3. SOURCES 3-1

3.1. OVERVIEW OF SOURCES 3-1

3.2. PULP AND PAPER MILLS 3-14

3.2.1. Bleached Chemical Wood Pulp and Paper Mills 3-14

3.2.2. Nonchemical and Nonwood Pulping and Bleaching Mills 3-18

3.2.3. Ongoing Regulatory Investigations 3-19

3.3. PUBLICLY OWNED TREATMENT WORKS (POTWs) 3-20

3.3.1. Sources of CDDs/CDFs 3-20

3.3.2. Releases of CDDs/CDFs 3-22

3.4. CHEMICAL MANUFACTURING AND PROCESSING SOURCES 3-25

3.4.1. Manufacture of Halogenated Organic Chemicals - Overview 3-25

3.4.1.1. Chlorophenols 3-25

3.4.1.2. Chlorobenzenes 3-28

3.4.1.3. Chlorobiphenyls 3-33

3.4.1.4. Aliphatic Chlorine Compounds 3-33

3.4.1.5. Dyes and Pigments 3-35

3.4.2. Manufacture of Halogenated Organic Chemicals - Dioxin/Furan Test Rule Data 3-37

3.4.3. Manufacture of Halogenated Organic Chemicals-Pesticide Data Call-In 3-44

3.4.4. Chlorine Production Using Graphite Electrodes 3-56

3.4.5. Petroleum Refining Catalyst Regeneration 3-59

3.4.6. Additional Chemical Manufacturing and Processing Sources 3-62

3.5. MECHANISMS OF FORMATION OF DIOXIN-LIKE COMPOUNDS DURING COMBUSTION OF ORGANIC MATERIALS 3-63

3.5.1. CDD/CDF Contamination in Fuel as a Source of Combustion Stack Emissions 3-64

3.5.2. Formation of CDDs/CDFs from Precursor Compounds 3-67

3.5.3. The de novo Synthesis of CDDs/CDFs During Combustion of Organic Materials 3-75

3.5.4 Theory on the Emission of Polychlorinated Biphenyls 3-91

3.5.5. Evaluation of Naturally Occurring CDD/CDFs by Examination of Sediment Core Data 3-92

3.5.6. Summary of Theories of CDD/CDF Emissions 3-94

3.6. COMBUSTION AND OTHER HIGH TEMPERATURE SOURCES 3-96

3.6.1. Municipal Solid Waste Incineration 3-97

3.6.2. Hazardous Waste Incineration 3-109

3.6.3. Medical Waste Incineration 3-112

3.6.4. Kraft Black Liquor Recovery Boilers 3-119

3.6.5. Sewage Sludge Incineration 3-120

3.6.6. Primary Nonferrous Metal Smelting/Refining 3-122

3.6.7. Secondary Nonferrous Metal Smelting/Refining 3-123

3.6.7.1 Secondary Aluminum Smelters and Refiners 3-124

3.6.7.2 Secondary Copper Smelters and Refiners 3-124

3.6.7.3 Secondary Lead Smelters and Refiners 3-125

3.6.8. Primary Ferrous Metal Smelting/Refining 3-128

3.6.9. Secondary Ferrous Metal Smelting/Refining 3-129

3.6.10. Scrap Electric Wire Recovery 3-129

3.6.11. Drum and Barrel Reclamation and Incineration 3-131

3.6.12. Tire Combustion 3-132

3.6.13. Motor Vehicle Fuel Combustion 3-134

3.6.14. Wood Burning at Residences 3-143

3.6.15. Industrial Wood-Burning Facilities 3-146

3.6.16. Wood Burned in Forest Fires 3-147

3.6.17. Coal Combustion 3-151

3.6.18. Combustion of Polychlorinated Biphenyls (PCBs) 3-152

3.6.19. Pyrolysis of Brominated Flame Retardants 3-153

3.6.20. Carbon Reactivation Furnaces 3-154

3.6.21. Cement Kilns 3-156

3.6.22. Additional Combustion and High Temperature Sources 3-165

3.7. RESERVOIR SOURCES 3-166

3.8. COMPARING SOURCE EMISSIONS TO DEPOSITION ESTIMATES 3-167

The purpose of this chapter is twofold: (1) to identify sources that release dioxin-like compounds into the environment and (2) to derive national estimates for releases from these sources in the United States. The dioxin-like compounds have been found in all media and all parts of the world. This ubiquitous nature of these compounds suggests that multiple sources exist and that long range transport can occur. An unresolved issue is how the relative impacts from local vs. distant sources compare at a particular location. Presumably, in industrial areas, local sources will dominate, and in rural areas, distant sources will dominate. However, site specific considerations such as stack height, wind patterns, magnitude of local sources, etc. could influence these comparisons.

The chlorinated and brominated dioxins and furans have never been intentionally produced other than on a laboratory-scale basis for use in chemical analyses. Rather, they are generated as byproducts from various combustion and chemical processes. PCBs were produced in relatively large quantities for use in such commercial products as dielectrics, hydraulic fluids, plastics, and paints. They are no longer produced in industrialized countries, but continue to be released to the environment through the use and disposal of these products.

Dioxin-like compounds are released to the environment in a variety of ways and in varying quantities depending upon the source. For example,

• Releases to the air occur primarily from combustors and appear to have the most direct influence on human exposure. As discussed in Chapter 4, atmospheric deposition and subsequent accumulation through the food chain appears to be the major pathway of human exposure to dioxin-like compounds.

• Solid residues such as combustor ash, still bottoms, etc. can contain high levels of CDD/CDFs and can collectively contain more of these compounds than are found in air or water discharges. However, these solid residues are not generally released to the environment in an uncontrolled manner. Rather, they are usually disposed at secure landfills and any leaching to ground water is minimal due to their very low water solubility.

• Water discharges from paper mills, sewage treatment plants, and possibly other industries can contain low levels of CDD/CDFs. These releases can bioaccumulate via the aquatic food chain and ultimately lead to human exposure via fish ingestion.

The major identified sources of environmental release have been grouped into four major types for the purposes of this report:

• Industrial/Municipal Processes: Dioxin-like compounds can be formed through the chlorination of naturally occurring phenolic compounds such as those present in wood pulp. The formation of CDDs and CDFs resulting from the use of chlorine bleaching processes in the manufacture of bleached pulp and paper has resulted in the presence of CDDs and CDFs in paper products as well as in liquid and solid wastes from this industry. Municipal sewage sludge has been found to frequently contain CDDs and CDFs. Influents from industrial facilities, stormwater runoff, microbial metabolism of chlorophenols, and domestic household wastewater have been identified by various researchers as the sources(s) of the CDDs/CDFs.

• Chemical Manufacturing/Processing Sources: Dioxin-like compounds can be formed as by-products from the manufacture of chlorine and such chlorinated compounds as chlorinated phenols, PCBs, phenoxy herbicides, chlorinated benzenes, chlorinated aliphatic compounds, chlorinated catalysts, and halogenated diphenyl ethers. Although the manufacture of many chlorinated phenolic intermediates and products, as well as PCBs, was terminated in the late 1970s in the United States, continued, limited use and disposal of these compounds can result in releases of CDDs, CDFs, and PCBs to the environment. High levels of CDFs have been found in sludge from graphite electrodes used in chloralkali process to manufacture chlorine.

• Combustion and Incineration Sources: Dioxin-like compounds can be generated and released to the environment from various combustion processes when chlorine donor compounds are present. These sources can include incineration of wastes such as municipal solid waste, sewage sludge, hospital, and hazardous wastes; metallurgical processes such as high temperature steel production, smelting operations, and scrap metal recovery furnaces; and the burning of coal, wood, petroleum products, and used tires for power/energy generation.

• Reservoir Sources: The persistent and hydrophobic nature of these compounds causes them to accumulate in soils, sediments, and organic matter and to persist in waste disposal sites. The dioxin-like compounds in these "reservoirs" can be redistributed by various processes such as dust or sediment resuspension resulting in the potential for exposure. Releases from these "reservoirs" are not original sources in a global sense, but can be on a local scale. For example, past air emissions causing deposition onto a watershed with subsequent erosion may have resulted in accumulation in downstream sediments. Future sediment dredging operations could result in short-term significant resuspension of dioxins that had accumulated over a much longer period of time. Similarly, leaf composting operations could lead to releases of the dioxins that had, over the course of a growing season, deposited on or been sorbed to the leaves of deciduous trees in an area. Such leaf reservoirs could also be resuspended during forest fires.

As awareness of these possible sources has grown in recent years, a number of changes have occurred that should reduce the release rates (Rappe, 1992a). For example, releases of dioxin-like compounds have 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, new emission standards and upgraded emission controls for incinerators, and reductions in the manufacture of chlorinated phenolic intermediates and products.

Some investigators have raised the possibility that major sources exist that have not yet been identified. This suggestion is acknowledged to be quite speculative, but is important to consider. Three studies addressing this issue are summarized below.

Travis and Hattemer-Frey (1991) used the Fugacity Food Chain (FFC) model to predict the contribution of municipal solid waste incinerators, motor vehicles, hospital waste incinerators, residential wood burning, and pulp and paper mill effluents, to the U.S. total environmental input of 2,3,7,8-TCDD. It was estimated that the total input from all five sources combined accounted for only 11 percent of the total 2,3,7,8-TCDD found in the different media in the United States. The authors concluded that this low value indicated: (1) the source term used in the FFC modeling exercise for 2,3,7,8-TCDD may have been too high; (2) some unidentified major source(s) of 2,3,7,8-TCDD exist; or (3) multiple environmental sources of 2,3,7,8-TCDD with no one source dominating the total input.

Rappe (1991) found discrepancies between estimated emissions from known sources of CDDs and CDFs into the Swedish environment and calculated aerial deposition rates. The total emissions in Sweden were estimated by Rappe (1991) as 100 to 250 g TEQ/yr. The deposition rates used by Rappe (1991) were derived from a study by Marklund (1990) who made measurements in rural areas of Sweden and found an average deposition rate of 5 ng of TEQ/m² - yr. [Later measurements by Andersson et al. (1992) indicate that deposition in this area has been reduced to about 1 ng of TEQ/m²-yr due to emissions reductions.] Rappe (1991) multiplied the deposition rate of 5 ng TEQ/m²-yr by the total land area of Sweden, yielding a total annual deposition for Sweden of 2,250 g of TEQ/yr, which appears to be 10 to 20 times higher (or 2 to 4 times higher using the deposition rate of Andersson et al. [1992]) than the total emissions from sources originating in Sweden. Possible explanations for this discrepancy are (1) uncertainty in the emission estimates, (2) uncertainty in the deposition estimates, (3) long-range transport of dioxin-like compounds from sources outside of Sweden, or 4) existence of unidentified sources. In an earlier publication, Rappe et al. (1987) compared congener patterns found in human and aquatic life tissue samples with the congener patterns found in various known emission sources and contaminated products. A poor correlation was observed between the congener patterns found in human and environmental samples and the respective potential sources. Rappe et al. (1987) speculated that the observed pattern in human and environmental samples could be the result of a combination of sources, coupled with environmental and biological degradation of the released congeners.

Harrad et al. (1992a; 1992b) have made similar estimates for the United Kingdom. They estimate that the average annual deposition from the atmosphere to the land surface in the United Kingdom is 250 kg of CDD/CDF (on a total mass basis, not TEQ), compared to about 29.1 kg/yr emitted from known sources. As with the other two studies, these discrepancies could be the result of inaccuracies in emission/deposition estimates, long- range transport from outside the country, or unidentified sources. The authors speculated that much of the discrepancy may be accounted for by secondary or "reservoir" sources (i.e. the remobilization and subsequent redeposition of CDD/CDFs already in the environment).

Table 3-1 presents CDD and CDF source-specific air emission estimates reported for West Germany (Fiedler and Hutzinger, 1992), Austria (Riss and Aichinger, 1993), the United Kingdom (ECETOC, 1992), the Netherlands (Koning et al., 1993), Switzerland (Schatowitz et al., 1993), and the United States (based on estimates generated in this document). The emission estimates for West Germany and Switzerland suggest that municipal waste incinerators and metal smelters/refiners are the largest sources of air emissions. In Austria, domestic combustion of wood is believed to be the largest source followed by emissions from the metallurgical industry. In the United Kingdom, municipal waste incinerators and coal combustion are estimated to be the major sources. Municipal waste incinerators are also estimated to be the largest source in the Netherlands. Rappe (1992a) and Lexen et al. (1992) have identified emissions from ferrous and nonferrous metals smelting and refining facilities as potentially the largest current source in Sweden. Rappe (1992a) reported that changes in various industrial practices have led to reductions in dioxin emissions in Sweden from 400 - 600 g of TEQ/yr in 1985 to 100 - 200 g TEQ/yr in 1991.

Similar nationwide emission estimates for the United States have not previously been compiled. This task has been attempted in this document, and the results are presented in Table 3-1 (air emissions only) and in Table 3-2. Table 3-2 lists emission estimates for the major known or suspected sources that could have releases of dioxin-like compounds to the environment. For each source listed in Table 3-2, estimated emissions to air, water, land, and product are listed where appropriate and where data are adequate to enable an estimate to be made. The term "product" in Table 3-2 is defined to include substances or articles (e.g., paper pulp or sewage sludge that is distributed/marketed commercially) that are known to contain dioxin-like compounds and whose subsequent use may result in releases to the environment. Figure 3-1 is a chart that visually displays the range of emission estimates to air that are reported in Table 3-2.

In order to make each source emission estimate, information was required concerning both the "emission factor" term for the source (e.g., grams TEQ per kg of material processed) and the "production" term for the source (e.g., kg of material processed annually in the United States). Because the quantity and quality of the available information for both terms for each emission source varies considerably, a confidence rating scheme was developed. This scheme is based on a consideration of the following factors:

Basis of Estimate - The basis for the emission estimate varied widely from expert judgement to detailed studies. The best studies involved direct emission measurements at multiple facilities. The representativeness of emission samples was evaluated on the basis of the variability in technologies and associated release rates among individual facilities in the source category. The more variability among facilities; the more important it is to test multiple facilities. In other cases, although no direct emission measurements were available, estimates could be derived using indirect techniques. Obviously, these "indirect" estimates are much more uncertain than those based on direct measurements.

Citation Quality - The quality of the supporting literature varied widely. Whenever possible, only peer reviewed final reports were used. In some cases, however,

draft reports that had undergone some review were used. In a few cases, unpublished references were used such as personal communication with experts.

The confidence rating scheme, presented in Table 3-3, provides criteria for assigning a "high," "medium," or "low" confidence rating for both the emission factor and production terms. As shown in Table 3-2, confidence ratings have been assigned to each emission estimate. The first rating applies to the "production" term, and the second rating applies to the "emission factor" term. In addition to the confidence rating, the uncertainty in these national release estimates is reflected by presenting, where possible, for each source category both a central or "best guess" value and a possible range from a lower to upper estimate. These lower and upper estimates are not intended to be absolute bounds, but reasonable estimates of how much higher or lower the true value might be. Insufficient data were available to statistically derive these ranges. Therefore, a judgement-based approach was developed. This approach uses the average or best guess estimate as the central value of the range (assumed to be a geometric average) and sets the width of the range on the basis of the confidence class as follows:

· Low confidence class: upper end of range is 10 times higher than lower end;

· Medium confidence class: upper end of range is 5 times higher than lower end;

· High confidence class: upper end of range is 2 times higher than lower end.

This approach initially assumes that the range of uncertainty is symmetrical about the central value. However, in some cases it may be more reasonable to shift the uncertainty range upwards or downwards. For example, it may be reasonable to shift the range downwards in cases where there is strong evidence that upgrades have occurred since the emissions testing. Alternatively, it is possible that the range should be shifted upwards if it can be shown that the tested facilities are more representative of the low emitting facilities than the high emitting facilities. It is emphasized that these ranges should be interpreted as judgements which are symbolic of the relative uncertainty among sources, not statistical measures. The remainder of this chapter reviews the available data for estimating CDD/CDF releases from specific source categories and provides the basis for the emission estimates presented in Table 3-2 for the United States.

During 1988, EPA and the U.S. pulp and paper industry jointly conducted a survey of 104 pulp and paper mills in the United States to measure levels of dioxins in effluent, sludge, and pulp (U.S. EPA, 1990a). This study, commonly called the 104-Mill Study, was managed by the National Council of the Paper Industry for Air and Stream Improvement, Inc. (NCASI) with oversight by EPA, and included all U.S. mills where chemically produced wood pulps are bleached with chlorine or chlorine derivatives.

In 1992, the pulp and paper industry conducted its own NCASI-coordinated survey. The collected data were summarized and analyzed in a report entitled "Summary of Data Reflective of the Pulp and Paper Industry Progress in Reducing the TCDD/TCDF Content of Effluents, Pulps, and Wastewater Treatment Sludges" (NCASI, 1993). Although the report is available from NCASI, it has not been peer reviewed nor published in an independent journal. The data used in the report were provided by individual pulp and paper companies and neither NCASI nor EPA can vouch for the accuracy or representativeness of the data. However, NCASI (1993) reports that the pulp and paper industry has taken numerous steps to reduce CDD/CDF releases since 1988, and that NCASI considers the 1992 survey to be more reflective of current conditions than the data generated in the 104-Mill Study (U.S. EPA, 1990a).

As part of its ongoing efforts to develop revised effluent guidelines and standards for the pulp, paper, and paperboard industry, EPA recently published the Development Document for the guidelines and standards being proposed for this industry (U.S. EPA, 1993d). The Development Document presents estimates of the 2,3,7,8-TCDD and 2,3,7,8-TCDF annual discharges in wastewater from the mills in this industry as of January 1, 1993. EPA used the most recent information about each mill from four data bases (104-Mill Study, EPA short-term monitoring studies at 13 mills, EPA long-term monitoring studies at 8 mills, and industry self-monitoring data submitted to EPA) to estimate these discharges. The 104-Mill Study data were used only for those mills that did not report making any process changes subsequent to the 104-Mill Study and did not submit any more recent effluent monitoring data. For the purpose of this report, the release estimates from NCASI (1993) and U.S. EPA (1990a) are presented to show the possible range of releases within recent years, but the U.S. EPA (1993d) estimates are believed to be most reflective of current conditions.

NCASI (1993) found that less than 10 percent of mills had 2,3,7,8-TCDD and 2,3,7,8-TCDF concentrations in effluent above the nominal detection limits of 10 ppq and 100 ppq, respectively. Similar results were obtained in the short- and long-term sampling reported for 18 mills in U.S. EPA (1993d). 2,3,7,8-TCDD was detected at four mills, and 2,3,7,8-TCDF was detected at nine mills. Wastewater sludges at most mills (75 to 90 percent) were reported by NCASI (1993) to contain less than 10 ppt of 2,3,7,8-TCDD and less than 100 ppt of 2,3,7,8-TCDF. U.S. EPA (1993d) reported similar results but did find detectable levels of 2,3,7,8-TCDD and 2,3,7,8-TCDF in sludges from 64 percent and 85 percent of the facilities sampled, respectively. NCASI (1993) reported that nearly 90 percent of the bleached pulps contained less than 2 ppt of 2,3,7,8-TCDD and less than 20 ppt of 2,3,7,8-TCDF. The final levels in white paper products would correspond to levels in bleached pulp, so bleached paper products would also be expected to contain less than 2 ppt of 2,3,7,8-TCDD. Overall, NCASI (1993) reports a 90 percent reduction in TEQ generation from 1988 to 1992.

