p-dioxins in Great Lakes fish: a baseline and interlake comparison. Environ. Toxicol. Chem. 8: 1013-1022.

deWit, C.; Jansson, B.; Strandell, M.; Jansson, P.; Bergqvist, P.A.; Bergek, S.; Kjeller, L.O.; Rappe, C.; Olsson, M.; Slorach, S. (1990) Results from the first year of the Swedish dioxin survey. Chemosphere 2 (10-12):1473-1480.

Edgerton, S.A.; Czuczwa, J.M.; Rench, J.D.; Hodanbosi, R.F.; Koval, P.J. (1989) Ambient air concentrations of polychlorinated dibenzo-p-dioxins and dibenzofurans in Ohio: sources and health risk assessment. Chemosphere 18(9/10):1713-1730.

Eitzer, B.D.; Hites, R.A. (1989) Polychlorinated dibenzo-p-dioxins and dibenzofurans in the ambient atmosphere of Bloomington, Indiana. Environ. Sci. Technol. 23(11):1389-1395.

Firestone, D. (1991) Determination of dioxins and furans in foods and biological tissues: review and update. J. Assoc. Off. Anal. Chem. 74(2):375-384.

Firestone, D.; Niemann, R.A.; Schneider, L.F.; Gridley, J.R.; Brown, D.E. (1986) Dioxin residues in fish and other foods. In: Chlorinated Dioxins and Dibenzofurans in Perspective; Rappe, C.; Choudhary, G., Keith, L.H., eds.; Lewis Publishing Co., Chelsea, MI. P. 355-365.

Frommberger, R. (1991) Polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans in fish from south-west Germany: River Rhine and Neckar. Chemosphere 22(1-2): 29-38.

Fürst, P.; Fürst, C.; Groebel, W. (1990) Levels of PCDDs and PCDFs in food-stuffs from the Federal Republic of Germany. Chemosphere 20(7-9):787-792.

Fürst, P.; Fürst, L.; Widmers, K. (1991) Body burden with PCDD and PCDF from food. In: Gallo, M.; Scheuplein, R.; Van der Heijden, K. eds. Biological basis for risk assessment of dioxins and related compounds. Banbury Report #35. Plainview, NY: Cold Springs Harbor Laboratory Press.

Gardner, A.M.; White, K.D. (1990) Polychlorinated dibenzofurans in the edible portion of selected fish. Chemosphere 21(1-2): 215-222.

Gotz, R.; Schumacher, E. (1990) Polychlorierte dibenzo-p-dioxine (PCDDs) und polychlorierte dibenzofurane (PCDFs) in sedimenten und fischen Aus Dem Hamburger Hafen. Chemosphere 20(1-2): 51-73.

Halsall, C.; Jones, K.C. (1993) PCBs in UK urban air. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Harless, R.L.; Lewis, R.G. (1991) Evaluation of a sampling and analysis method for determination of polyhalogenated dibenzo-p-dioxins and dibenzofurans in ambient air. Presented at: Dioxin '91, 11th International Symposium on Chlorinated Dioxins and Related Compounds, Research Triangle Park, NC; September 1991.

Harless, R.L.; McDaniel, D.D.; Dupuy, A.E. (1990) Sampling and analysis for polychlorinated dibenzo-p-dioxins and dibenzofurans in ambient air. Proceedings of the Tenth International Symposium on Chlorinated Dioxins and Related Compounds, Bayreuth, Germany, September 10-14, 1990.

Henry, S.; Cramer, G.; Bolger, M.; Springer, J.; Scheuplein, R. (1992) Exposures and risks of dioxin in the U.S. Food supply. Chemosphere 25(1-2):235-238.

Hoff, R.M.; Muir, D.C.G.; Grift, N.P. (1992) Annual cycle of polychlorinated biphenyls and organohalogen pesticides in air in southern Ontario. 1. Air concentration data. Environ. Sci. Technol. 26(2):266-275.

Huckins, J.N.; Schwartz, T.R.; Petty, J.D.; Smith, L.M. (1988) Determination, fate, and potential significance of PCBs in fish and sediment samples with emphasis on selected AHH-inducing congeners. Chemosphere 17(10): 1995-2016.

Hülster, A.; Marschner, H. (1993) Soil-plant transfer of PCDD/PCDF to vegetables of the cucumber family (Cucurbitaceae). Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Hunt, G.T.; Maisel, B. (1990) Atmospheric PCDDs/PCDFs in wintertime in a

northeastern U.S. urban coastal environment. Chemosphere 20: 1455-1462.

Hunt, G.; Maisel, B.; Hoyt, M. (1990) Ambient concentrations of PCDDs/PCDFs in the South Coast air basin. California Air Resources Board. Contract No. A6-100-32. Document No. 1200-005-700.

Jobb, B.; Uza, M.; Hunsinger, R.; Roberts, K.; Tosine, H.; Clement, R.; Bobbie, B.; LeBel, G.; Williams, D.; Lau, B. (1990) A survey of drinking water supplies in the province of Ontario for dioxins and furans. Chemosphere 20(10-12): 1553-1558.

Kitunen, V.H.; Salkinoja-Salonen, M.S. (1990) Soil contamination at abandoned sawmill areas. Chemosphere 20: 1671-1677.

Kjeller, L.O.; Kulp, S.E.; Bergetk, S.; Bostrom, M.; Bergquist, P.A.: Rappe,

C. ; Jonsson, B.; de Wit, C.; Jansson, B.; Olsson, M. (1990) Levels and possible sources of PCDD/PCDF in sediment and pike samples from Swedish lakes and rivers. (Part one). Chemosphere 20 (10-12): 1489-1496.

Koistinen, J.; Passivirta, J.; Sarkka, J. (1990) Organic chlorine compounds in lake sediments. IV. Dioxins, furans and related chloroaromatic compounds. Chemosphere 21(12): 1371-1379.

König, J.; Theisen, J.; Günther, W.J.; Liebl, K.H.; Büchen, M. (1993a) Ambient air levels of polychlorinated dibenzofurans and dibenzo(p)dioxins at different sites in Hessen. Chemosphere 26:851-861.

König, J.; Balfanz, E.; Günther, W.J.; Liebl, K.H.; Büchen, M. (1993b) Ambient air levels of polychlorinated biphenyls at different sites in Hessen, Germany. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

LaFleur, L.; Bousquet, T.; Ranage, K.; Brunck, B.; Davis, T.; Luksemburg, W.; Peterson, B. (1990) Analysis of TCDD and TCDF on the ppq-level in milk and food sources. Chemosphere 20(10-12):1657-1662.

Liebl, K.; Büchen, M.; Ott, W.; Fricke, W. (1993) Polychlorinated dibenzo(p)dioxins and dibenzofurans in ambient air; concentration and deposition measurements in Hessen, Germany. Presented at Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Lugar, R.M. (1993) Results of monitoring for PCDDs and PCDFs in ambient air at McMardo Station, Antarctica. EG&G Idaho, Inc. Idaho Falls, ID.

Maisel, G.E.; Hunt, G.T. (1990) Background concentrations of PCDDs/PCDFs in

ambient air. A comparison of toxic equivalency factor (TEF) models. Chemosphere 20: 771-778.

McKee, P.; Burt, A.; McCurvin, D.; Hollinger, D.; Clement, R.; Sutherland, D.; Neaves, W. (1990) Levels of dioxins, furans and other organic contaminants in harbour sediments near a wood preserving plant using pentachlorophenol and creosote. Chemosphere 20(10-12): 1679-1685.

Mes, J.; Newsome, W.H.; Conacher, H.B.S. (1991) Levels of specific polychlorinated biphenyl congeners in fatty foods from five Canadian cities between 1986 and 1988. Food Addit. Contam. 8(3):351-361.

Meyer, C.; O'Keefe, D.; Hilker, D.; Rafferty, L.; Wilson, L.; Connor, S.; Aldous, K.; Markussen, K.; Slade, K. (1989) A survey of twenty community water systems in New York State for PCDDs and PCDFs. Chemosphere 19(1-6): 21-26.

Ministry of the Environment (1988) Polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans and other organochlorine contaminants in food. Ontario, Canada: Ministry of Agriculture and Food.