The 104-Mill Study and the NCASI study measured only 2,3,7,8-TCDD and 2,3,7,8-TCDF because these two congeners are the primary contributors (90 percent or more) to the TEQ total found in pulp, sludge, and effluent (U.S. EPA, 1990b). Ninety-four mills participated in the NCASI study, and the remaining 10 (of 104) were assumed by NCASI to be operating at the same levels as measured in the 1988 104 Mill Study. All not detected values were counted as half the detection limit. If detection limits were not reported, they were assumed to be 10 ppq for effluent and 1 ppt for sludge or bleached pulp.

The U.S. annual discharge rates of 2,3,7,8-TCDD, 2,3,7,8-TCDF, and TEQs due to these two compounds are summarized in Table 3-4 for each study. As stated previously, the 1993 discharge estimate for effluent (U.S. EPA, 1993d) is believed to be the best estimate of current emissions. During the period between the conduct of the 104 Mill Study and the issuance of the U.S. EPA Development Document (U.S. EPA, 1993d), the U.S. pulp and paper industry has reduced releases of CDD/CDFs primarily by instituting numerous process changes to reduce the formation of CDD/CDFs during the production of chemically bleached wood pulp. U.S. EPA (1993d) did not provide extensive sampling of sludge and pulp samples from bleached chemical wood pulp and paper mills comparable to that provided for effluents. However, because most of the reduction between 1988 and 1993 can be attributed to process changes of a pollution prevention nature, the percentage reduction observed in effluent emissions (from 356 g TEQ/yr to 105 g TEQ/yr or 70 percent reduction) is likely representative of the reduction that has been achieved in sludge and pulp emissions over this same time period. Table 3-4 presents best estimates of emissions in sludge and pulp of 100 g TEQ/yr and 150 g TEQ/yr, respectively, using this assumption. The confidence ratings for these release estimates were judged to be H/H based on the fact that direct measurements have been made at virtually all facilities, indicating a high level of confidence in both the production and emission factor estimates. Based on these high confidence ratings, the estimated ranges of potential annual emissions for effluent, sludge, and pulp are assumed to vary by a factor of 2 between the low and high ends of the ranges. Assuming that the best estimates of annual emissions (i.e., the 1993 discharge-based estimates presented in Table 3-4) are the geometric means of these ranges, then the ranges are calculated to be 74 to 150 g TEQ/yr for effluent, 71 to 140 g TEQ/yr for sludge, and 105 to 210 g TEQ/yr for pulp.

In 1990, approximately 20.5 percent or 500 million dry kg of the pulp and paper mill wastewater sludge generated by facilities employing chlorine bleaching of pulp were incinerated (U.S. EPA, 1993e). The majority of the wastewater sludge generated by these facilities is landfilled or placed in surface impoundments (79.5 percent) with the remainder incinerated (20.5 percent), applied to land directly or as compost (4 percent), or distributed as a commercial product (less than 1 percent) (U.S. EPA, 1993e). Black liquor recovery boilers used in the Kraft process for the production of paper pulp are potential sources of

CDDs/CDFs. Estimates of potential CDD/CDF emissions to air from these sources are discussed in Section 3.6.4.

Although the EPA Office of Water does not believe that secondary fiber mills (i.e., mills using recycled paper as a source of pulp) are significant sources of CDDs and CDFs, EPA is considering whether to establish effluent limitations guidelines and standards for CDD/CDFs for these mills based primarily upon data generated for the Development Document (U.S. EPA, 1993d). These data, collected by EPA or provided to EPA by industry, indicate detectable levels of 2,3,7,8-TCDD in the effluents of 2 of the 12 mills with reported monitoring data and detectable levels of 2,3,7,8-TCDF in the effluents of 4 of the 7 mills with data (U.S. EPA, 1993d).

Data on the presence of more chlorinated (i.e., penta-through octachlorinated) CDDs and CDFs in the effluents of these facilities were not generated for the Development Document (U.S. EPA, 1993d). However, Berry et al. (1993) reports that trace levels of these higher chlorinated homologs were commonly observed in the effluents from Canadian pulp mills that use recycled paper for fiber furnish (i.e., the raw materials used to manufacture pulp) and/or that do not practice chlorine bleaching. Similar results were reported by Rappe et al. (1990). The congener profile observed is not dominated by the tetra-CDDs/CDFs, as is the case with bleach plant wastewater, but rather by the higher chlorinated congeners more consistent with the congener profile found in ambient air, soil, and adipose tissue. These results lead to the hypothesis that paper and paperboard products, during their useful life, can accumulate trace amounts of CDD/CDFs from the ambient environment.

As a step in evaluating this hypothesis, Berry et al. (1993) analyzed the CDD/CDF content of pulp and paper samples from Canadian mills that use neither chlorine-containing bleaching compounds nor fibers that have been bleached with chlorine-containing compounds as well as papers from mills that use recycled paper as a furnish. All samples analyzed had detectable levels of one or more CDD/CDFs. The congener profiles of the samples were similar with the higher chlorinated congeners dominating in terms of concentration. The order of degree of contamination on a TEQ basis, from high to low, is recycled linerboard (1 sample--2.5 ng TEQ/kg) > "totally chlorine-free" bleached kraft paper (1 sample--0.35 ng TEQ/kg) > pulp from de-inked recycled paper (1 sample--0.19 ng TEQ/kg) > newsprint (17 samples--mean = 0.07 ng TEQ/kg) > unbleached kraft paper (2 samples--mean = 0.02 ng TEQ/kg). Rappe et al. (1990) also reported finding higher levels of CDD/CDFs, particularly the hepta- and octa-chlorinated congeners, in recycled paper pulps than in virgin bleached and unbleached pulps. Based on the results of their study, Berry et al. (1993) concluded that, although it may be possible to produce a dioxin-free pulp, it is likely that all papers will become contaminated during their first life cycle by contact with dioxin-laden dust, and contamination is inevitable if they are recycled multiple times.

The U.S. EPA is currently under court order to develop revised effluent guidelines (i.e., Best Available Technology and Pretreatment Standards for Existing Sources) for the chemical pulping and bleaching subcategories of the pulp and paper industry. These revised effluent guidelines and standards which address control of CDDs and CDFs from bleached chemical wood pulp and paper mills were proposed by EPA on December 17, 1993 (Federal Register, 1993a). In addition, the Clean Air Act Amendments of 1990 require EPA to promulgate Most Achievable Control Technology (MACT) standards for hazardous air pollutants from this industry by 1997. To that end, the Office of Air and Radiation, in coordination with the Office of Water, proposed control technology standards for non-combustion sources on December 17, 1993, (Federal Register, 1993a) and will propose control technology standards for combustion sources by October 1994 with promulgation of both by September 1995 (U.S. EPA, 1992d).

Based on the results of an in-depth risk assessment, EPA's Office of Solid Waste concluded that dioxin contained in pulp and paper mill sludges does not pose an unreasonable probability of adverse effects on human health and the environment when disposed in landfills and surface impoundments and that further regulation of these facilities under Subtitle D of the Resource Conservation and Recovery Act (RCRA) to reduce potential dioxin-related risks was not warranted (U.S. EPA, 1991a).

However, EPA did find that improper land application of pulp and paper mill sludge for soil conditioning purposes can pose a significant risk to wildlife. In 1991, EPA proposed a regulation under the Toxic Substances Control Act (TSCA) to limit the concentration of CDDs/CDFs in soil conditioned with sludge and also to establish site management practices for land application of the sludge. EPA deferred finalizing the rule until issuance of the final integrated regulations for effluent guidelines and MACT standards. These regulations could make TSCA rulemaking unnecessary. In the interim, EPA is negotiating a voluntary agreement with the American Forest and Paper Association to establish CDD/CDF standards and management practices for the use of sludge as a conditioner (U.S. EPA, 1993b).

CDD/CDFs have been measured in sewage sludge, though the origins have not been well established. In fact, Oberg et al. (1992) reported that low levels of HpCDDs and OCDD are formed, probably as a result of microbial action, in aerated sewage sludge spiked with pentachlorophenol. Potential sources of the CDD/CDFs include industrial inputs, runoff to sewers from lands or urban surfaces contaminated by product uses or deposition of emissions from combustion sources, household wastewater, chlorination operations within the wastewater treatment facility, or a combination of all the above (Rappe, 1992a; Rappe et al., 1989; Horstmann et al., 1992). The major source(s) for a given treatment plant is likely to be site-specific. For example, Rieger and Ballschmiter (1992) traced the origin of CDDs and CDFs found in municipal sewage sludge in Ulm, Germany, to metal manufacturing and urban sources. The characteristics of both sources were similar and suggested generation via thermal processing. The presence of CDD/CDFs in sewage sludge suggests that CDD/CDFs may also be present in the wastewater effluent discharges of POTWs; however, no published studies reporting the results of effluent analyses for CDD/CDFs could be found.

In a series of recent studies, Horstmann et al. (1992; 1993a; 1993b) and Horstmann and McLachlan (1994) demonstrated that wastewater from household washing machines could be the major source at many, if not all, POTWs that serve primarily residential populations. Horstmann et al. (1992) provided initial evidence that household wastewater could be a significant source. Horstmann et al. (1993a) measured CDD/CDF levels in the effluent from four different loads of laundry from two different domestic washing machines. The concentrations of total CDD/CDF in the four samples ranged from 3,900 to 7,100 pg/L and were very similar in congener profile with OCDD being the dominant congener followed by the hepta- and hexa-CDDs. Based on the similar concentrations and congener profiles found, Horstmann et al. (1993a) concluded that the presence of CDD/CDF in washing machine wastewater is widespread. A simple mass balance performed using the results showed that the CDD/CDFs found in the four washing machine wastewater samples could account for 27 to 94 percent of the total CDD/CDF measured in the sludge of the local wastewater treatment plant (Horstmann and McLachlan, 1994).

Horstmann et al. (1993a) also performed additional experiments that showed that detergents, commonly used bleaching agents, and the washing cycle process itself were not responsible for the observed CDD/CDFs. Rappe and Andersson (1992) had previously reported that wastewater from clothing and dish washing machines in which sodium hypochlorite-containing detergents were used contained low levels of CDD/CDFs.

To determine if the textile fabric or fabric finishing processes could account for the observed CDD/CDFs, Horstmann et al. (1993b) analyzed the CDD/CDF content of eight different raw (unfinished) cotton cloths containing fiber from different countries and five different white synthetic materials (acetate, viscose, bleached polyester, polyamide, and polyacrylic). The maximum concentrations found in the textile fabrics were 30 ng/kg in the cotton products and 45 ng/kg in the synthetic materials. Also, a cotton finishing scheme was developed in which one of the cotton materials was subjected to a series of 16 typical cotton finishing processes; one sample was analyzed following each step. The fabric finishing processes showing the greatest effect on CDD/CDF concentration were the application of an indanthrene dye and the "wash and wear" finishing process which together resulted in a CDD/CDF concentration of about 100 ng/kg. Based on the concentrations found, the authors concluded that neither unfinished new fabrics nor common cotton finishing processes can explain the CDD/CDF levels found in wastewater.

Horstmann and McLachlan (1994) analyzed 35 new textile samples, primarily cotton products, for CDD/CDFs. Low levels were found in many cases (total CDD/CDF less than 50 ng/kg). However, several colored T-shirts from a number of clothing producers had extremely high levels, with concentrations up to 290,000 ng/kg. Because the concentrations in identical T-shirts purchased at the same store varied by up to a factor of 20, the authors concluded that the source of CDD/CDFs is not a textile finishing process because a process source would have resulted in a more consistent level of contamination.

Horstmann and McLachlan (1994) conducted additional experiments that demonstrated that the small percentage of clothing items with high CDD/CDF levels could be responsible for the quantity of CDD/CDFs observed in household wastewater and sewage sludge. They were able to demonstrate that the CDD/CDFs can be gradually removed from the fabric during washing, can be transferred to the skin and subsequently transferred back to other textiles and then washed out, or can be transferred to other textiles during washing and then removed during subsequent washings.

EPA conducted the National Sewage Sludge Survey in 1988 to obtain national data on sewage sludge quality and management. As part of this survey, EPA analyzed sludges from 175 POTWs for CDD/CDF content; sludges from 15 of the POTWs had detectable levels of 2,3,7,8-TCDD. All sludges had detectable levels of at least one CDD/CDF congener (Rubin and White, 1992). TEQ concentrations ranged from 0.7 to 1,816 ng TEQ/kg dry weight. If all not detected values are assumed to be zero, then the mean and median concentrations are 50 and 9 ng TEQ/kg, respectively. If the not detected values are set equal to the detection limit, then the mean and median concentrations are 86 and 50 ng TEQ/kg, respectively (Rubin and White, 1992).

Approximately 5.4 million dry metric tons of sewage sludge are estimated by EPA to be generated annually in the United States (Federal Register, 1993b). Table 3-5 lists the volume of sludge disposed annually by use and disposal practices. Table 3-5 also lists the estimated amount of TEQs that may be present in sewage sludge and potentially be released to the environment. These values were estimated using the mean TEQ concentration value (not detected values assumed to be zero) reported by Rubin and White (1992) (i.e., 50 ng TEQ/kg). Multiplying this mean concentration by the sludge volumes generated, yields an annual potential total release of 208 grams of TEQ for nonincinerated sludges. Of this 208 grams of TEQ, 3.6 grams enter commerce as a product for distribution and marketing. The remainder is applied to land or is landfilled.

This release estimate is assigned a H/H confidence rating indicating high confidence in both the production and emission factor estimates. The high rating was based on the judgement that the 175 tested facilities were reasonably representative of the variability in the POTW technologies and sewage characteristics. Based on this high confidence rating, the estimated range of potential annual emissions is assumed to vary by a factor of 2 between the low and high ends of the range. Assuming that the best estimate of annual emission to land (105 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 145 to 290 g TEQ/yr. Assuming that the best estimate of 3.6 g TEQ annual emissions in product (i.e., the fraction of sludge that is distributed and marketed as a product) is the geometric mean of the range, then the range is calculated to be 2.5 to 5.0 g TEQ/yr.

An additional 10 to 52 grams of TEQ (central estimate of 23 g TEQ/yr) are estimated to be released to the atmosphere annually by the incineration of sewage sludge. The basis of these incineration release estimates is presented in Section 3.6.5. It is interesting to note that CDDs and CDFs detected in ambient air in Ohio have been linked to sewage sludge combustion (Edgerton et al., 1989). In this study, total CDD/CDF in ambient air ranged from 1,900 to 9,900 fg/m³; no 2,3,7,8-TCDD was detected in any of the samples with a detection limit of less than 240 fg/m³.

Several chemical production processes have been shown to generate CDDs and CDFs (Versar, 1985; Hutzinger and Fiedler, 1991a). CDDs and CDFs can be formed during the manufacture of chlorophenols, chlorobenzenes, and chlorobiphenyls (Versar, 1985; Ree et al., 1988). Consequently, disposal of industrial wastes from manufacturing facilities producing these compounds may result in the release of CDDs and CDFs to the environment. Also, the products themselves may contain these compounds, and when used/consumed, may result in additional releases to the environment. CDD and CDF congener distribution patterns indicative of noncombustion sources have been observed in sediments in southwest Germany and the Netherlands. The congener patterns found suggest that wastes from the production of chlorinated organic compounds may be important sources of CDD and CDF contamination in these regions (Ree et al., 1988). The production and use of many of the chlorophenols, chlorophenoxy herbicides, and PCB products have been banned or strictly regulated in most countries. However, these products may have been a source of the environmental contamination that occurred prior to the 1970s and may continue to be a source of environmental releases based on limited use and disposal conditions (Rappe, 1992a).

The two major manufacturing processes used to produce chlorophenols include: (1) electrophilic chlorination of phenol by chlorine gas in the presence of catalytic amounts of aluminum chloride and organic chlorination promoters and stabilizers; and (2) alkaline hydrolysis of chlorobenzenes using aqueous methanolic sodium hydroxide and heat (Ree et al., 1988). CDD and CDF formation is promoted by the high temperatures and/or alkaline conditions used in these processes. CDDs and CDFs may be formed by nucleophilic substitution, radical reactions, and pyrolysis mechanisms (Versar, 1985; Ree et al., 1988). The major CDD/CDF congeners generated by chlorophenol manufacture are the hexa- through octa-chlorinated congeners (Versar, 1985).

The concentrations of CDD/CDFs in chlorophenols analyzed in the 1970s and early 1980s were assembled and summarized by Versar (1985) and Hutzinger and Fiedler (1991a). Hagenmaier and Brunner (1987) reported the results of analyses of four pentachlorophenol products commercially available during the late 1980s; the total TEQ concentrations in these four products ranged from 0.08 to 2.32 mg/kg. Table 3-6 presents a summary of the data from these three studies. No more recent data on concentrations of CDDs and CDFs in chlorophenols could be found in the literature. However, the mono- through tetra- substituted chlorophenols and bromophenols are subject to reporting under the Dioxin/Furan Test Rule (discussed in Section 3.4.2) and/or the Dioxin/Furan Pesticide Data Call-In. (See Section 3.4.3.) CDDs and CDFs have also been found in numerous chlorophenol-based biocides according to Versar (1985) and Hutzinger and Fiedler (1991a). (See Section 3.4.3 for information on current EPA efforts to obtain data on contamination levels in pesticides.)

Several studies have provided evidence of localized environmental contamination resulting from the production or use of chlorophenols. For example, Tong et al. (1990) observed that sediment samples collected from a site near a chemical manufacturing facility where 2,4,5-T had been synthesized were highly contaminated with CDDs and CDFs. In addition, the CDD and CDF congener distribution pattern in the sediment was similar to that of 2,4,5-T, suggesting the manufacture found in 2,4,5-T as a primary source of contamination.

As indicated in Table 3-6, pentachlorophenol (PCP) products have been reported to be the most contaminated chlorophenol products. The major congener found in PCP is OCDD, but lower chlorinated congeners are also found (Rappe et al., 1987; Hutzinger and Fiedler, 1991a). High levels of CDD/CDFs have also been found in sludges from the production of PCP (Versar, 1985; Hutzinger and Fiedler, 1991a). McKee et al. (1990) surveyed harbor sediments adjacent to a wood preserving plant in Ontario, Canada, that uses PCP and creosote. Sediments were contaminated with hexa-, hepta-, and octa-chlorinated CDD/CDFs. The highest levels observed were: 5.7 ng/g HxCDD, 320 ng/g HpCDD, 980 ng/g OCDD, 6.5 ng/g HxCDF, and 53 ng/g HpCDF for a site 13 meters from the facility's dock and 400 ng/g OCDF for a site 78 meters from the dock. CDD/CDFs have also been found in composts from a yard waste composting facility in the United Kingdom (Harrad et al., 1991). Past use of PCP-based biocides was suggested as the major source of contamination, based on isomer patterns and empirical evidence.

In the mid-1980s, EPA's Office of Solid Waste promulgated land disposal restrictions on wastes (i.e., wastewaters and non-wastewaters) resulting from the manufacture of chlorophenols (40 CFR 268). Table 3-7 lists all solid wastes in which CDDs and CDFs are regulated as hazardous constituents by EPA, including chlorophenol wastes. The regulations prohibit the land disposal of these wastes until they have been treated to a level below the routinely achievable detection limit of 1 ppb in the waste extract for each of the following congener groups: TCDDs, PeCDDs, HxCDDs, TCDFs, PeCDFs, and HxCDFs (standards for waste code F039 apply only to TCDDs and TCDFs). The treatment standard of 1 ppb is based on incineration to 99.9999 percent destruction

and removal efficiency. Section 3.4.3 of this report describes regulatory actions taken by EPA to control the manufacture and use of chlorophenol-based pesticides.

EPA's Office of Water has promulgated effluent limitations for facilities that manufacture chlorinated phenols and discharge treated wastewater (40 CFR 414.70). Although these effluent limitations do not specifically address CDDs and CDFs, the treatment processes required to control the chlorinated phenols that are regulated (2-chlorophenol and 2,4,-dichlorophenol) are expected to control releases of CDDs and CDFs to minimal levels. The effluent limitations for the individual regulated chlorinated phenols are less than or equal to 39 µg/l for facilities that utilize biological end-of-pipe treatment.

Chlorobenzenes are manufactured by electrophilic substitution reactions of gaseous chlorine and benzene (Ree et al., 1988). CDD/CDFs may form during the production of these chemicals, but with less probability than in chlorophenol manufacturing (Hutzinger and Fiedler, 1991a). CDD/CDFs form by nucleophilic substitution and pyrolysis mechanisms (Ree et al., 1988). The factors contributing to the production of CDD/CDFs are: (1) using oxygen as a nuclear substituent; (2) producing or purifying the substance under alkaline conditions; and (3) using reaction temperatures above 150° C (Hutzinger and Fiedler, 1991a).

The concentrations of CDD/CDFs found in single samples of chlorobenzenes by researchers in Germany (Hagenmaier and Brunner, 1987; Hutzinger and Fiedler, 1991a) are listed in Table 3-8. In di-, tri-, tetra-, and penta-chlorobenzene, CDD/CDFs have been detected in the sub-µg/kg range. In hexachlorobenzene, CDD/CDFs have been detected in the µg-mg/kg range. No more recent data on concentrations of CDDs and CDFs in chlorobenzenes could be found in the literature. The limited available published information on CDD/CDF concentrations in chlorobenzene products is not sufficient in quantity (i.e., number of samples) or in detail (i.e., congener-specific results) to enable a reliable estimate to be made of the mass of CDDs/CDFs present in chlorobenzene products even though reliable annual production volume information is available for some products (e.g., 107,526 metric tons of monochlorobenzene and 63,104 metric tons of dichlorobenzene were produced in the United States in 1990) (U.S. ITC, 1991). However, the mono-, di-, and trichlorobenzenes are subject to reporting under the Dioxin/Furan Test rule (Section 3.4.2) and/or the Dioxin/Furan Pesticide Data Call-In (Section 3.4.3).

EPA's Office of Solid Waste has promulgated land disposal restrictions on wastes (i.e., wastewaters and non-wastewaters) resulting from the manufacture of chlorobenzenes (40 CFR 268). Table 3-7 lists all solid wastes in which CDDs and CDFs are regulated as hazardous constituents by EPA, including chlorobenzene wastes. The regulations prohibit the land disposal of these wastes until they have been treated to a level below the routinely achievable detection limit of 1 ppb in the waste extract for each of the following congener groups: TCDDs, PeCDDs, HxCDDs, TCDFs, PeCDFs, and HxCDFs (standards for waste code F039 apply only to TCDDs and TCDFs). The treatment standard of 1 ppb is based on incineration to 99.9999 percent destruction and removal efficiency.

EPA's Office of Water has promulgated effluent limitations for facilities that manufacture chlorinated benzenes and discharge treated wastewater (40 CFR 414.70). Although these effluent limitations do not specifically address CDDs and CDFs, the treatment processes required to control the chlorinated benzenes that are regulated (chlorobenzene; 1,2-dichlorobenzene; 1,3-dichlorobenzene; 1,4-dichlorobenzene; 1,2,4-trichlorobenzene; and hexachlorobenzene) are expected to control releases of CDDs and CDFs to minimal levels. The effluent limitations for the individual regulated chlorinated benzenes are less than or equal to 77 µg/l for facilities that utilize biological end-of-pipe treatment and are less than or equal to 196 µg/l for facilities that do not employ biological end-of-pipe treatment.

PCBs are manufactured by the direct chlorination of biphenyl in the presence of a catalyst. HpCDDs, OCDD, and CDFs, particularly the tetra-, penta-, and hexa-chlorinated CDF congeners, have been detected in commercial PCB formulations (Hagenmaier, 1987) However, the production of PCBs in the United States has been banned under TSCA and the use of in-service PCBs has been dramatically reduced. CDFs can be formed from PCBs under pyrolytic conditions, or by nonpyrolytic conditions via chlorine substitutions on the ortho-positions in the PCB molecule (Ree et al., 1988). Combustion of PCB-containing materials in transformers and capacitors may be a source of PCB-associated CDFs. (See Section 3.5.17.)

Aliphatic chlorine compounds are used as monomers in the production of plastics, as solvents and cleaning agents, and as precursors for chemical synthesis (Hutzinger and Fiedler, 1991a). These compounds are produced in large quantities. In 1990, 13.2 million metric tons of chlorinated aliphatic hydrocarbons were produced (U.S. ITC, 1991). The production of 1,2-dichloroethane and vinyl chloride accounted for 85 percent of this total production. Highly chlorinated CDDs and CDFs (i.e., hexa- to octa-chlorinated congeners) have been found in samples of 1,2-dichloroethane (55 ppb of OCDF), tetrachloroethane (47 ppb of OCDD), and epichlorohydrin (88 ppb of CDDs and 33 ppb of CDFs) (Hutzinger and Fiedler, 1991a). Because no more recent or additional data could be found in the literature to confirm these values, no estimates have been made of the mass of CDDs/CDFs present in these products manufactured annually.

Greenpeace recently issued a report (Greenpeace, 1993) on dioxin emissions associated with the production of ethylene dichloride (EDC) and vinyl chloride monomer (VCM). The Vinyl Institute has responded with a critique of the report (ChemRisk, 1993). Both of these studies are discussed below.

Greenpeace (1993) estimated that plants producing EDC and VCM release 1.8 kg of TEQ/yr to the environment (air, water, and ground combined - possible releases in the final products were not discussed). This estimate was based on an emission factor of 5 to 10 g TEQ/100,000 tons of VCM produced and a worldwide estimate of PVC (and thus VCM) production of 18 million metric tons/yr. This estimate represents the total emissions from all plants in the world but was based on data from only four European plants. Greenpeace (1993) cited some specific information on CDD/CDF formation or releases from a lengthy list of primary references. While most of the specific data came from studies conducted or sponsored by industry, in no case was the information offered by Greenpeace (1993) complete enough to allow calculation of all process or waste stream-specific emission factors to particular environmental media for a given plant.

European PVC manufacturers claim the emission factor is 0.01 to 0.5 g TEQ/100,000 metric tons of VCM, resulting in global emissions from EDC/VCM production as 0.002 to 0.09 kg TEQ/yr (Miller, 1993). There is no apparent dispute between the industry and Greenpeace regarding the formation of CDDs/CDFs during the production process, nor that some CDDs/CDFs are released to various environmental media. However, both European and U.S. manufacturers strongly dispute the total emission factors used in Greenpeace (1993) in arriving at their estimated total of 1.8 kg TEQ/yr emitted world-wide by the PVC industry.

Greenpeace (1993) cites the same specific monitoring information as industry but argues in several case studies that "diffuse emissions" of products and byproducts containing unspecified amounts of CDDs/CDFs constitute a very significant additional source to several environmental media. This appears to be the only rationale presented by Greenpeace (1993) to justify increasing the overall emission factor of 0.01 to 0.5 g TEQ/100,000 metric tons of VCM produced, which is accepted by European manufacturers, to Greenpeace's 5 to 10 g TEQ/100,000 metric tons.

PVC production in the United States is 4.5 million metric tons per year (ChemRisk, 1993). No data could be found on dioxin levels in waste streams or air emissions from PVC plants in the United States. Applying the worldwide emission factors discussed above to the U.S. PVC industry, gives a range of dioxin emissions of 0.45 to 23 g TEQ/yr (based on the industry emission factors) to 230 to 450 g TEQ/yr (based on the Greenpeace emission factors). It is unclear whether EDC/VCM/PVC production and emission control methods are sufficiently similar worldwide to know whether these factors should apply in the United States. Considering this unknown and the lack of measurement data in general and for U.S. facilities in particular, this report does not endorse either of these emission estimates nor is an independent emission estimate presented. Also, insufficient information was provided to indicate how these emissions, if present in the United States, would separate among media. Monitoring efforts to collect these data are highly recommended.

EPA's Office of Water has promulgated effluent limitations for facilities that manufacture chlorinated aliphatic chlorine compounds and discharge treated wastewater (40 CFR 414.70). Although these effluent limitations do not specifically address CDDs and CDFs, the treatment processes required to control the chlorinated aliphatic compounds that are regulated (e.g., 68 µg/l for 1,2-dichloroethane and 22 µg/l for tetrachloroethylene) are expected to control releases of CDDs and CDFs to minimal levels.

CDD/CDF contamination of dioxazine dyes and pigments available in Canada has been observed (Williams et al., 1992). As shown in Table 3-9, OCDD and OCDF concentrations in the µg/g range, and HpCDD, HxCDD, and PeCDD concentrations in the ng/g range were found in Direct Blue 106 dye (3 samples) and Direct Blue 108 dye (1 sample) dyes and Violet 23 pigments (6 samples)(Williams et al., 1992). Dioxazine pigments (e.g., Violet 23 pigment) and dioxazine dyes (e.g., Direct Blue 106 and 108) are derived from chloranil, which has been found to contain high levels of CDD/CDFs and has been suggested as the source of contamination among these dyes (Christmann et al., 1989; Williams et al., 1992; U.S. EPA, 1992b). In May 1990, EPA received test results showing that chloranil was heavily contaminated with dioxins; levels as high as 3,065 ppb TEQ were measured (U.S. EPA, 1992b). (See Section 3.4.2 for analytical results.)

Between 1990 and 1992, EPA learned that dioxin TEQ levels in chloranil could be reduced by more than two orders of magnitude (to less than 20 ppb) through manufacturing feedstock and process changes. EPA's Office of Pollution Prevention and Toxics (OPPT) subsequently began efforts to complete an industry-wide switch from use of the contaminated chloranil to low-dioxin chloranil. Although no chloranil is manufactured in the United States, significant quantities are imported. As of June 1993, EPA had negotiated agreements with all chloranil importers and domestic dye/pigment manufacturers known to EPA who use chloranil in their products to switch to low-dioxin chloranil. EPA will issue a significant new use rule (SNUR) under Section 5 of TSCA when U.S. stocks of chloranil with high levels of CDDs/CDFs are depleted. The SNUR will require industry to notify EPA at least 90 days prior to the manufacture, import, or processing, for any use, of chloranil containing total CDDs/CDFs at a concentration greater than 20 ug/kg (Cash, 1993; U.S. EPA, 1993c).

CDD/CDFs (tetra-, penta-, and hexa-chlorinated congeners) in the ppt range were found in Ni-phthalocyanine when several commercial phthalocyanine dyes were analyzed (Hutzinger and Fiedler, 1991a). Phthalocyanine dyes and diarylide yellow pigments have also been observed to contain PCBs in the ppm range. The PCBs are believed to be generated during manufacture because of the use of high-boiling chlorinated aromatic solvents (Hutzinger and Fiedler, 1991a). EPA, however, has prohibited the processing or distribution in commerce of any diarylide and phthalocyanine pigments that contain 50 ppm or more of PCBs (40 CFR 762.20).

Based on evidence that halogenated dioxins and furans may be formed as by-products during chemical manufacturing processes (Versar, 1985), EPA proposed a rule under Section 4 of the TSCA that would require chemical manufacturers and importers to test for the presence of chlorinated and brominated dioxins and furans in certain commercial organic chemicals (Federal Register, December 19, 1985). The final rule (Federal Register, June 5, 1987) listed 12 manufactured or imported chemicals for which testing was required and 20 chemicals not currently being manufactured or imported that would require testing if manufacture or importation resumed. These chemicals are listed in Table 3-10. The specific dioxin and furan congeners for which quantitation is required and the target limits of quantitation (LOQ) specified in the Rule are listed in Table 3-11. Under Section 8(a) of TSCA, the final rule also required that chemical manufacturers submit data on manufacturing processes and reaction conditions for chemicals produced using any of the 29 precursor chemicals listed in Table 3-12. The rule stated that subsequent to this data gathering effort, testing may be proposed for additional chemicals if any of the manufacturing conditions used favored the production of dioxins and furans.

To date, data have been submitted to the EPA TSCA Docket for 10 of the 12 chemicals requiring testing, however, not every manufacturer/importer has submitted data for every applicable product (Cash, 1993). Manufacture/import of the other two substances have stopped since the test rule was promulgated. [NOTE: All data and reports in the EPA TSCA Docket are available for public review/inspection at EPA Headquarters in Washington, DC.]

The results of analytical testing for dioxins and furans for the eight chemicals for which data are available in the TSCA docket are presented in Table 3-13. Data submitted for pentabromodiphenyloxide and tetra-bromobisphenol A-bisethoxylate are currently under EPA review. Dioxins/furans were found in four of these eight chemicals. The chemicals for which positive results were obtained are: 2,3,5,6-tetrachloro-2,5-cyclohexadiene-1,4-dione (chloranil), octabromodiphenyloxide, decabromodiphenyloxide, and tetrabromobisphenol-A. Table 3-14 presents the quantitative analytical results for the four submitted chloranil samples as well as the results of verification sampling/analysis performed on chloranil by EPA.

It should be noted that although testing conducted under this test rule for 2,4,6-tribromophenol indicated no halogenated dioxins or furans above the LOQs, Thoma and Hutzinger (1989) reported detecting BDDs and BDFs in a technical grade sample of this substance. Total TBDD, TBDF, and PeBDF were found at 84 m g/kg, 12 m g/kg, and 1 m g/kg, respectively. No hexa-, hepta-, or octa-BDFs were detected. Thoma and Hutzinger (1989) also analyzed analytical grade samples of two other brominated flame retardants, pentabromophenol and tetrabromophthalic anhydride; no BDDs or BDFs were detected (detection limits not reported).

In the early 1980s, attention began to focus on pesticides as potential sources of CDDs and CDFs in the environment. Historically, no regulation had been placed on CDD and CDF levels in end-use pesticide products. Certain pesticide active ingredients were known or suspected, however, to be contaminated with CDDs and CDFs (e.g., pentachlorophenol (PCP), Silvex, and 2,4,5-T). During the mid and late 1980s, EPA took several actions to investigate and control CDD/CDF contamination of pesticides. In 1983, the sale of Silvex and 2,4,5-T was canceled for all uses by EPA (Federal Register, October 18, 1983). EPA entered into a Settlement Agreement in 1987 with PCP manufacturers to allow continued registrations for wood uses (Federal Register, January 2, 1987) but which set tolerance levels for HxCDD and 2,3,7,8-TCDD. TCDD levels were not allowed to exceed 1.0 ppb in any product, and after February 2, 1989 (a gradually phased in requirement), any manufacturing-use PCP released for shipment could not contain HxCDD levels that exceeded an average of 2 ppm over a monthly release or a batch level of 4 ppm. EPA then issued a Final Determination and Intent to Cancel and Deny Applications For Registrations of Pesticide Products Containing Pentachlorophenol (Including but not limited to its salts and esters) For Non-Wood Uses which prohibited the registration of PCP for nonwood uses (Federal Register, January 21, 1987).

In addition to these cancellations and product standards, EPA's Office of Pesticide Programs (OPP) issued two Data Call-Ins (DCIs) in June 1987. Pesticide manufacturers are required to register their products with EPA in order to market them commercially in the United States. Through the registration process, mandated by FIFRA (Federal Insecticide, Fungicide and Rodenticide Act), EPA can require that the manufacturer of each active ingredient generate a wide variety of scientific data through several mechanisms. The most common process is the five phase reregistration effort to which the manufacturers (i.e., registrants) of older pesticide products must comply. In most registration activities, registrants must generate data under a series of strict testing guidelines, 40 CFR 158--Pesticide Assessment Guidelines (U.S.EPA, 1988). FIFRA accommodates the fact that some pesticide active ingredients may require additional data, outside of the norm, to adequately develop effective regulatory policies for those products. Therefore, EPA can require additional data, where needed, through various mechanisms as noted above including the DCI process.

The purpose of the first DCI (June 6, 1987), Data Call In Notice For Product Chemistry Relating to Potential Formation of Halogenated Dibenzo-p-dioxin or Dibenzofuran Contaminants in Certain Active Ingredients, was to identify chemicals that may contain halogenated dibenzo-p-dioxin and dibenzofuran contaminants and to quantify and eventually minimize exposure to these contaminants. The requirements made in this DCI parallel requirements established in the Dioxin/Furan Test Rule promulgated under Sections 4 and 8 of TSCA. (See Section 3.4.2.) The list of pesticide active ingredients to which this DCI applied along with their corresponding Shaughnessey and Chemical Abstract code numbers are presented in Table 3-15. [Note: the Shaughnessey code is an internal EPA tracking system--it is of interest because chemicals with similar code numbers are similar in chemical nature (e.g., salts, esters and acid forms of 2,4-D)]. All registrants supporting these chemicals were subject to the requirements of this DCI unless their product qualified for a Generic Data Exemption (i.e., a registrant exclusively used a registered product(s) as the source(s) of the active ingredient(s) identified in Table 3-15 in formulating their product(s)). Registrants whose products did not meet the Generic Data Exemption were required to submit the types of data listed below to assess the formation of tetra- through hepta-halogenated dibenzo-p-dioxin or dibenzofuran contaminants during manufacture. Registrants, however, did have the option to voluntarily cancel their product or "reformulate to remove an active ingredient," described in Table 3-15, to avoid compliance with the DCI.

· Product Identity and Disclosure of Ingredients: EPA required submittal of a Confidential Statement of Formula (CSF) based on the requirements specified in 40 CFR 158.108 and 40 CFR 158.120 - Subdivision D:Product Chemistry. Registrants who had previously submitted still current CSFs were not required to resubmit this information.

· Description of Beginning Materials and Manufacturing Process: Based on the requirements mandated by 40 CFR 158.120 - Subdivision D, EPA required submittal of a manufacturing process description for each step of the manufacturing process, including specification of the range of acceptable conditions of temperature, pressure, or pH at each step.

· Discussion of the Formation of Impurities: Based on the requirements mandated by 40 CFR 158.120 - Subdivision D, EPA required submittal of a detailed discussion/assessment of the possible formation of halogenated dibenzo-p-dioxins and dibenzofurans.

The second DCI (dated June 15, 1987), Data Call-In For Analytical Chemistry Data on Polyhalogenated Dibenzo-p-Dioxins/Dibenzofurans (HDDs and HDFs), was issued for a variety of pesticide active ingredients to the individual manufacturers of each ingredient. (See Table 3-16.) All registrants supporting these pesticides were subject to the requirements of this DCI unless the product qualified for various exemptions or waivers. Pesticides regulated by the second DCI were strongly suspected to be contaminated with detectable levels of HDDs/HDFs.