Müller, J.F.; Hülster, A.; Päpke, O.; Ball, M.; Marschner, H. (1993) Transfer of PCDD/PCDF from contaminated soils into carrots, lettuce, and peas. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Naf, C; Broman, D.; Ishaq, R.; Zebuhr, Y. (1990) PCDDs and PCDFs in water, sludge and air samples from various levels in a waste water treatment plant with respect to composition changes and total flux. Chemosphere 20: 1503-1510.

Nestrick, T.J.; Lamparski, L.L.; Frawley, N.N.; Hummel, R.A.; Kocher, C.W.; Mahle, N.H.; McCoy, J.W.; Miller, D.L.; Peters, T.L.; Pillepich, J.L.; Smith, W.E.; Tobey, S.W. (1986) Perspectives of a large scale environmental survey for chlorinated dioxins: overview and soil data. Chemosphere 15: 1453-1460.

Niimi, A.J.; Oliver, B.G. (1989a) Assessment of relative toxicity of chlorinated dibenzo-p-dioxins, dibenzofurans, and biphenyls in Lake Ontario salmonids to mammalian systems using toxic equivalent factors (TEF). Chemosphere 18(7-8): 1413-1423.

Niimi, A.J.; Oliver, B.G. (1989b) Distribution of polychlorinated biphenyl congeners and other halocarbons in whole fish and muscle among Lake Ontario salmonids. Environ. Sci. Technol. 23: 83-88.

Norwood, C.B.; Hackett, M.; Powell, R.J.; Butterworth, B.C.; Williamson, K.J.; Naumann, S.M. (1989) Polychlorinated dibenzo-p-dioxins and dibenzofurans in selected estuarine sediments. Chemosphere 18(1-6): 553-560.

Oehme, M.; Mane, S.; Brevik, E.M.; Knutzen, J. (1989) Determination of polychlorinated dibenzofuran (PCDF) and dibenzo-p-dioxin (PCDD) levels and isomer patterns in fish, crustacea, mussel and sediment samples from a fjord region polluted by Mg-production. Fresenius Z. Anal. Chem. 335: 987-997.

Oliver, B.G.; Niimi, A.J. (1988) Trophodynamic analysis of polychlorinated biphenyl congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ. Sci. Technol. 22: 388-397.

Pearson, R.G.; McLaughlin, D.L.; McIlveen, W.D. (1990) Concentrations of PCDD and PCDF in Ontario soils from the vicinity of refuse and sewage sludge incinerators and remote rural and urban locations. Chemosphere 20: 1543-1548.

Petty, J.D.; Smith, C.M.; Bergquist, P.A.; Johnson, J.L.; Stalling, D.L.; Rappe, C. (1982) Chlorinated dioxins, dibenzofurans total environment, [Proc. Symp.], 1982. Edited by Choudhary, Keith, Rappe, and Butterworth. Boston, Massachusetts.

Prats, D.; Ruiz, F.; Zarzo, D. (1992) Polychlorinated biphenyls and organo chlorine pesticides in marine sediments and seawater along the coast of Alicante, Spain. Marine Pollution Bulletin 24(9):441-446.

Rappe, C. (1991) Sources of human exposure to PCDDs and PCDFs. In: Gallo, M.; Scheuplein, R.; Van der Heijden, K. eds. Biological basis for risk assessment of dioxins and related compounds. Banbury Report #35. Plainview, NY: Cold Spring Harbor Laboratory Press.

Rappe, C. (1992) Sources of PCDDs and PCDFs. Introduction. Reactions, levels, patterns, profiles and trends. Chemosphere 25(1-2):41-44.

Rappe, C.; Kjeller, L.O. (1987) PCDDs and PCDFs in environmental samples, air, particulates, sediments, and soil. Chemosphere 16(8/9): 1775-1780.

Rappe, C.; Nygren M.; Lindstrom G. (1987) Polychlorinated dibenzofurans and dibenzo-p-dioxins and other chlorinated contaminants in cow milk from various locations in Switzerland. Environ. Sci. Technol. 21(10): 964-970.