Under the second DCI, registrants whose products did not qualify for an exemption or waiver were required to generate and submit the following types of data in addition to the data requirements of the first DCI:

· Quantitative Method For Measuring HDDs or HDFs: Registrants were required to develop an analytical method for assessing the HDD/HDF contamination of their products. The DCI established a regimen for defining the precision of the analytical method (i.e., for internal standard--precision within +/- 20 percent and recovery range of 50 to 150 percent, also a signal to noise ratio of at least 10:1 was required). Target quantification limits were established in the DCI for specific HDD and HDF congeners. (See Table 3-11.)

· Certification of Limits of HDDs or HDFs: Registrants were required to submit a "Certification of Limits" in accordance with 40 CFR 158.110 and 40 CFR 158.120 - Subdivision D. Analytical results were required that met the guidelines described above.

Registrants could select one of two options to comply with the second DCI. The first option was to submit relevant existing data, develop new data, or share the cost to develop new data with other registrants. The second option was to alleviate the DCI requirements through several exemption processes including a Generic Data Exemption, voluntary cancellation, reformulation to remove the active ingredient of concern, an assertion that the data requirements do not apply, or the application/award of a low-volume, minor-use waiver.

The data contained in CSFs, as well as any other data generated under Subdivision D, are typically considered Confidential Business Information (CBI) under the guidelines prescribed in FIFRA because they usually contain information regarding proprietary manufacturing processes. In general, all analytical results submitted to EPA in response to both DCIs are considered CBI and cannot be released by EPA into the public domain. Summaries based on the trends identified in that data as well as data made public by EPA are provided below.

To date, more than 100 submissions have been reviewed in response to the two DCIs. The majority have been manufacturing process data in support of waiver requests, analytical method protocols, and sample collection protocols (telephone conversation between S. Funk, EPA - Office of Pesticide Programs (OPP), and J. Dawson, Versar, Inc. on 2/18/93). Analytical results on the levels of tetra- through hepta- HDDs/HDFs have been received and reviewed for 16 distinct pesticide active ingredients (Table 3-17). In general, the analyses have not revealed HDD/HDF concentrations in excess of the LOQs specified in Table 3-11. For those products in which LOQs are exceeded, the identified contamination levels were generally within an order of magnitude of the LOQ and apply only to one or two congeners per product (telephone conversation between S. Funk, EPA/

OPP, and J. Dawson, Versar, Inc. on 2/18/93). Table 3-18 presents a summary of results recently reported by EPA for CDDs and CDFs in eight technical 2,4-D herbicides.

The production and use of chlorine gas has involved processes that result in the generation of CDFs (Rappe, 1992a). Chlorine is commonly produced via electrolysis of brine in mercury cells. High levels of CDFs have been found in the graphite electrode sludge from this chemical process and may have been responsible for occupational exposures among workers who handled these sludges. Svensson et al. (1992) evaluated the relationship between blood CDF levels in chloralkali plant workers and direct exposure of these workers to electrode sludges and to dust and earth contaminated with graphite electrode sludge. Subjects who had been exposed by handling graphite electrode sludge had higher levels of 2,3,7,8-substituted PeCDFs and HxCDFs than reference subjects. Evaluations of congener distribution patterns have demonstrated that the 2,3,7,8-substituted CDFs are the major congeners formed during the chloralkali process (Rappe et al., 1990; Rappe, 1992a).

Until the late 1970s, graphite electrodes were the primary type of anode used in the chloralkali industry (Curlin and Bommaraju, 1991). Since then, metal anodes have been developed to replace graphite electrodes because of production problems associated with their use (U.S. EPA, 1982; Curlin and Bommaraju, 1991). Currently, no U.S. facilities are believed to use graphite electrodes in the production of chlorine gas (telephone conversation between L. Phillips, Versar, Inc., and T. Fielding, U.S. EPA, Office of Water, February 1993). Although the use of graphite electrodes has been eliminated, the potential for CDD/CDF releases from dump sites containing contaminated sludges may still exist (Svensson et al., 1992; Rappe, 1992a).

Catalyst regeneration in the petroleum refinery reforming process has been identified as a source of CDDs and CDFs based on testing conducted in Canada (Thompson et al., 1990). According to Thompson et al. (1990), "catalytic reforming is a refinery process which is used to produce high octane gasoline. The reforming process occurs at high temperature and pressure and requires the use of a catalyst. During the catalytic process, a complex mixture of aromatic compounds known as coke is formed and deposited onto the catalyst. As coke deposits onto the catalyst, its activity is decreased. The high cost of the catalyst necessitates its regeneration. Catalyst regeneration is achieved by removing the coke deposits via burning and activating the catalyst using chlorinated compounds. Burning of the coke produces flue gases which contain CDDs and CDFs along with other combustion products." Thompson et al. (1990) reported total CDD and CDF concentrations of 8.9 ng/m³ and 210 ng/m³, respectively, in stack gas samples from petroleum refinery reforming operations (Table 3-19). It was also found that the CDD and CDF congener distribution patterns observed were similar to those found in municipal waste incinerator ash and stack samples. Because flue gases may be scrubbed with water, internal effluents may also be contaminated with CDD/CDFs. Thompson et al. (1990) observed CDDs and CDFs in the internal wash water from a scrubber of a periodic/cyclic regenerator (Table 3-20).

The Canadian Ministry of the Environment detected concentrations of CDDs in an internal wastestream of spent caustic in a petroleum refinery that ranged from 1.8 to 22.2 ppb, and CDFs ranging from 4.4 to 27.6 ppb. The highest concentration of 2,3,7,8-TCDD was 0.0054 ppb (Maniff and Lewis, 1988). CDDs were also observed in the refinery's biological sludge at a maximum concentration of 74.5 ppb, and CDFs were observed at a

maximum concentration of 125 ppb (Maniff and Lewis, 1988). The concentration of CDD/CDFs in the final combined refinery plant effluent was below the detection limits.

Insufficient data are available to evaluate CDD/CDF releases from these sources in the United States. However, Beard et al. (1993) conducted a series of benchtop experiments to investigate the mechanism(s) of CDD/CDF formation in the catalytic reforming process. A possible pathway for the formation of CDFs was found, but the results could not explain the formation of CDDs. Analyses of the flue gas from burning coked catalysts revealed the presence of unchlorinated dibenzofuran (DBF) produced in quantities of up to 220 ng/g of catalyst. Chlorination experiments indicated that dibenzofuran and possibly biphenyl and similar hydrocarbons act as CDF precursors and can become chlorinated in the catalyst regeneration process. Corrosion products on the steel piping of the process plant seem to be the most likely chlorinating agent. Furthermore, CDFs can form by de novo synthesis from chlorinated hydrocarbons like trichloroethylene, methylene chloride, and carbon tetrachloride in the presence of FeCl₃ and HCl or Cl₂.

Rappe et al. (1989) reported the formation of CDFs (tetra- through octa-chlorinated CDFs) when tap water and double-distilled water were chlorinated using chlorine gas. The CDF levels found in the single samples of tap water and double-distilled water were 35 and 7 pg TEQ/L, respectively. The water samples were chlorinated at a dosage rate of 300 mg of chlorine per liter of water which is considerably higher (by a factor of one to two orders of magnitude) than the range of dosage rates typically used to disinfect drinking water. Rappe et al. (1989) hypothesized that the CDFs or their precursors are present in chlorine gas. It should be noted, however, that although few surveys of finished drinking water for CDD/CDF levels have been conducted, the few that have been published only rarely report the presence of any CDD/CDF even at low pg/L detection limits and in those cases the CDD/CDFs were also present in the untreated water. (See Section 4.3.)

Several recent studies have been conducted to identify the source(s) of CDD/CDFs found in textiles and at dry cleaning facilities. Horstmann and McLachlan (1994) analyzed 35 new textiles and found total CDD/CDF levels generally less than 50 pg/g; however, some items were as high as 290,000 pg/g. The authors conclude that textile finishing processes are not likely to be the source of the high CDD/CDF levels found because of the apparent randomness of the textiles with high CDD/CDF levels. However, the authors hypothesize that the use of pentachlorophenol to preserve cotton, particularly when it is randomly strewed on bales of cotton as a preservative during sea transport, is the likely source of the high levels occasionally observed. As discussed in Section 3.4.3, the use of pentachlorophenol (PCP) for nonwood uses has been prohibited in the United States since 1987. However, Horstmann and McLachlan (1994) comment that PCP is still used in developing countries, especially for purposes of preserving cotton during sea transport. As discussed in Section 3.4.1.5, certain dyes and pigments have also been observed to contain CDD/CDFs and may also contribute to levels found in textiles. Horstmann and McLachlan (1994) also summarize recent research concerning CDD/CDFs in dry cleaning residues and reach the conclusion that new textiles are the source of the CDD/CDFs found.

3.5. MECHANISMS OF FORMATION OF DIOXIN-LIKE COMPOUNDS DURING COMBUSTION OF ORGANIC MATERIALS

The specific molecular mechanisms by which CDDs and CDFs are initially formed and then emitted from combustion sources remain largely unknown and are theoretical. The theoretical basis for conjecture is derived primarily from direct observations in municipal solid waste incinerators and from well conducted laboratory studies. Municipal solid waste incinerators (MSWIs) have been heavily studied from the perspective of eventually finding the specific formation mechanism transpiring within the system, and determining ways to either significantly reduce such opportunities or ultimately hinder the formation kinetics to preclude evolution of these chemicals. Although much has been learned from these studies, it is still not known how to completely block the formation of CDDs/CDFs during the combustion of certain organic materials in the presence of a source of chlorine. Adding to this complexity is the wide variability of organic materials that are incinerated and thermally processed by a wide spectrum of combustion technologies having variable temperatures, residence times, and oxygen requirements. However, it is possible to identify the central chemical events participating in the formation of CDDs and CDFs by evaluating emission test results from MSWIs in combination with laboratory experiments.

The emission of CDDs and CDFs can be explained by three principal theories, which should not be regarded as being mutually exclusive. The first is that CDD/CDFs are present as contaminants in the combusted organic material. This theory is discussed in Section 3.5.1. The second is that CDDs/CDFs are ultimately formed from the thermal breakdown and molecular rearrangement of precursor compounds, which are defined as chlorinated aromatic hydrocarbons having a structural resemblance to the CDD/CDF molecule. This theory is discussed in Section 3.5.2. The third theory, similar to the second and described in Section 3.5.3, is that CDDs/CDFs are synthesized de novo; this means they are formed from organic and inorganic substrates comprised of singular or mixtures of molecules bearing little resemblance to the molecular structure of CDDs or CDFs. Section 3.5.4 discusses the generation of coplanar PCBs. Section 3.5.5 discusses the evaluation of naturally occurring CDDs/CDFs by examinations of sediment core data, and Section 3.5.6 provides a closing summary of the three principal theories of formation.

The first theory states that CDD and CDF compounds present as contaminants in the fuel or waste products that are fed into the combustion chamber are responsible for dioxin and dibenzofuran emissions out the stack of the combustion process. Most work in this area has involved the study of municipal solid waste incineration (MSWI) in which case CDDs and CDFs have been analytically detected in the raw refuse fed into the MSWI. Tosine, et al. (1983) first reported detecting trace amounts of HpCDD and OCDD in the MSW fed into an MSWI in Canada. HpCDD ranged in concentration from 100 ppt to 1 ppb, and OCDD ranged from 400 to 600 ppt. Wilken et al. (1992) separated the various solid waste fractions of MSW collected from municipalities in Germany and analyzed them for the presence of CDDs/CDFs and other organochlorine compounds. Total CDDs/CDFs were detected in all MSW fractions in the following range of concentrations: paper and cardboard = 3.1 to 45.5 ppb; plastics, wood, leather, textiles combined = 9.5 to 109.2 ppb; vegetable matter = 0.9 to 16.9 ppb; and "fine debris" (defined as particles < 8 mm) = 0.8 to 83.8 ppb. Ozvacic (1985) measured CDDs/CDFs in the raw MSW fed into two MSWIs operating in Canada. In one MSWI, CDDs were detected in the refuse in a range of concentration from 10 to 30 ppb, but no CDFs were detected (detection limit: 1 pg/g). In the MSW fed to the second MSWI, CDDs were detected in a range of 75 to 439 ppb, and CDFs were detected only in one of three samples at a total concentration of 11 ppb. EPA has reported on the detection of CDDs/CDFs in refuse derived fuel (RDF) burned in a large, urban MSWI (Federal Register, 1991). From 13 MSW samples taken prior to incineration, CDDs were detected in a range of 1 to 13 ppb, and CDFs were measured in a range of 0 to 0.6 ppb. In these samples, OCDD predominated, and the lower chlorinated congeners were not detected.

Despite these findings, the conditions of thermal stress imposed by the incineration process discounts the likelihood that the total magnitude of CDDs and CDFs, as measured in the raw MSW, can explain the total magnitude of concentration as an emission from the stack of the MSWI (Clement et al., 1990; Commoner, 1990). Contamination, however, may partially contribute to the stack release. Clement and coworkers (1988) performed a mass balance involving an input versus output of dioxin at two operational MSWIs in Canada. These mass balance calculations clearly demonstrated that the mass of CDDs and CDFs emitted at the point of the stack was much greater than the mass in the raw MSW incinerated at the MSWIs, and that the profiles of the distributions of CDD/CDF congeners were strikingly different. Primarily, higher chlorinated congeners were detected as contaminants in the waste, whereas the total array of tetra - octa CDDs/CDFs were emitted from the stack.

Commoner and coworkers (1984; 1985; 1987) evaluated the test data of a mass burn MSW incinerator for the concentration of CDDs and CDFs at multiple sampling points during the combustion process: (1) exit to the furnace; (2) entry to the heat exchanger; (3) inlet to the electrostatic precipitator (ESP); (4) exit to the ESP; and (5) exit to the smokestack. Lowest or nondetectable concentrations of CDDs/CDFs were found at sampling point (1), and highest concentrations were measured at sampling point (5). From these sampling data, Commoner concluded that: CDDs/CDFs were not formed within the furnace region where the waste material was combusted and that usually only OCDD and OCDF were detected in extremely low concentrations at the point of exit to the furnace (if dioxins were detected at all). It was also concluded that the CDDs/CDFs were mostly formed as a synthesis process catalyzed by the properties of fly ash in combination with chlorine, and that this probably transpired within areas downstream of the combustion zone where the combustion offgases had cooled to less than 400° C. Commoner et al. (1984, 1987) ruled out the effectiveness of combustion as a major factor in CDD/CDF emissions from the stack; this would be expected if waste contamination was solely responsible for the emission. This phenomena was independently observed by Environment Canada in a series of tests of a modular MSW incinerator (Hay, et al., 1986; Environment Canada, 1985). On a mass balance basis, the concentration of CDDs and CDFs measured at the stack was approximately two orders of magnitude higher as compared to the inlet to the boiler just after exiting the secondary furnace. The temperatures of the combustion gases at these two points of measurement were 130 and 740° C at the stack and boiler inlet, respectively (Environment Canada, 1985). For the most part, only OCDD was present in the hot gases exiting the furnace, whereas all the congeners were present in the stack emissions, thus giving further evidence that CDDs/CDFs are formed after the combustion zone. Using similar protocols, EPA and Environment Canada (1991) jointly evaluated the emission of CDDs and CDFs from a refuse-derived fuel MSWI operating in the United States. It was found that approximately 5 milligrams of total CDDs and CDFs per metric ton of MSW burned by the facility were measured in the raw MSW prior to combustion, but no CDDs nor CDFs were detected at the point of exit to the furnace prior to the inlet to the economizer (i.e., the heat exchanger used to extract additional heat from the hot gases). Once heat in the combustion gas was extracted for energy purposes and the gases were further cooled to less than 400° C, the total array of tetra- through octa-CDDs and CDFs could be detected.

These series of experiments in which the mass balance of CDD/CDF was estimated within the entire combustor, beginning with the waste and ending with the stack, discount the first theory of dioxin formation (i.e., that dioxin in the feed accounts for all emissions of dioxin from the stack to the air). Moreover, it is expected that the conditions of thermal stress imposed by typical incineration and other combustion sources would destroy and reduce the CDDs and CDFs present as contaminants in the waste to levels that are 0.0001 to 10 percent of the initial concentration, depending on the performance of the combustion source and the level of combustion efficiency. Stehl et al. (1973) demonstrated that the moderate temperature of 800° C enhances the decomposition of CDDs at a rate of about 99.95 percent, but that lower temperatures result in a higher survival rate. Theoretical modeling has shown that unimolecular destruction of CDDs/CDFs at 99.99 percent can occur at the following temperatures and retention times within the combustion zone: 977° C with a retention time of 1 second; 1000° C at a retention time of 1/2 second; 1227° C at a retention time of 4 milliseconds; and 1727° C at a retention time of 5 microseconds (Schaub and Tseng, 1983). Thus, CDDs and CDFs would have to be in parts per million concentration in the feed to the combustor to be found in the part per billion or part per trillion levels in the stack gas emission (Shaub and Tseng, 1983). However, it cannot be ruled out is that CDDs/CDFs in the waste or fuel may contribute (up to some percentage) to the overall concentration leaving the stack.

The second theory states that the production of CDDs and CDFs is a direct result of in-situ thermal degradation of precursor compounds during or after combustion of organic materials. Present theory is mostly derived from laboratory experiments involving the heating of suspect precursors in quartz ampules under starved-air conditions, and in experiments investigating the role that combustion fly ash has in promoting the formation of CDD/CDFs from precursor compounds.

Liberti and Brocco (1982) postulated that the general reaction that may be taking place in a typical combustion process is a thermolytic synthesis and interaction between two families of precursors indicated by A and B. Precursors A are aromatic compounds having a definite phenolic structure (e.g., phenol and polychlorinated phenols), and precursors B are chemical species that can act as a chlorine donor (e.g., PVC and HCl). Esposito et al. (1980) offered a chemical basis for defining a dioxin precursor:

1. The compound is comprised of an ortho-substituted (positions 1 and 2 on the compound) benzene ring in which one of the substituents is an oxygen atom directly attached to the ring, and

2. It then must be possible for the two substituents on the benzene ring to react with each other to form a new and independent compound under the influences of heat and pressure (i.e., dioxin).

Dickson and Karasek (1987) further refined this definition to be consistent with the formation kinetics thought to occur within combustion processes. In their definition, the term "precursor" refers specifically to chlorinated aromatic compounds that are either already present on the surface of combustion fly ash, or are present in the gas phase prior to entering a critical region outside the combustion zone where the gases have cooled and where heterogeneous catalyzed reactions take place that form CDDs/CDFs. Chlorophenols and chlorobenzenes were identified as ideal precursor compounds in these reaction pathways.

Controlled laboratory combustion experiments involving the thermal degradation of aromatic compounds, either singly or in mixtures, have provided useful data in identifying ideal precursor compounds. For example, Jansson and coworkers (1977) generated CDDs through the pyrolysis of wood chips treated with tri-, tetra-, and penta-chlorophenol in a bench-scale furnace operated at 500-600° C. Stehl and Lamparski (1977) combusted grass and paper treated with the herbicide 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) in a bench-scale furnace at 600-800° C and generated ppm_v levels of TCDD. Ahling and Lindskog (1982) have reported on the formation of CDDs during the combustion of tri- and tetrachlorophenol formulations at temperatures of 500-600° C. Decreases in oxygen during combustion generally increased the yield, and the addition of copper salts to the tetrachlorophenol formulation significantly enhanced the yield of CDDs. Combustion of pentachlorophenol resulted in low yields of CDDs except when burned with an insufficient supply of oxygen. In that case, the investigators noted the formation of tetra- through octa-chlorinated congeners. Buser (1979) generated CDDs/CDFs on the order of 0.001-0.08 percent (by weight) by heating tri-, tetra-, and pentachlorobenzenes at 620° C in quartz ampules in the presence of oxygen. It was noted that chlorophenols were formed as combustion by-products, and Buser (1979) speculated that these were acting as reaction intermediates in the formation of CDDs/CDFs.

Recently it has been demonstrated that CDDs and CDFs are formed from aromatic precursor compounds adsorbed onto the reactive surface of fly ash (particulate matter) entrained in the combustion plasma. Moreover, formation occurs outside and downstream of the combustion zone of a furnace to a combustion source in regions where the temperature of the combustion offgases has cooled to between 200 and 400° C (Vogg et al., 1987; Bruce et al., 1991; Cleverly et al., 1991; Gullet et al., 1990a; Commoner et al., 1987; Dickson and Karasek, 1987; Dickson et al., 1992). Vogg and coworkers (1987) have shown that inorganic chloride ions, such as copper chloride, present in the combustion gas may act as a catalyst to promote surface reactions on particulate matter to convert aromatic precursor compounds to chlorinated dioxins and dibenzofurans. After carefully extracting organics from MSWI fly ash, Vogg et al. (1987) added a known concentration of isotopically labeled CDDs/CDFs to the matrix. The MSWI fly ash was then heated in a laboratory furnace at varying temperatures for 2 hours. The treated fly ash was exposed to increasing temperatures in 50° C increments in a temperature range of 200 to 400° C. Table 3-21 summarizes these data.

Because the relative concentration of CDDs/CDFs increased while exposed to varying temperature, Vogg, et al. (1987, 1992) concluded that formation of CDDs and CDFs from precursor compounds on the surface of fly ash transpires during MSW incineration within a specific range of temperature, 250 to 450° C. Within this range, the concentration of CDDs/CDFs increases to some maxima, and outside this range the concentration diminishes. Vogg et al. (1987) proposed an oxidation reaction pathway giving rise to the formation of CDDs and CDFs in the post-furnace regions of the incinerator in the following order: (1) hydrogen chloride gas (HCl) is thermolytically derived as a product of the combustion of heterogeneous fuels containing abundant chlorinated organic chemicals and chlorides; (2) oxidation of HCl, with copper chloride (CuCl₂) as a catalyst, yields free gaseous chlorine; (3) phenolic compounds (present from combustion of lignin in the waste or other sources) entrained in the combustion plasma are substituted on the ring structure by contact with the free chlorine; and (4) the chlorinated precursor to dioxin (e.g., chlorophenol) is further oxidized (with copper chloride as a catalyst) to yield CDDs and CDFs and chlorine.

Gullett and coworkers (1990a; 1990b; 1991a; 1991b; 1992) have studied the formation mechanisms through extensive combustion research at EPA, and have verified the observations of Vogg et al. (1987). It was proven that CDDs and CDFs could be ultimately produced from low temperature reactions (i.e., 350° C) between Cl₂ and a phenolic precursor combining to form a chlorinated precursor, followed by oxidation of the

chlorinated precursors (catalyzed by a copper catalyst such as copper chloride) as in examples (1) and (2), below.

(1) The initial step in the formation of dioxin is the formation of chlorine from HCl in the presence of oxygen (the Deacon process), as follows (Vogg et al., 1987; Bruce et al., 1991):

_

2HCl + 1/2 O₂ ^{——————>} H₂O + Cl₂

(2) Phenolic compounds adsorbed on the surface of fly ash are chlorinated to form the dioxin precursor, and the dioxin is formed as a product from the breakdown and molecular rearrangement of the precursor. The reaction is promoted by the presence of heat and copper chloride acting as a catalyst (Vogg et al., 1987; Gullett et al., 1992):

(a) phenol + Cl₂ ————> chlorophenol (dioxin precursor)

CuCl₂

(b) 2-chlorophenol + 1/2 O₂ ————> dioxin + Cl₂

The major direct source of chlorine available for participating in the formation of CDDs/CDFs is gaseous HCl, which is initially formed as a combustion by-product from the chlorine and chlorinated organic chemicals contained in the MSW (and other fuels) (Vogg et al., 1987; Bruce et al., 1991; Cleverly, 1984; Commoner et al., 1987). MSW contains approximately 0.45-0.90 percent (by weight) chlorine (Domalski et al., 1986). MSW incinerators are a major stationary combustion source of air emissions of HCl, which average between 400 to 600 ppm in the combustion gas (U.S. EPA, 1987). HCl is converted to chlorine vapor by the Deacon process, and the vapor phase chlorine directly chlorinates a dioxin precursor along the aromatic ring structure. Oxidation of the chlorinated precursor in the presence of an inorganic chloride metal catalyst (of which copper chloride was found to be the most active) yields CDDs and CDFs. Increasing the yield of chlorine in vapor phase from the oxidation of HCl generally causes an increase in the rate of formation of CDDs/CDFs. Formation kinetics are most favored at temperatures between 200 to 350° C. Reductions in chlorine production, either by limiting initial HCl concentration or by shortening the residence time in the Deacon process temperature window, should result in decreases in the rate and magnitude of formation of CDDs and CDFs (Bruce et al., 1991; Gullett et al., 1990b; Commoner et al., 1987). Bruce and coworkers (1991) observed a general increase in the formation of CDDs and CDFs with increases in the vapor phase concentration of chlorine. Figure 3-2 shows the apparent dependence of the extent of formation of CDDs and CDFs upon chlorine concentration in

the vapor phase. Bruce et al. (1991) verified a dependence on the concentration and availability of gaseous chlorine in the thermolytic formation of CDDs/CDFs.

In the testing of a variety of industrial stationary combustion sources during the National Dioxin Study in 1987, EPA made a series of qualitative observations on the relationship between total chlorine present in the fuel/waste and the magnitude of emissions of CDDs and CDFs from the stack of the tested facilities (U.S. EPA, 1987). In general, combustion units with the highest CDD emission concentrations had greater quantities of chlorine in the fuel, and, conversely, sites with the lowest CDD emission concentrations contained only trace quantities of chlorine in the feed. The typical chlorine content of various combustion fuels has been reported by Lustenhouwer et al. (1980) as: coal: 1,300 µg/g; MSW: 2,500 µg/g; leaded gasoline: 300-1,600 µg/g; unleaded gasoline: 1-6 µg/g.

The role that temperature plays in the formation kinetics has been investigated by Oberg et al. (1989) on a full-scale hazardous waste incinerator operating in Sweden. Oberg confirmed that the formation of CDDs/CDFs occurs after the furnace. Most of the formation transpired in the boiler used to extract heat for co-generation of energy. In this investigation, significant increases in total concentration of dioxin TEQ occurred between temperatures of 280-400° C, and concentrations declined at temperatures above 400° C. This is in agreement with the experimental evidence of the temperature range defined as the "window of opportunity" for catalytic formation of CDDs/CDFs on the surfaces of fly ash particles.

Dickson and Karasek (1987) have demonstrated that CDDs/CDFs can be directly formed from the thermal conversion and oxidation of chlorinated precursors, in particular chlorophenols, on the surface of MSWI fly ash while heated in a bench-scale furnace. Their experiment was designed to mimic conditions of MSW incineration; to identify the step-wise chemical reactions involved in converting a precursor compound into dioxin, and to determine if MSWI fly ash could promote these reactions. MSWI fly ash was obtained from a facility in Canada and a facility in Japan. The MSWI fly ash was extensively solvent-extracted for any organic constituents prior to initiating the experiment. Twenty grams of fly ash were introduced into a bench-scale oven (consisting of a simple flow-tube combustion apparatus) and heated at 340° C overnight to desorb any remaining organic compounds from the matrix. ¹³C₁₂ -labeled pentachlorophenol (PCP) and two trichlorophenol isotopes (¹³C₁₂- 2,3,5-T and 3,4,5-T) were added to the surface of the clean fly ash matrix, and placed into the oven for 1 hour at 300° C. Pure inert nitrogen gas (flow rate of 10 ml/min) was passed through the flow tube to maintain constant temperatures. Tetra- through octa- CDDs were formed from the labeled pentachlorophenol experiment; over 100 µg/g of total CDDs were produced. The congener pattern was similar to the congener pattern found in MSWI emissions. The 2,4,5-T experiment primarily produced HxCDDs and very small amounts of tetra- and octa-CDD. The 3,4,5-T experiment mainly produced OCDD and 1,2,3,4,6,7,8-HpCDD. Dickson and Karasek (1987) proposed that the chlorinated phenol may undergo molecular rearrangement or isomerization as a result of dechlorination, dehydrogenation, and trans-chlorination before condensation occurs to ultimately form CDDs on the fly ash surface. These reactions ultimately dictate the types and amounts of CDDs that are formed.

Nestrick and coworkers (1987) reported on the thermolytic reaction between benzene (an unsubstituted precursor) and iron (III) chloride on a silicate surface to yield CDDs/CDFs at temperatures ³ 150° C. The experimental protocol was to introduce 100 - 700 mg of native and ¹³C₆-benzene into a macro-reactor system consisting of a benzene volatilization chamber connected to a glass tube furnace. The investigators noted the relevance of this experiment to generalizations about combustion processes because benzene is the usual combustion by-product of organic fuels. Inert nitrogen gas was used to carry the benzene vapor to the furnace area. The exit to the glass tubing to the furnace was plugged with glass wool, and silica gel was introduced from the entrance end to give a bed depth of 7 cm to which the FeCl₃ was added to form a FeCl₃/silica reagent. The thermolytic reaction took place in a temperature range of 150-400° C at a residence time of 20 minutes. Although di- through octa-CDD/CDF were formed by this reaction at all the temperatures studied, the percent yields were extremely small. Table 3-22 summarizes these data.

The third and last theory states that CDDs/CDFs are formed in combustion processes from materials and/or compounds that are not structurally related to CDDs/CDFs on a molecular level. As in Theory 2, synthesis is believed to occur in regions outside of the furnace zone of the combustion process where the combustion plasma has cooled to a range of temperatures considered favorable to formation kinetics. A key component to de novo synthesis is the production of intermediate compounds (either halogenated or non-halogenated) that are precursors to dioxin formation. However, research in this area has produced CDDs/CDFs directly from the heating of carbonaceous fly ash in the presence of

an inorganic ion without the apparent generation of reactive intermediates. Thus, the specific steps involved in the de novo process have not been fully and succinctly delineated. Laboratory experimentation has proven that MSWI fly ash, itself, is not an inert substrate, and the matrix can actually participate in the formation kinetics. Typically the fly ash is composed of an alumina-silicate construct with 5-10 percent concentrations of silicon, chlorine (as inorganic chlorides), sulfur, and potassium (NATO, 1988). Twenty percent of the weight of fly ash particles are carbon, and the particles have specific surface areas in the range of 2-4 m² (NATO, 1988). The distinguishing feature of the de novo synthesis over the precursor synthesis is the thermolytic breakdown and molecular rearrangement of chemical species unrelated to CDDs/CDFs at the start of the process to yield precursor compounds. Theory 2 starts with the precursor compounds already adsorbed onto the surface of fly ash or present in the gas phase (Dickson et al., 1992). By this distinction, however, one could argue that Theory 3 is really an augmentation to Theory 2 because the generation of CDDs/CDFs may still require the formation of a dioxin precursor. Nevertheless, a distinction is presented here for purposes of describing

additional pathways that have been suggested for the thermal formation of these compounds.

To delineate the de novo synthesis of CDDs/CDFs from unrelated matter, Stieglitz and coworkers (1989a) have conducted experiments involving the heating of particulate carbon containing adsorbed mixtures of Mg and Al-silicate in the presence of copper chloride as a catalyst to the reaction. The authors described annealing mixtures of Mg-Al silicate with activated charcoal (4 percent by weight), chloride as potassium chloride (7 percent by weight), and 1 percent copper chloride (CuCl₂) (in water) in a glass tube at 300° C. The retention time was varied at 15 minutes, 30 minutes, and 1, 2, and 4 hours to obtain differences in the amounts of CDDs/CDFs that could be formed. The results are summarized in Table 3-23.

In addition to the CDDs/CDFs formed as primary products of the de novo synthesis, the investigators observed the formation of precursors at the varying retention times of the experiment. In particular, similar yields of tri- though hexa-chlorobenzenes, tri- through hepta-chloronaphthalenes, and tetra- through hepta-chlorobiphenyls, were quantified which were seen as highly suggestive of the role these compounds may play as intermediates in the continued formation of CDDs/CDFs. Table 3-24 summarizes the experimental yields of chlorinated benzenes as a function of the annealing time at 300° C. Stieglitz et al. (1989a) made the following observations:

1. The de novo synthesis of CDDs/CDFs via the reaction of carbonaceous particulate matter exposed to a temperature of 300° C was clearly demonstrated. Additionally, the experiment yielded ppb-ppm concentrations of chlorinated benzenes, chlorinated biphenyls, and chlorinated napthalenes through a similar mechanism. When potassium bromide was substituted for potassium chloride as a source of halogen for the organic compounds in the reaction, polybrominated dibenzo-p-dioxins and dibenzofurans were formed as reaction products.

2. Copper chloride catalyzed the de novo synthesis of CDDs/CDFs on the surface of particulate carbon in the presence of oxygen to yield carbon dioxide and chlorinated/brominated aromatic compounds.

3. Particulate carbon, which is characteristic of combustion processes, may act as the source for the direct formation of CDDs/CDFs as well as other chlorinated organics.

More recently, Stieglitz and coworkers (1991) investigated the role that particulate carbon plays in the de novo formation of CDDs/CDFs from fly ash containing appreciable quantities of organic chlorine. Stieglitz et al. (1991) found that the fly ash contained 900 µg/g of bound organic chlorine. Only 1 percent of the organic chlorine was extractable. Annealing the fly ash at 300-400° C for several hours caused the carbon to oxidize leading to a reduction in the total organic chlorine in the matrix, and a corresponding increase in the total extractable organic chlorine (TOX) (e.g., 5 percent extractable TOX at 300° C and 25-30 percent extractable total organic chlorine at 400° C). From this, Stieglitz et al. (1991) concluded that the oxidation and degradation of carbon in the fly ash are the source for the formation of CDDs/CDFs, and, therefore, are essential in the de novo synthesis of these compounds.

Addink et al. (1991) conducted a series of experiments to observe the de novo synthesis of CDDs/CDFs in a carbon-fly ash system. In this experiment, 4 grams of carbon-free MSWI fly ash were combined with 0.1 gram of activated carbon and placed into a glass tube between two glass wool plugs. The glass tube was then placed into a furnace at a specific temperature in the range 200 to 400° C. This was repeated for a series of retention times and temperatures. The investigators observed that the formation of CDDs/CDFs was optimized at the temperature of 300° C and at the furnace retention times of 4-6 hours. Figure 3-3 displays the relationship between retention time, temperature and the production of CDDs/CDFs from the heating of carbon particulate. Addink et al. (1991) also investigated the relationship between temperature of the furnace

and the production of CDDs/CDFs from the annealing of carbonaceous fly ash. Figure 3-4 displays this relationship. In general, the concentration began to increase at 250° C and crested at 350° C, with a sharp decrease in concentration above 350° C. The authors also noted a relationship between temperature and the CDD/CDF congener profile; at 300° C to 350° C, the lower chlorinated tetra- and penta-CDD/CDF congeners increased in concentration, while hexa-, hepta-, and octa-CDD/CDF congeners either remained the same or decreased in concentration. The congener profile of the original MSWI fly ash (not subject to de novo experimentation) was investigated with respect to changes caused by either temperature or residence time in the furnace. No significant changes occurred, leading the authors to propose an interesting hypothesis for further testing: after formation of CDDs/CDFs occurs on the surface of fly ash, the congener profile remains fixed and insensitive to changes in temperature or residence time indicating some form of equilibrium is reached in the formation kinetics.

Gullett et al. (1994) developed a pilot-scale combustor to study the effect on CDD/CDF formation of varying the combustion-gas composition, temperature, residence time, quench rate, and sorbent (Ca[OH]₂) injection. The fly ash loading was simulated by the injection on fly ash collected from a full-scale MSWI. Sampling and analysis indicated CDD/CDF formation or the injected fly ash at levels representative of those observed at full-scale MSWIs. A statistical analysis of the results showed that, although the effect of combustor operating parameters of CDD/CDF formation is interactive and very complicated, substantial reduction in CDD/CDF formation can be realized with high temperature sorbent injection to reduce HCl or Cl₂ concentrations, control of excess air (also affects ratio of CDDs to CDFs formed), and increased quench rate.

The de novo theory also considers the generation of CDDs/CDFs from the combustion of PVC resin. Key to the de novo synthesis of CDDs/CDFs is the initial formation of HCl from combustion. Paciorek and coworkers (1974) thermally degraded pure PVC resin at 400° C and produced 550 mg/g HCl vapor as a primary thermolysis product, which was observed as being 94 percent of the theoretical amount based on the percent weight chlorine on the molecule. Ahling et. al. (1978) have concluded that HCl can act as a chlorine donor to ultimately yield chlorinated aromatic hydrocarbons from the thermolytic degradation of pure PVC, and that these yields are a function of transit time, percent oxygen, and temperature. The data they observed from 11 separate experiments conducted with a range of temperatures from 570-1130° C indicated that significant quantities of various isomers of dichloro-, trichloro-, tetrachloro-, and hexachlorobenzenes could be produced. Choudhry and Hutzinger (1983) proposed that the radical species Cl× and H× generated in the incineration process may attack the chlorinated benzenes thus formed, and abstract hydrogen atoms to produce ortho-chlorine substituted chlorophenol radicals. These intermediate radical species then react with molecular oxygen to yield ortho-substituted chlorophenols. As a final step, the ortho-substituted chlorophenols act as ideal precursors to yield CDDs/CDFs with heat and oxygen.

Although most of the aforementioned experiments have involved the pyrolysis of anthropogenic substances, the de novo formation of CDDs/CDFs is theoretically proposed to include the combustion of autochthonous (naturally occurring) organic substances (Choudhry and Hutzinger,1983) in the presence of a chlorine donor. This possibility was first advanced by scientists at Dow Chemical Co. in 1978 in a proposed working hypothesis known as "the trace chemistries of fire" (Crummett, 1982). This proposed working hypothesis was based on the following observations:

1. Combustion processes are seldom more than 99.9 percent efficient in converting carbonaceous fuel into carbon dioxide.

2. The remaining 0.1 percent of the fuel is converted into traces of organic species including complex halogenated aromatic hydrocarbons. Most of these compounds have not been identified in combustion emissions.

3. Municipal solid waste and fossil fuels contain complex mixtures of diverse chemical species at variable concentrations.

4. Combustion fuels contain chlorine in a range of 1-5000 parts per million.

5. Particulate matter that is emitted from oil-fired heating and power plants contain vanadium and nickel. Particulates emitted from coal-fired power plants contain vanadium, nickel, iron, and manganese. In combination with silicon and unburned carbon, these species can act as catalysts in the combustion process to form halogenated aromatic hydrocarbons.

6. Chemical reactions that occur in flames include pyrolysis, oxidation, and reduction. Ions, electrons, free radicals, and free atoms interact in a continuously changing environment.

7. Dow scientists found traces of CDDs/CDFs in all particulate matter samples taken from areas that were in close proximity to combustion sources.