Rappe, C.; Bergqvist, P.A.; Kjeller, L.A. (1989a) Levels, trends and patterns of PCDDs and PCDFs in Scandinavian environmental samples. Chemosphere 18(1-6): 651-658.

Rappe, C.; Kjeller, L.O.; Andersson, R. (1989b) Analysis of PCDDs and PCDFs in sludge and water samples. Chemosphere 19(1-6): 13-20.

Rappe, C.; Bergquist, P.A.; Kjeller, L.O.; Swanson, S.; Belton, T.; Ruppel, B.; Lockwood, K.; Kahn, P.C. (1991) Levels and patterns of PCDD and PCDF contamination in fish, crabs, and lobsters from Newark Bay and the New York Bight. Chemosphere 22(3-4): 239-266.

Reed, L.W.; Hunt, G.T.; Maisel, B.E.; Hoyt, M.; Keefe, D.; Hackney, P. (1990) Baseline assessment of PCDDs/PCDFs in the vicinity of the Elk River, Minnesota generating station. Chemosphere 21(1-2): 159-171.

Rotard, W.; Christmann, W.; Knoth, W. (1993) Background levels of PCDD/F in soils of Germany. Unpublished.

Ryan, J.J.; Lizotte, R.; Sakuma, T.; Mori, B. (1985) Chlorinated dibenzo-p-dioxins, chlorinated dibenzofurans, and pentachlorophenol in Canadian chicken and pork samples. J. Agric. Food Chem. 33: 1021-1026.

Schecter, A. (1991) Dioxins and related compounds in humans and in the environment. In: Biological Basis for Risk Assessment of Dioxins and Related Compounds. Banbury Report #35. Edited by M. Gallo, R. Scheuplein and K. Van der Heijden. Cold Spring Harbor Laboratory Press. Plainview, NY.

Schecter, A.; Fürst, P.; Fürst, C.; Groebel, W.; Constable, J.D.; Kolesnikar, S.; Belm, A.; Boldonor, A.; Trubitsun, E.; Veasor, B.; Cau, H.D.; Dai, L.C.; Quynh, H.T. (1990) Levels of chlorinated dioxins, dibenzofurans and other chlorinated xenobiotics in food from the Soviet Union and the South of Vietnam. Chemosphere 20 (7-9): 799-806.

Schecter, A.; Papke, O.; Ball, M.; Startin, J.R.; Wright, C.; Kelly, M. (1992) Dioxin and dibenzofuran levels in food from the United States as compared to levels in food from other industrial countries. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Schecter, A.; Papke, O.; Ball, M.; Startin, J.R.; Wright, C.; Kelly, M. (1993) Dioxin levels in food from the U.S. with estimated daily intake. Submitted to Dioxin '93.

Sherman, R.K.; Clement, R.E.; Tashiro, C. (1990) The distribution of polychlorinated dibenzo-p-dioxins and dibenzofurans in Jackfish Bay, Lake Superior, in relation to a kraft pulp mill effluent. Chemosphere 20(10-12): 1641-1648.

Sievers, S.; Friesel, P. (1989) Soil contamination patterns of chlorinated organic compounds: looking for the source. Chemosphere 19(1-6): 691-698.

Smith, L.M.; Schwartz, T.R.; Feltz, K. (1990) Determination and occurrence of AHH-active polychlorinated biphenyls, 2,3,7,8-tetrachloro-p-dioxin and 2,3,7,8-tetrachlorodibenzofuran in Lake Michigan sediment and biota. The question of their relative toxicological significance. Chemosphere 21(9): 1063-1085.

Smith, R.M.; O'Keefe, P.W.; Hilker, D.R.; Aldous, K.M.; Mo, S.H.; Stelle, R.M. (1989) Ambient air and incinerator testing for chlorinated dibenzofurans and dioxins by low resolution mass spectrometry. Chemosphere 18: 585-592.

Smith, R.M.; O'Keefe, P.W.; Aldous, K.M.; Valente, H.; Connor, S.P.; Donnelly, R.J. (1990) Chlorinated dibenzofurans and dioxins in atmospheric samples from cities in New York. Environ. Sci. Technol. 24(10):1502-1506.

Smith, R.M.; O'Keefe, P.; Aldous, K.; Briggs, R.;