8. Precursors for the formation of CDDs have been experimentally proven, and have been identified to be primarily chlorinated phenols and chlorinated benzenes. Because the pyrolysis of polyvinyl chloride (PVC) produces chlorobenzenes, the combustion of PVC may cause the formation of CDDs/CDFs.

Dow Chemical Co. invited the scientific community at large to give advice on ways in which the trace chemistries of fire hypothesis could be tested. The following studies were proposed as a means of testing the hypothesis (Crummett, 1982):

1. Determine if CDDs/CDFs are present in soils (having a relatively high carbon content) taken from drill cores beneath ancient lake beds at depths corresponding to 5, 12, and 35,000 years of sedimentation and deposition.

2. Determine if CDDs/CDFs are present in ice core samples taken from the center of an ancient glacier.

3. Determine if CDDs/CDFs are present in volcanic ash.

4. Determine if CDDs/CDFs are present in sea breezes from remote islands in the South Pacific.

5. Determine if CDDs/CDFs can be formed by the combustion of fossil fuels in the presence of chlorine or inorganic chloride.

6. Determine if CDDs/CDFs can be detected in fish species taken from rivers remote from chemical manufacturing but close to incinerators and fossil-fueled power plants.

Although these studies were proposed in 1978, only items (3), (5) and (6) have even been partially addressed. Thus the "trace chemistries of fire" remains largely a working hypothesis that is in need of further testing and proving through well designed and conducted field sampling and laboratory research programs. Nevertheless, there exists some empirical evidence in this area.

Liberti et al. (1983) showed that CDDs/CDFs could be produced from the combustion of pure vegetable extracts in the presence of chlorine gas and oxygen. Pyrolytic degradation of extracts of chestnut, mimosa, and tannic acid was accomplished in a bench-scale thermal reactor. When combustion proceeded without chlorine gas, phenolic compounds and cresol were formed as primary thermolysis products. When the vegetable extracts were burned in association with chlorine gas or PVC plastic, chlorophenols and CDDs/CDFs were formed. Liberti et al. (1983) postulated that the PVC was acting as a chlorine donor in the formation of CDDs/CDFs from phenolic compounds, and that the chlorine gas directly formed the chlorinated precursor from a phenolic (pre-dioxin) ring structure. Table 3-25 summarizes these experiments.

There is some empirical evidence that the burning of wood, in the presence of chlorine or inorganic chlorides, may form CDDs/CDFs, although the evidence is not conclusive. Few of these experiments had ruled out contamination of the wood fiber by known chlorinated precursors through extraction and chemical analysis. None of the cited experiments attempted to determine if the wood fiber was contaminated by CDDs/CDFs prior to the conduct of the experiment. If the atmosphere serves to widely distribute CDDs/CDFs, and if CDDs/CDFs can exist in the vapor and particle phases in the ambient air, then trees and other biomass can become reservoirs of CDD/CDF contamination by means of particle deposition onto and vapor diffusion into the biomass. Until these possibilities have been addressed and their impacts, if any, are quantified, experiments in which CDDs/CDFs are generated from the combustion of wood must be interpreted with a certain degree of caution, especially with regard to proving that CDDs/CDFs can be formed in nature without human intervention.

Ahling and Lindskog (1982) demonstrated that the combustion of untreated wood in an open fire can generate relatively high concentrations of chlorinated aromatic hydrocarbons in the emissions. These compounds include established dioxin precursors such as di- through hexa-chlorobenzene and tetra- and penta- chlorinated phenols in ppb_v-ppm_v concentrations. In addition, ppm_v levels of benzene were produced. The presence of chlorinated precursors indicates that inorganic chlorides in the plant may be capable of chlorinating unsubstituted aromatic structures. Reaction kinetic experiments involving the formation of HCl vapor (Olie et al., 1983; Choudhry and Hutzinger, 1983) have shown that

HCl can be formed from inorganic chlorides as a result of a reaction between sulfur dioxide and sodium chloride during combustion, as follows:

_

2 NaCl + SO₂ + 1/2 O₂ + H₂O -----> Na₂SO₄ + HCl

or: HCl can be liberated by the catalytic reaction of NaCl and a metal oxide. The general reaction is:

_

NaCl + H₂O + (A) -----> (B) + HCl

where: (A) is a metal oxide, i.e.: Al₂O₃, Fe₂O₃.

Olie et al. (1983) conducted wood burning experiments in a bench-scale combustion unit. Wood treated with pentachlorophenol was incinerated to generate CDDs/CDFs in one experiment, and 60-year-old wood from the demolition of a residence was burned in a separate experiment. The authors alleged that the 60-year-old wood predated the manufacture and use of phenoxy wood preservatives and, therefore, probably was absent any dioxin precursors; however, this was not analytically confirmed. They did not directly monitor the smoke emissions for the presence of CDDs/CDFs, only the collected fly ash. CDDs/CDFs were detected in the fly ash in ppb_w concentrations. However, the authors noted that the quantified CDDs/CDFs could have occurred as a consequence of the previous tests of burning wood treated with pentachlorophenol.

Nestrick and Lamparski (1983) conducted studies on residential wood combustion to evaluate the possibility that CDDs may form. This was accomplished through the evaluation of soot scrapings from the chimneys of wood burning stoves. Samples were taken at random from the eastern, central, and western regions of the United States. Average total CDD levels in the chimney flue scrappings were: 8.3 ppb in the eastern region, 42.5 ppb in the central region, and 9.9 ppb in the west.

EPA tested a freestanding noncatalytic residential wood stove for chimney flue gas emissions of CDDs/CDFs during the combustion of pine and oak (U.S. EPA, 1987d). Through a series of tests, it was determined that the wood fiber was free of known chlorinated precursors (e.g., PCBs, chlorinated benzenes, and chlorinated phenols). The total chloride concentration was found to be 125 ppm for the oak and 49 ppm for the pine wood prior to burning in the wood stove. The combustion of the wood generated ppm levels of aromatic hydrocarbon compounds. This relatively high loading of emissions on the sampling device interfered with any speciation of CDD/CDF compounds in the emissions. However, combustion ash samples provided an alternative matrix for evaluation. The analysis of ash samples from the unit showed that only OCDD was present as a dioxin contaminant at a maximum concentration of 0.09 ppb (by weight). Wipe samples were also taken from inside the chimney flue. OCDD and hepta-CDD were detected in the chimney soot at a maximum concentration of 0.6 ppb and 0.04 ppb, respectively. No lower chlorinated CDDs nor any CDFs were found in the ash or soot wipes.

Choudhry and Hutzinger (1983) have postulated that the complex structure of lignin in wood fiber can pyrolyze to generate CDDs/CDFs if a chlorine donor is present. This theory is based on the experiments of Kirsbaum et al. (1972) in which two lignin preparations (spruce and asp) were thermally degraded in glass tubes at 475° C to yield an array of hydrocarbons, including phenol. If phenol could be formed, and if gaseous forms of inorganic or organic chlorine are available, then the phenol could be chlorinated to form a chlorophenol compound. The latter is then a precursor to the ultimate formation of CDDs/CDFs. In addition, Choudhry and Hutzinger (1983) indicated that continued pyrolysis of other hydrocarbons identified as thermolysis products in the combustion of lignin could ultimately yield benzene. Nestrick et al. (1987) demonstrated that benzene can react with an inorganic chloride in the presence of heat to produce a variety of chlorinated aromatic compounds including CDDs/CDFs. If these thermolytic pathways are operational in lignin pyrolysis, then, in theory, it is possible that forest fires can generate CDDs/CDFs in the smoke, which has been proposed by Clement and Tashiro (1991). Because of the potential importance of lignin pyrolysis as a potential, yet unverified, combustion source of CDDs/CDFs in the environment, additional research should be directed in this area. In the conduct of combustion experiments involving the pyrolysis of lignin, attention should be given to the identification of any CDDs/CDFs or precursor compounds that may exist as contaminants. Only after sample contamination has been completely ruled out can the researcher draw convincing conclusions from the experiment.

Coal is a naturally occurring substance having the potential to form CDDs when combusted. Mahle and Whiting (1980) first reported on the results of high temperature combustion of bituminous coal in a bench scale furnace with the addition of HCl, NaCl, or Cl_2­ and air to yield CDDs. Table 3-26 summarizes these experiments. In experiment III, tetra through octa-CDDs were formed from the oxidation of coal by air which had been bubbled through a solution of hydrochloric acid. In a review of this experiment, Choudhry and Hutzinger (1983) postulated that the hydrochloric acid aided in the chlorination of aromatic hydrocarbons produced as combustion byproducts. This was also the case in experiment IV in which chlorine gas was introduced into the oxidation of coal. Experiment IV produced the highest yield of tetra- through octa-CDDs. When coal was combusted only with air, the exothermic reaction did not generate detectable quantities of tetra- or

hexa- CDDs. Experiment I yielded only hepta- and octa-CDD in quantities close to the detection limit.

The air emission of polychlorinated biphenyls (PCBs) from MSW incinerators is less understood. There are virtually no theories explaining the detection of these compounds in incinerator emissions nor other combustion sources, the exception being the intentional destruction of PCBs in hazardous waste incinerators in which case 99.9999 percent destruction rated efficiency (DRE) must be achieved. When this occurs, 0.0001 percent of the initial amount of PCBs fed into the hazardous waste incinerator may be emitted out the stack. This may indicate that some small fraction of the PCBs present in the fuel fed into an incineration process may result in emissions of PCBs from the stack of the process.

PCBs have been measured as contaminants in the raw refuse prior to incineration in an MSWI (Choudhry and Hutzinger, 1983; Federal Register, 1991). It is possible to use this information to test Theory 1 involved in CDD/CDF emissions: that the PCB contamination present in the fuel is responsible for emissions from the stack. The mass balance of total PCB beginning with measurement in the raw refuse and ending with measurement at the stack to an RDF MSW incinerator (Federal Register, 1991) can be used to calculate the destruction rated efficiency (DRE) of incineration of the PCB contaminated MSW. Using results from test number 11 at the RDF facility (Federal Register, 1991), a computation of DRE can be made with the following equation (Brunner, 1984):

where:

W_i = mass rate of contaminant fed into the incinerator system

W_o = mass rate of contaminant exiting the incinerator system

In test 11, 811 nanograms of total PCBs/gram of refuse (ng/g) were measured in the MSW fed into the incineration system, and 9.52 ng/g of total PCB were measured at the inlet to the pollution control device (i.e., outside the furnace region, but preceding emission control). From these measurements, a DRE of 98.8 percent can be calculated. Therefore, it appears that PCB contamination in the raw MSW that was fed into this particular incinerator may have accounted for the emission of PCBs from the stack of the MSW incinerator.

PCBs can be thermolytically converted into CDFs (Choudhry and Hutzinger, 1983; U.S. EPA, 1984). This process occurs at temperatures somewhat lower than typically measured inside the firebox of an MSWI. Laboratory experiments conducted by EPA (U.S. EPA, 1984) indicate that the optimum conditions for CDF formation from PCBs are near a temperature of 675° C in the presence of 8 percent oxygen and a residence time of 0.8 seconds. This resulted in a 3 to 4 percent efficiency of conversion of PCBs into CDFs. Because 1 to 2 percent of the PCBs present in the raw refuse may survive the thermal stress imposed in the combustion zone to the incinerator (Federal Register, 1991), then it is reasonable to presume that PCBs in the MSW may contribute to the total mass of CDF emissions released from the stack of the incinerator. This is speculative, and more definitive research is needed in this area before strong conclusions can be made regarding the causes of PCB emissions during combustion.

3.5.5. Evaluation of Naturally Occurring CDD/CDFs by Examination of Sediment Core Data

In the review of these theories, a question arises as to the contribution made by combustion of synthetic organic substances produced by humans versus the contribution made by natural sources to the overall thermolytic synthesis of CDDs and CDFs. This question can be partially addressed using the results from analyses of the temporal distribution of CDDs/CDFs in sediment core samples taken from lakes located near the cradle of the U.S. industrial revolution (Czuczwa et al., 1984; Czuczwa and Hites, 1985; Czuczwa and Hites, 1986; Smith et al., 1992). Czuczwa and Hites (1985) analyzed sediment core samples taken in Lake Huron by the Great Lakes Research Station of the University of Michigan. Sedimentation rates within the core samples were determined by using Cs-137 and Pb-210 techniques of Robbins and Edgington (1973). These rates were used as a basis of relating depth of core sample to era. CDDs/CDFs were detected in the core samples, and the results showed no appreciable degradation of CDDs/CDFs in the sediments over time. The most abundant CDDs/CDFs were OCDDs, and HpCDDs/CDFs. Analysis of depth of core sample versus era showed that the CDDs/CDFs increased steadily in concentration beginning at about 1940 and leveled off at about 1960. Comparisons were made between this trend and the total production of synthetic chlorinated organic chemicals, as well as the total volume of coal combusted for energy production. If it is theoretically possible that the combustion of coal produces air emissions of CDDs/CDFs, then this should be reflected in the sediments. The pattern of levels of CDDs and CDFs in the sediment cores seemed to track the total volume production of synthetic chloro-aromatics by the petrochemical industry in the United States, whereas the consumption of coal did not show a good correlation. Czuczwa and Hites (1985) concluded that the history of sedimentation rates of CDDs/CDFs in core samples from Lake Huron were reflective of atmospheric deposition from the combustion of synthetic chloro-aromatics, and, therefore, could only have come from the combustion of anthropogenic substances.

In a separate study, Czuczwa et al. (1984) and Czuczwa and Hites (1986) reported on the temporal variability of CDDs/CDFs in sediment core samples taken from a wilderness lake located in an uninhabited/undeveloped island (Siskiwit Lake, Isle Royale) in Lake Superior. Comparisons were made between the congener profiles found in the lake sediments to congener profiles found in urban air particulates. A near perfect correlation was found (correlation coefficient = 0.998), leading to the observation that CDDs/CDFs entered the lake system from aerial transport and deposition. The historical record of CDD/CDF concentration in the core samples showed that CDDs/CDFs were virtually absent from the sediments until around 1940, therefore ruling out any significant contribution to background from natural sources such as forest fires.

Using a similar study design, Smith et al. (1992) investigated the temporal distribution by era in sediment core samples taken from Green Lake, New York, near Niagara Falls. Green Lake is a State Park, and removed from direct discharge of CDDs/CDFs into the system. The investigators found a similar congener profile as Czuczwa et al. (1984) and Czuczwa and Hites (1986) in the sediments, and found excellent agreement with measured deposition flux of OCDD into the lake and the concentration of OCDD in the most recent sediment layer. This supports the observation of Hites (1991) on the importance of atmospheric transport and deposition as a major pathway of entry of CDDs/CDFs into the aquatic environment. Smith et al. (1992) found that CDDs/CDFs could be detected in sediments dating back to 1860 - 1865, although concentrations were found to be low (e.g., CDDs = 7.0 ppt with 98 percent being OCDD; CDFs = 2.1 ppt with 75 percent being OCDF). These low concentrations remained essentially steady until about 1920 when concentrations significantly increased. Between 1920 and 1940, CDDs increased from about 1-10 ppt to about 250 ppt, and CDFs increased from 5 to about 100 ppt. Between 1940 and 1960, CDDs increased to about 680 ppt, and CDFs increased to about 300 ppt. From 1960 to 1980, CDDs continued to increase to approximately 950 ppt, whereas CDFs significantly declined to about 150 ppt. From the work of Czuczwa et al. (1984), Czuczwa and Hites (1986), and Smith et al. (1992), it appears that anthropogenic combustion sources, taken in their entirety, probably represent the largest mass flux of CDDs/CDFs into the environment, and that natural combustion activity (i.e., forest fires) probably is insignificant by comparison. For example, if it is assumed that the 7 ppt CDDs in sediments dating back to the 1860s is reflective more of natural sources, and if the CDDs that were found to be 950 ppt in 1980 are more reflective of human sources, then a simple comparison would indicate that anthropogenic sources may exceed natural sources by a factor of 100:1.

The above section discussed the likelihood that anthropogenic sources explain the bulk of CDD/CDFs currently in the environment. Still, the considerable research on the complex chemistries of combustion that transpire to ultimately yield CDDs/CDFs remains largely theoretical. The three primary theories being advanced are:

Theory 1: CDDs/CDFs present as contaminants in the combusted organic materials or that are thermally treated by a combustion process explain the emissions of CDDs/CDFs out of the stack. It is proposed that some quantity of this initial contamination survives the thermal stress imposed by the heat of the incineration or combustion process and is subsequently emitted from the stack.

Theory 2: CDDs/CDFs are ultimately formed from the thermal breakdown and molecular rearrangement of precursor compounds. Precursor compounds are chlorinated aromatic hydrocarbons having a structural resemblance to the CDD/CDF molecule. Among the precursors that have been identified are polychlorinated biphenyls (PCBs), chlorinated phenols (CPs), and chlorinated benzenes (CBs). The formation of CDDs/CDFs is believed to occur after the precursor has condensed and adsorbed onto binding sites on the surface of fly ash particles. The active sites on the surface of fly ash particles somehow promote the chemical reactions forming CDDs/CDFs as products of this reaction, which has been observed to be catalyzed by the presence of inorganic chlorides sorbed to the particulate. Heat in a range of 250-450° C has been identified as a necessary condition for these reactions to occur, with either lower or higher temperatures inhibiting the process. Therefore, the precursor theory focuses on the region of the combustor that is downstream and away from the high temperature zone of the furnace or combustion chamber. This is a location where the gases and smoke derived from combustion of the organic materials have cooled down because of heat losses during conduction through flue ducts; passing through heat exchanger and boiler tubes to recover the heat from combustion for the co-generation of energy; after passing through some air pollution control equipment, and while convected up the stack to be discharged to the atmosphere.

Theory 3: CDDs/CDFs are synthesized de novo in the same region of the combustion process as described in Theory 2 (i.e., the so-called cool zone). De novo refers to the formation of CDDs/CDFs from organic and inorganic substrates comprised of singular or mixtures of molecules bearing little resemblance to the molecular structure of CDDs and CDFs. In broad terms, these are nonprecursors and include such diverse substances as petroleum products, chlorinated plastics (PVC), nonchlorinated plastics (polystyrene), cellulose, lignin, coke, coal, particulate carbon, and hydrogen chloride gas. Formation of CDDs/CDFs requires the presence of a chlorine donor, a molecule that partakes a chlorine atom to the predioxin molecule, and the formation and chlorination of a chemical intermediate that is a precursor. The production of a chemical intermediate that is a dioxin precursor does not neatly differentiate Theory 2 from Theory 3, and indeed introduces some confusion into the explanation of the primary chemical events. The primary distinction is that Theory 3 begins with the combustion of diverse substances that are not defined as precursors, then the nonprecursor reacts on the carbonaceous particulate to eventually yield a chlorinated aromatic hydrocarbon that is a precursor to finally yield CDDs/CDFs. By this distinction, Theory 3 may be viewed as an augmentation to Theory 2 as explained by the initial steps in the synthesis of CDDs/CDFs.

This review has shown that all of these theories are possible and, therefore, taken as a whole, should not be seen as being mutually exclusive. One or more of these theories may act in combination during the combustion of carbonaceous matter. Theory 2 and Theory 3 require the presence of chlorine in the combustible material and in the gaseous state in the combustion plasma.

A summary of the major combustion sources that produce CDDs and CDFs is presented in the following sections. The development of combustor emission estimates has been coordinated with a similar effort ongoing in EPA's Office of Air Quality Planning and Standards (OAQPS). The OAQPS effort is part of a larger EPA effort to inventory air emissions of various toxic substances. To date, OAQPS has completed emission inventories for about 20 chemicals and has completed a draft document entitled "Locating and Estimating Air Emissions from Sources of Dioxins and Furans" (U.S. EPA, 1993a). This draft OAQPS document summarizes dioxin emissions data for a variety of combustor types. OAQPS is preparing a second report (draft not yet complete) entitled "Emissions Inventory of Section 112 (c)(6) Pollutants: 2,3,7,8-TCDD, 2,3,7,8-TCDF and 2,3,7,8-TCDD Toxic Equivalents." This report will combine the emission factor information in the "Locating and Estimating" report with production values to develop actual emission rate estimates. Since the OAQPS efforts will not be completed until after this draft report is issued, it is possible that the second OAQPS report may reach somewhat different conclusions. However, the Agency is striving to coordinate these efforts and make the outcomes as consistent as possible. Readers interested in further details about the OAQPS efforts are encouraged to contact OAQPS directly at their location in Research Triangle Park, NC.

Municipal Solid Waste Incinerators (MSWI) operating in the United States can be classified into four general design categories: mass burn, modular, refuse-derived fuel, and fluidized-bed (U.S. EPA, 1992h). The first type is called mass burn because the waste is combusted without any preprocessing other than removal of items too large to go through the feed system. In a typical mass burn combustor, refuse is placed on a grate that moves through the combustor. These facilities typically range in combustion capacity from 90 to 2,700 metric tons of MSW per day. Subcategories of mass burn technologies include refractory-walled, rotary kiln, and water-wall facilities. Refractory- walled represent an older class of MSWIs generally built in the late 1970's to early 1980's that were designed to reduce the volume of waste in need of disposal by 70 to 90 percent. These facilities generally lacked boilers to recover the heat of combustion for energy purposes. In the refractory design, the MSW is delivered to the combustion chamber by a traveling grate. Combustion air in excess of stoichiometric amounts is supplied both below and above the grate. Mass burn water-wall facilities represent substantial design improvements over the refractory-walled incinerators. The water-wall refers to a series of steel tubes running vertically along the walls of the furnace. The tubes contain water, which when heated by combustion, acts as a boiler and transfers energy to produce steam. The steam is then used either to drive an electrical turbine generator or for other industrial needs. Because a secondary purpose is to generate energy to sell to a customer, significant improvements over refractory-walled MSWIs in terms of increased combustion efficiency have been fostered. The third subcategory of mass burn MSWIs is the rotary kiln. The rotary kiln lacks a traveling or reciprocating grate system to deliver MSW into the furnace. Rather it employs a water-cooled rotary combustor that is essentially a rotating combustion barrel mounted at a slight angle of decline into which the refuse is pushed by a hydraulic ram (Donnelly, 1992). Preheated combustion air is delivered to the kiln through various portals. The slow rotation of the kiln (i.e., 10 to 20 rotations/hr) causes the MSW to tumble thereby exposing more surface area for complete burn-out. These systems are also equipped with boilers for energy recovery.

As with the mass burn type, modular incinerators also burn waste without preprocessing. Modular MSWIs consist of two combustion chambers (e.g., a primary and secondary chamber mounted in a vertical array). Modular combustors generally range in combustion capacity from 4.5 to 270 metric tons/day. One of the most common types of modular systems is the starved air (or controlled air system). In these systems, air is supplied to the primary chamber at sub-stoichiometric levels. The incomplete combustion products entrained in the combustion gases from the primary combustion chamber pass into the secondary combustion chamber where excess air is added and combustion is completed by elevated temperatures sustained by auxiliary fuel.

The third major type of MSWI technology is designed to combust refuse-derived fuel (RDF). RDF is a general term describing MSW from which relatively noncombustible items have been removed thereby enhancing the combustibility of the MSW. RDF is commonly prepared by shredding, sorting, and separating metals to create a dense MSW fuel in a pelletized form having a uniform size. RDF fuel is typically burned in a spreader stoker-type combustion chamber (Donnelly, 1992). In the United States, RDF facilities range in total combustion capacity from 227 to 2,720 metric tons/day. These MSWIs are typically steam production facilities that generate salable energy.

The fourth type of MSWI is the fluidized-bed design. In this design, the waste burns in a turbulent bed of noncombustible material, usually sand. The MSW may be fed into the incinerator either as unprocessed waste or as a form of RDF. There are two basic design concepts to the technology: (1) a bubbling-bed incineration unit and (2) a circulating-bed incineration unit. Fluidized-bed MSWIs typically have capacities ranging from 184 to 920 metric tons/day. These systems are usually equipped with boilers to produce steam.

Currently, there are about 170 to 190 MSWI facilities located in 37 states in the United States. (Berenyi and Gould, 1993; Burton and Kiser, 1993). This range in number of facilities reflects the fact that, for any given point in time, the exact population of operating facilities is unknown. However, the best estimate is that 171 MSWI facilities are in operation (Berenyi, 1993; Berenyi and Gould, 1993). About one-half of the operating MSWIs were built since 1988 (Berenyi and Gould, 1993). The states with the greatest number of facilities are: New York (16), Florida (14), Minnesota (14), Massachusetts (8), Virginia (8), and Connecticut (7) (Berenyi and Gould, 1993). In the most recent reporting year, 1991, EPA estimated that approximately 29.35 million metric tons of MSW were combusted by all operating MSWIs; this represents approximately 17 percent of the annual generation of MSW in the United States (U.S. EPA, 1992c).

Gould (1991) estimated the average annual utilization capacity of typical MSWI designs. Utilization capacity is defined as the percentage of days a facility operates during the course of the year (U.S. EPA, 1992h). Gould (1991) estimated that existing mass burn, modular, and RDF MSWIs had average annual utilization capacities of 87.5, 84.2, and 83.3, respectively.

An estimated 85 percent of existing MSWIs are equipped with one or more air pollution control devices (APCD) to remove some class of pollutants prior to release from the stack (e.g., particulate matter, heavy metals, acid gases, and/or organic constituents) (U.S. EPA, 1992h). These APCDs include electrostatic precipitators (ESPs), fabric filters (FFs), dry sorbent injection (DSI), spray dryer adsorption (SDA), and wet scrubbers (WS). The ESP is generally used to collect and control particulate matter derived from combustion. This is accomplished by introducing a strong electrical field in the flue gas stream, which, in turn, imparts a charge to the particles entrained in the combustion gases (Donnelly, 1992). Large collection plates are given an opposite charge to attract and collect the particles. Fabric filters are also particulate matter control devices. Six- to eight-inch diameter bags made from woven fiberglass material are arranged in series. The combustion gases are forced through the tightly woven fabric. The porosity of the fabric is such that the bags act as a filter medium and retain small particles comprising the particulate matter. Dry sorbent injection is designed for the control of MSWI acid gases. DSI involves the injection of hydrated lime or soda ash into the gas stream to react with and neutralize the acid gas constituents (Donnelly, 1992). Spray dryer adsorption involves both acid gas and particulate matter control. In a typical SDA system, hot combustion gases enter a reactor where atomized hydrated lime slurry is introduced at a controlled velocity (Donnelly, 1992). The flue gas temperature is significantly decreased, and the acid gas constituents quickly react with the reagent. The reaction evaporates the moisture to produce a dried product that is removed from the bottom of the spray dryer. In general, SDAs are used in combination with either ESPs or FFs. Greater than 95 percent reduction and control of CDDs/CDFs in MSWI emissions has routinely been achieved with FF/SDA systems (U.S. EPA, 1992h). Wet scrubber devices (WS) are designed for acid gas removal, and are more common to MSWIs in Europe than in the United States. Wet scrubber devices consist of two-stage scrubbers whereby the first stage removes HCl and the second stage removes SO₂ (Donnelly, 1992). Water is used to remove the HCl, and either caustic or hydrated lime is added to remove SO₂ from the combustion gases. Table 3-27 summarizes the current estimated distribution of operating MSWIs by design category and installed APCDs.

Estimation of MSWI Dioxin Emissions Using an Emission Factor Approach The approach used here to estimate emissions is based on an emission factor. Emission factors are estimates of the mass of CDD/CDF emitted from the stack per kg of waste combusted. As shown in Table 3-28, these factors were estimated for each design category, multiplied by the amount of waste burned within the design category and then summed to get the total emissions.

The first step in this process is to collect emission test data representative of each design category. EPA's Office of Air Quality Planning and Standards (OAQPS) has already collected such a data set (U.S. EPA, 1993a). This summary presents emission testing for dioxin-like compounds for 30 existing MSWI facilities. These tests have been reviewed by OAQPS and determined to have used appropriate stack testing and laboratory protocols and to have been conducted under normal operating conditions. These 30 facilities represent a mix of MSWI designs and technologies as well as air pollution control devices (APCDs) in actual use, providing a basis for extrapolating to all U.S. facilities. For design

categories where more than one test was available, the concentrations for each congener were averaged across tests.

The emissions data (U.S. EPA, 1993a) are presented in concentration units of nanogram of CDD/CDF per dry standard cubic meter of combustion gas (ng/dscm) corrected to 7% oxygen. Emission factors were computed for each MSWI design category by multiplying the average CDD/CDF concentration by the volume of combustion gas that is produced per kg of waste incinerated. The gas production factor was derived considering the typical heat content of the refuse as follows (Federal Register, 1987c):

1. Assume the heat content of typical MSW = 4500 B.t.u./lb of MSW.

2. Assume that 2.57E-7 dscm are produced per joule value of the MSW.

3. One joule = 9.47E-04 B.t.u.

4. One pound = 0.4536 kg

Then:

dscm/kg of MSW = (4500 B.t.u./lb) x (1 joule/9.47E-04 B.t.u) x (lb/.4536 kg) x (2.57E-07 dscm/joule)

dscm/kg of MSW = 2.69

As indicated above an emission factor was estimated for each design class and multiplied by the amount of waste burned to get the emission rate. The emission rate estimates shown in Table 3-28 reflect all congeners of CDD and CDF. These values can be converted to TEQs by applying a ratio of total CDD/CDF to TEQ. EPA has reviewed the congener-specific emissions profiles of twelve MSWI technologies and has determined that, although variable, the average ratio appears to be about 60:1 (i.e., the total mass of CDD/CDF is roughly 60 times greater than the computed TEQ) (Radian, 1994). As a measure of variability, the standard deviation from this analysis was +/- 20 from the mean ratio (i.e., a ratio of 40:1 to 80:1). As noted in Table 3-28, the total CDD/CDF mass emission from all operating 171 MSWIs is 1.8E+05 grams. The TEQ mass emission is estimated to be 3,000 grams TEQ/yr (assuming the average ratio of 60:1 for the TEQ conversion from the total CDDs/CDFs).

The procedure used to estimate national emissions of dioxin from the MSWI industry involves uncertainties that could cause the estimate to be lower or higher than the true value. The emission estimates were derived on the basis of emissions testing from 30 facilities. As discussed below much of the uncertainty revolves around the representativeness of these facilities.

• How well do the 30 facilities represent the whole population of 171 facilities in terms of technologies? The 30 facilities were selected to be representative of the range of MSWI designs and air pollution control systems. As indicated in Table 3-28, only one of the 13 design classes was not represented. Facilities of particular concern are those that use ESPs which operate in a temperature range of 200° - 400° C. As discussed in Section 3.5 these conditions can promote the formation of CDDs/CDFs. Over the past few years, some of the facilities with "hot-sided ESPs" have made changes in operating conditions or equipment to address this problem. Although, the 30 tested facilities do include some with "hot-sided ESPs" it is not clear if they are representative of current conditions at all such facilities.

• How well do the 30 facilities represent the whole population of 171 facilities in terms of timing? The emissions were largely derived from stack tests conducted during the period 1988 to 1991. Since 1991, new facilities may have become operational or changes may have been made to existing ones. Therefore emissions today may be somewhat lower, reflecting continued improvements in combustor design.

• For individual facilities, how representative are emission tests of long term performance? The average emissions from a single facility are typically derived from 3 - 4 days of testing over the year. It is not known to what extent such short-term testing may truly reflect long-term emissions, e.g., through the life of the facility. Most stack testing data were collected while the MSWI was operating according to design specifications, e.g., under normal operating conditions. Using these data would not reflect any additional emissions that may occur during upsets in the combustion zone, poor operations, equipment malfunctions, or degradation in the effectiveness of the pollution control systems.

• How accurate is the approach used to convert stack concentrations to emission factors? As discussed earlier in this section, the approach used to convert concentration in the combustion gas to an emission factor is based on an assumption that 2.69 dscm of combustion gas are produced per kg of MSW burned. This quantity is variable among facilities and is dependent on such factors as the temperature of combustion, the amount of air supplied to combustion chamber in excess of stoichiometric requirements, the moisture of the feed material being burned, and the heat value of the feed material being combusted. For some technologies with relatively high amounts of excess air delivered to the combustion chamber, the gas volume may be as high as 5.0 - 6.0 dscm/kg.

• How accurate is the procedure used to convert total CDD/CDF emissions to Toxic Equivalents (TEQs)? The conversion ratio was based on a review of emissions from 12 MSWIs. In actuality, the ratio of total CDD/CDF to TEQ is variable from one facility to another. It is influenced by the composition of the MSW and the operating conditions of the combustor. It is not known how representative the generic ratio of 60:1 is of dioxin emissions from all existing MSWIs.

Although MSWIs have the strongest emission data base of all combustion sources evaluated in this document, it still must be considered uncertain for the reasons stated above. Therefore, the estimated emission factors are given a "medium" confidence rating. The amount of MSW that is annually combusted by various MSWI technologies (see Table 3-28) is given a "high" confidence rating. These estimates are based on a recently conducted and comprehensive survey (U.S. EPA, 1992h). Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual emissions (3,000 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 1,300 to 6,700 g TEQ/yr.

EPA will soon propose revised emission standards for all existing and new MSWIs with unit capacities greater than 30 metric tons per day. Once these standards have been promulgated, and the States have fully enforced the emission limits, then EPA expects to reduce the national emissions of dioxin emissions from all existing MSWIs by about 99 %. All existing facilities combined should then be emitting about 30 g TEQ/yr. Full implementation and enforcement of the rules should be achieved by the year 2000. As the compliance date approaches and facilities are upgraded, EPA expects that emissions from these facilities will decline significantly from current levels.

An estimated 7 million metric tons of total ash (bottom ash plus fly ash) are generated annually by MSWIs (telephone conversation between J. Loundsberry, U.S. EPA Office of Solid Waste, and L. Brown, Versar Inc., on February 24, 1993). U.S. EPA (1991b) indicates that 2.8 to 5.5 million tons of total ash are produced from MSWIs with fly ash comprising 5 to 15 percent of the total. U.S. EPA (1990c) recently reported the results of analyses of MSWI ash samples for CDDs and CDFs. Ashes from five state-of-the-art facilities located in different regions of the United States were analyzed for all 2,3,7,8-substituted CDDs and CDFs. The TEQ levels in the ash (fly ash mixed with bottom ash) ranged from 106 ng/kg to 466 ng/kg with a mean value of 258 ng/kg. CDD/CDF levels in fly ash are generally much higher than in bottom ash. For example, Fiedler and Hutzinger (1992) report levels of 13,000 ng TEQ/kg in fly ash. Multiplying the mean TEQ total ash concentration by the estimated volume of MSWI ash generated annually (7 million metric tons) yields an estimated annual TEQ in MSWI ash of 1,800 g TEQ/yr.

The total ash generation estimate is given a "medium" confidence rating since it is based on an expert opinion and is about twice as high as earlier published estimates. The emission factor is given a "medium" confidence rating because it is based on direct measurements at five facilities, although these five facilities may not be representative of all technologies in the United States. Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual emissions (1,800 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 810 to 4,000 g TEQ/yr.

Each of the five facilities sampled in U.S. EPA (1990c) had companion ash disposal facilities equipped with leachate collection systems or some means of collecting leachate samples. Leachate samples were collected and analyzed for each of these systems. Detectable levels were only found in the leachate at one facility (TEQ = 3 ng/l); the only detectable congeners were HpCDDs, OCDD, and HpCDF.

EPA estimates that there are 190 Hazardous Waste Incinerators (HWIs) in the United States. This total includes both operating facilities and facilities that are not operating but have filed an application with EPA (Helble, 1993). The four principal technologies employed for the combustion of hazardous waste in the United States are: liquid injection, rotary kiln, fixed hearth and fluidized-bed incinerators (Dempsey and Oppelt, 1993). Liquid injection incinerators are designed to burn pumpable liquid hazardous waste. These incinerators are typically simple refractory-lined cylinders (either horizontally or vertically aligned) equipped with one or more waste burners. The liquid waste is injected into the combustion chamber through an atomizer, and the liquid droplets are exposed to high temperatures in suspension. Rotary kiln incinerators are the more common design. They have the added versatility of being able to combust hazardous waste in any physical form (i.e., liquid, semi-solid, or solid). The rotary kiln is a horizontal cylinder lined with refractory material. Rotation of the cylinder on a slight slope provides for transport of the waste through the kiln, as well as enhanced mixing and exposure to the heat of combustion. The combustion gases emanating from the kiln are usually passed through a high temperature afterburner chamber to more completely destroy organic pollutants arising from combustion. Fixed hearths, the third principal hazardous waste incineration technology, are starved air or pyrolytic incinerators. These are two-stage combustion units. Waste is ram-fed into the primary chamber and incinerated at about 50 to 80 percent of stoichiometric air requirements. The resulting smoke and pyrolytic combustion products are then passed though a secondary combustion chamber where relatively high temperatures are maintained by the combustion of auxiliary fuel. Oxygen is introduced into the secondary chamber to promote complete thermal oxidation of the organic molecules entrained in the gases. The fourth hazardous waste incineration technology is the fluidized-bed incinerator. This technology is similar in design to that employed in MSW incineration. (See Section 3.6.1).

Dempsey and Oppelt (1993) summarized the results of EPA-sponsored stack testing at six full-scale HWIs, three PCB incinerators, and one incinerator burning PCP waste. CDD/CDFs were detected at all three PCB incinerators with TEQ emission rates ranging from 0.3 to 1.63 ng TEQ/dscm (@ 7 percent oxygen). CDD/CDFs were detected at three of the HWIs with TEQ emission rates ranging from 0.57 to 17.7 ng TEQ/dscm (@ 7 percent oxygen).

Helble (1993) reviewed recent data from trial burn reports on CDD/CDF emissions from 15 HWIs. CDD/CDFs were detected in the stack emissions of 11 of the 15 facilities at total CDD/CDF emission rates ranging from 0.1 to 1,600 ng/dscm (@ 7 percent oxygen) with most facilities between 1 and 100 ng/dscm. Based on his evaluation of the emissions data, Helble (1993) concluded that the CDD/CDFs observed in emissions from HWIs are formed catalytically under low temperature conditions either through catalytic chlorination or through catalytic condensation of dioxin-like precursors such as chlorobenzenes and PCBs.

Emission factors are estimated based on the results of the emission tests reported by Helble (1993). Homologue-specific emissions data, waste feed rates, and stack flow rates (dscm @ 7 percent oxygen) were available for six of the HWIs evaluated by Helble (1993). From these data, total CDD/CDF emission factors were calculated for each facility (range: 10 to 6,830 ng/kg of waste feed; mean: 1,550 ng/kg of waste feed). For those facilities with more than one test run reported, the total CDD/CDF emission rates for the individual runs were averaged to obtain a facility average emission rate. These total CDD/CDF emission factors were converted to TEQ emission factors using a conversion factor of 1.75 ng TEQ/ng of total CDD/CDF that was developed by EER, Inc. for EPA's Office of Solid Waste (EER, 1993). The resulting mean TEQ emission factor is 27.2 ng TEQ/kg waste feed (range = 0.18 to 119 ng TEQ/kg waste feed).

Dempsey and Oppelt (1993) estimate that between 216 and 249 million metric tons of hazardous waste were generated in 1987 (the year for which the most comprehensive data on waste management are available). Of this total amount, Dempsey and Oppelt (1993) estimate that between 1.0 and 1.3 million metric tons of hazardous waste were incinerated. Based on an estimated 1.3 million metric tons of hazardous waste incinerated per year in the United States and the mean emission factor derived above, it is estimated that 2,000 grams CDD/CDF per year and 35 grams TEQ/yr are emitted from HWIs.

A "low" confidence rating is ascribed to the emission factors derived above because stack test data were available for only 6 of the 190 HWIs in the United States and the stack test data used represent only one hazardous waste technology (rotary kiln). The "production" estimate has been assigned a "medium" confidence rating because it is based on a thorough review of the various studies and surveys which have been conducted in recent years to assess the quantity and types of hazardous waste generated in the United States, as well as the quantities and types of wastes being managed by various treatment, storage and disposal facilities. Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual TEQ emissions (35 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 11 to 110 g TEQ/yr.

Buonicore (1992b) has reviewed the primary incineration technologies used to burn medical and pathological wastes in the United States. These medical waste incinerations (MWI) fall within three broad technology categories: retort, controlled-air, and rotary kiln. Retort incinerators are multiple chamber combustors characteristic of the "older" existing technology. The medical waste is fed into a primary combustion chamber, and the gases from combustion are passed into a secondary chamber. In the secondary chamber, secondary auxiliary fuel is burned to sustain higher temperatures and to more completely burn the organic pollutants entrained in the combustion gas from primary combustion. Combustion air, 100 to 300 percent in excess of stoichiometric requirements, is usually added to the secondary chamber. Gases exiting the secondary chamber are directed to an incinerator stack. The second principal technology is the controlled-air incinerator. This is the most common technology used to incinerate hospital/medical waste and is often referred to as modular incineration. Like retort incinerators, combustion occurs in two stages. Medical waste is fed into a primary combustion chamber where air is delivered at less than stoichiometric requirements. Under these conditions, the waste is pyrolyzed and volatile compounds are released. A secondary chamber is located on top of the primary unit. Auxiliary fuel is added to sustain high temperatures in a controlled-air environment. These systems are usually automated with computer-directed controllers that are integrated with a thermocouple. Thus, the quality of combustion is superior to the retort technology. The third type of MWI is the rotary kiln. This is the same technology as employed in both municipal and hazardous waste incineration. (See Sections 3.6.1 and 3.6.2).

EPA has estimated that about 4.3 million metric tons (4.76 million short tons) of hospital/medical wastes are generated annually in the United States (U.S. EPA, 1991d). Table 3-29 summarizes the types and number of facilities that generate medical waste, and their corresponding annual generation rate of medical wastes. There are about 6,700 MWIs operating nationwide combusting approximately 3.72 million metric tons of medical waste annually (U.S. EPA, 1991d). Table 3-30 summarizes the estimated population of MWI currently operating in the United States.

CDDs and CDFs have been identified in the stack gas emissions of MWIs located at hospitals in the United States (U.S. EPA, 1993a). Although operating on a smaller-scale, the mechanism of CCD/CDF formation in hospital waste incineration is similar to that described for MSWI in Section 3.6.1. To support future rulemaking, EPA has developed a summary of annual emissions of dioxin from all existing hospital waste incinerators operating in the United States (U.S. EPA, 1991e). This summary represents an analysis and review of dioxin emissions measured at the stack from six MWIs (Radian, 1991a; 1991b; 1991c; McCormack, 1990; Lew et al., 1988; Lew et al., 1989). From these reports, EPA derived an average emission factor of total amount of dioxin released to the air per kg of medical waste combusted in a typical MWI (g total CDD/CDF per kg waste), based largely on uncontrolled emissions (U.S. EPA, 1993a). The average uncontrolled emission factor of total CDD/CDF is 8.53E-05 g/kg (U.S. EPA, 1993a). This factor can be compared with the average controlled emission factor of total CDD/CDF of 4.46E-06 g/kg from one facility equipped with acid gas controls and a fabric filter (U.S. EPA, 1993a). The ratio of controlled to uncontrolled emissions is a factor of 1:20. Table 3-31 summarizes the emission factors developed for this analysis.

In computing an estimate of national emissions of dioxin from 6,700 existing MWIs, EPA applied an average emission factor developed for uncontrolled MWIs. Uncontrolled emissions are defined as emissions from a MWI facility not equipped with add-on air pollution control devices (APCD) (e.g., electrostatic precipitator, scrubber, fabric filter, etc.). However, MWIs are modular designs consisting of both a primary and secondary combustion chamber. The purpose of the secondary combustion chamber is the continued destruction of organic compounds emanating from the primary chamber. Therefore MWIs are not actually totally uncontrolled. EPA expects that the majority of the existing MWIs are uncontrolled with respect to dioxin control measures (U.S. EPA, 1991e). EPA believes that the selection of an average emission factor derived from uncontrolled emissions

currently represents the most accurate means of estimating the magnitude of potential dioxin release from all 6,700 operating MWIs (U.S. EPA, 1991e).

In order to estimate national emission of total dioxin to the air from all operating facilities, EPA categorized the population of MWIs according to the operating duty, the size of the combustor, and the amount of medical or pathological waste combusted per year. Table 3-32 summarizes the estimate of total dioxin emitted (g/yr) from all operating MWIs in the United States according to this disaggregation. Emissions to air of total CDD/CDF (i.e., tetra-chlorinated through octa-chlorinated compounds) from approximately 6,700 existing medical waste incinerators are estimated to be 3.18E+05 grams/yr. This emission estimate was derived from tests conducted at six facilities (considered to be representative of the major design types), extrapolating average emissions nationwide using the amount of waste burned in each of the design classes.

For purposes of deriving an estimate of emissions in terms of TEQ, it is necessary to convert the total CDD/CDF emission into an estimated TEQ emission. This is done by assuming a ratio of TEQ to total CDD/CDF using existing data on emissions from existing facilities. The State of California Air Resources Board (CARB) has stack tested a number of hospital waste incinerators in southern California (CARB, 1990a). Congener-specific emissions of CDD/CDFs were measured in the stack gas emissions of seven facilities. From these data, the ratio of TEQ to total CDD/CDF is 0.016 as an overall average of five tested facilities. Multiplication of the annual emissions of CDD/CDF (in grams per year) by this ratio yields an estimate of 5,100 g TEQ emitted (grams per year) for all existing MWIs in the United States.

U.S. EPA (1993a) reports emissions testing at a number of controlled-air medical waste incinerators with a variety of emission controls. These tests yielded a lower range of emission factors. Based on these data, it appears possible that the national releases from medical waste incinerators could be much lower than the "average" value identified above. It is difficult to say how much lower, since it is unknown how representative these tested facilities are of all 6,700 facilities in the United States.

A "medium" confidence rating is assigned to the estimate of amount of hospital waste burned since it is based on a detailed study specific to the United States; however, the large number of these facilities makes it difficult to estimate precisely. The emission factor used to extrapolate to a national basis is given a "low" confidence rating, because the average was derived from the stack sampling at a small sample of the large numbers of MWI facilities (6 of 6,700). Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (5,100 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 1,600 to 16,000 g TEQ/yr.

In 1987, the U.S. EPA stack tested three kraft black liquor recovery boilers for the emission of dioxin in conjunction with the National Dioxin Study (U.S. EPA, 1987). These boilers are associated with the production of pulp in the making of paper using the Kraft process. In this process, wood chips are cooked in large vertical vessels called digesters at elevated temperatures and pressures in an aqueous solution of sodium hydroxide and sodium sulfide (Someshwar and Pinkerton, 1992). Wood is broken down into two phases: a soluble phase containing primarily lignin, and an insoluble phases containing the pulp. The spent liquor (called black liquor) from the digester contains sodium sulfate and sodium sulfide that the industry finds beneficial in recovering for reuse in the Kraft process. In the recovery of black liquor chemicals, weak black liquor is first concentrated in multiple-effect evaporators to about 65 percent solids. The concentrated black liquor also contains 0.5 to 4 percent chlorides by weight (U.S. EPA, 1987). Recovery of beneficial chemicals is accomplished through combustion in a Kraft black liquor recovery furnace. The concentrated black liquor is sprayed into a furnace equipped with a heat recovery boiler. The bulk of the inorganic molten smelt that forms in the bottom of the furnace contains sodium carbonate and sodium sulfide in a ratio of about 3:1 (Someshwar and Pinkerton, 1992). The combustion gas is usually passed through an electrostatic precipitator that collects particulate matter prior to being vented out the stack. The particulate matter can be processed to further recover and recycle sodium sulfate.

The three sites that were stack tested by EPA (U.S. EPA, 1987) were judged to be typical of Kraft black liquor recovery boilers. The following emission factors of dioxin were derived from the stack emissions data: average CDD/CDF = 9.34E-03 µg/kg (range: 4.88E-03 to 1.67 E-02 µg/kg), and average TEQ = 9.71E-05 µg/kg (range: 3.33E-05 to 2.06E-04 µg/kg). A "medium" confidence rating is ascribed to these emission factors because the emission factors were derived from the stack testing of three Kraft black liquor recovery boilers that were judged to be fairly representative of technologies used at Kraft pulp mills in the U.S.

In 1989, EPA estimated that approximately 28.2 million metric tons of black liquor solids were burned in Kraft black liquor recovery boilers in the U.S. (U.S. EPA, 1992g). This production estimate was given a confidence rating of "high" because it is based on a recent industry-wide survey conducted by EPA. Assuming this is the quantity of black liquor that is combusted each year, then it is estimated that 264 grams of CDD/CDF and 2.7 grams of TEQ are emitted to the U.S. atmosphere annually. Based on the confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual TEQ emissions (2.7 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 0.9 to 4.3 g TEQ/yr.

As discussed in Section 3.2, approximately 500 million dry kg of pulp and paper mill wastewater sludge were incinerated in 1990 by facilities employing chlorine bleaching of pulp (U.S. EPA, 1993e). However, insufficient data are currently available to estimate emission factors for dioxin-like compounds from pulp and paper mill incinerators. As discussed in Section 3.2, EPA proposed control technology standards that address CDD/CDF emissions for non-combustion pulp and paper mill sources in December 1993 (Federal Register, 1993a) and will propose control technology standards for combustion sources by October 1994 (U.S. EPA, 1992d).

Brunner (1992) reviewed the four principal combustion technologies used to incinerate sewage sludge in the U.S.: multiple-hearth incinerator, fluidized-bed incinerator, electric furnace, and cyclone furnace. All of these technologies are "excess-air" processes (i.e., they combust sewage sludge with oxygen in excess of theoretical requirements). Of the four types of technologies, multiple-hearth incinerators are the most common. They constitute approximately 60 percent of the 199 existing sewage sludge incineration facilities operational in the U.S. (Federal Register, 1993b). The furnace consists of refractory hearths arranged vertically in series, one on top of the other. Dried sludge cake is fed to the top hearth of the furnace. The sludge is mechanically moved from one hearth to another through the length of the furnace. Moisture is evaporated from the sludge cake in the upper hearths of the furnace. The center hearths are the burning zone to the furnace where gas temperatures reach 871° C. The bottom hearths are the burn-out zone where the sludge solids become ash. A waste-heat boiler is usually included in the burning zone where steam is produced to provide supplemental energy at the sewage treatment plant. Air pollution control measures typically include a wet scrubber system for particulate matter control (U.S. EPA, 1987).

The fluidized-bed incinerator is a cylindrical refractory-lined shell with a steel plate structure that supports a sand bed near the bottom of the furnace (Brunner, 1992). Air is introduced through openings in the bed plate supporting the sand. This causes the sand bed to undulate in a turbulent air flow, hence the sand appears to have a fluid motion when observed through furnace portals. Sludge cake is added to the furnace at a position just above this fluid motion of the sand bed. The fluid motion promotes mixing in the combustion zone. Sludge ash exists the furnace with the combustion gases, therefore air pollution control systems typically consist of high-energy venturi scrubbers.

Electric furnaces are sometimes called infrared furnaces (Brunner, 1992). This incineration system consists of a long rectangular refractory-lined chamber. A belt conveyer system moves the sludge cake through the length of the furnace. To promote combustion of the sludge, supplemental heat is added by electric infrared heating elements within the furnace that are located just above the travelling belt. Electric power is required to initiate and sustain combustion.

Cyclonic furnaces consist of a refractory-lined cylindrical shell with a domed top (Brunner, 1992). Air is blown in at tangential burner ports on the furnace shell which causes a violent swirling pattern. This motion promotes good mixing of combustion air with the sludge feed. Sludge is fed into the furnace chamber by screw conveyor. Combustion gases exit at the top of the swirling vortex at the top of the furnace dome.

EPA has confirmed that dioxin can be emitted from sewage sludge incineration based on the testing of three multiple-hearth sewage sludge incinerators (U.S. EPA, 1987). Emission factors for dioxin were developed from these data. The average emission factor of CDD/CDF was estimated to be 1.26E+00 µg/kg of dry sewage sludge (range: 8.80E-02 to 3.37E+00 µg/kg). The average emission factor for TEQ was estimated to be 2.69E-02 µg/kg of dry sewage sludge (range: 1.17E-03 to 3.04E-02 µg/kg) assuming perfect congener distribution within the total CDD/CDFs measured.

In 1992, approximately 199 sewage sludge incineration facilities combusted about 0.865 million metric tons of dry sewage sludge (Federal Register, 1993b). Given this mass of sewage sludge incinerated/yr, the best estimates of emissions to air are 1,090 grams of CDD/CDF per year and 23 grams TEQ per year from all sewage sludge incineration facilities.

A "medium" confidence rating is ascribed to the emission factors because they were developed from the stack testing of three multiple hearth incinerators. Although multiple hearth incinerators are the dominant technology in use in the U.S. today, some uncertainty exists as to the representativeness of these derived emission estimates to possible emissions from other sewage sludge incineration technologies. The production estimate is assigned a "high" confidence rating because it is based on an extensive EPA survey to support rulemaking activities. Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual emissions (23 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 10 to 52 g TEQ/yr.

Nonferrous metals include aluminum, copper, nickel and magnesium. Insufficient information is available for evaluating CDD/CDF emissions, if any, from primary smelting/refining of nonferrous metals in the United States. However, several European investigators have investigated the presence of CDD/CDFs at some facilities in this industry.

Oehme et al. (1989) reported that the production of magnesium leads to the formation of CDDs and CDFs. Oehme et al. (1989) estimated that 500 g of TEQ are released in wastewater to the environment and 6 g TEQ are released to air annually from a magnesium production facility studied in Norway; CDFs predominated with a CDF to CDD concentration ratio of 10 to one. The magnesium production process involves a step in which MgCl₂ is produced by heating MgO/coke pellets in a pure chlorine atmosphere to about 700 to 800° C. The MgCl₂ is then electrolyzed to metallic magnesium and Cl₂. The Cl₂ excess from the MgCl₂ process and the Cl₂ formed during electrolysis is collected by water scrubbers and discharged to the environment.

Oehme et al. (1989) also report that certain primary nickel refining processes generate CDDs and CDFs, primarily CDFs. Although the current low temperature process used at the Norwegian facility studied is estimated to release only 1 g TEQ per year, a high temperature NiCl₂/NiO conversion process that had been used for 17 years at the facility is believed to have resulted in much more significant releases based on the ppb levels of CDFs detected in aquatic sediments downstream of the facility (Oehme et al., 1989).

Lexen et al. (1993) reported that samples of filter powder and sludge from a lagoon at the only primary aluminum production plant in Sweden showed no or little CDD/CDF.

Secondary smelters/refiners are establishments primarily engaged in the recovery of nonferrous metals and alloys from new and used scrap and dross. The principal metals of this industry both in terms of volume and value of product shipments are aluminum, copper, lead, zinc, and precious metals (U.S. DOC, 1990a). Scrap metal and metal wastes may contain organic impurities such as plastics, paints, and solvents. Secondary smelting/refining processes for some metals (e.g., aluminum, copper, and magnesium) utilize chemicals such as NaCl, KCl, and other salts. The combustion of these impurities and chlorine salts in the presence of various types of metal during reclamation processes can result in the formation of CDDs and CDFs as evidenced by the detection of CDDs and CDFs in the stack emissions of secondary aluminum, copper, and lead smelters (Aittola et al., 1992; U.S. EPA, 1987; 1994b; 1994c; 1994d).

Levels of 2,3,7,8-TCDF in stack gas from an aluminum reclamation facility in the Finnish city of Vyborg have been measured at approximately 43 ng/m³ (Aittola et al., 1992). However, no studies of CDD/CDF emissions from secondary aluminum smelters located in the United States have been reported. Aluminum is processed at more smelters than any other nonferrous metal in the United States. Also more aluminum undergoes secondary smelting than any other nonferrous metal. An estimated 1.7 million metric tons of aluminum were produced by secondary smelters in 1987 in the United States (U.S. DOC, 1990a).

Stack emissions of CDD/CDFs from a secondary copper smelter were measured by EPA during the National Dioxin Study (U.S. EPA, 1987). The tested facility recovers copper and precious metals from copper and iron-bearing scrap. The copper and iron-bearing scrap are fed in batches to a cupola blast furnace, which produces a mixture of slag and black copper. Four to five tons of metal-bearing scrap were fed to the furnace per charge, with materials typically being charged 10 to 12 times per hour. Coke was used to fuel the furnace, and represented approximately 14 percent by weight of the total feed. During the stack tests, the feed consisted of electronic telephone scrap and other plastic scrap, brass and copper shot, iron-bearing copper scrap, precious metals, copper bearing residues, refinery by-products, converter furnace slag, anode furnace slag, and metallic floor cleaning material. Oxygen enriched combustion air for combustion of the coke was blown through tuyeres at the bottom of the furnace. At the top of the blast furnace were four natural gas-fired afterburners to aid in completing combustion of the exhaust gases. Particulate emissions were controlled by fabric filters, and the flue gas then was discharged into a common stack. The estimated emission factors derived for this one site are: CDD/CDF = 3.89E+04 ng/kg of scrap metal smelted (range: 3.31E+04 to 4.05E+04 ng/kg); TEQ = 7.79E+02 ng/kg of scrap metal smelted (range: 7.64E+02 to 1.04E+03 ng/kg).

More than 0.3 million metric tons of copper were produced by the 24 secondary copper smelters operating in the United States in 1987 (U.S. DOC, 1990a). If the emission rates derived above are assumed to be representative of all secondary copper smelters, then the best estimate of annual air emission of CDD/CDF released by secondary copper smelting operations in the United States is 1.17E+04 grams per year and the best estimate of TEQ emission is 2.34E+02 grams per year. A "high" confidence rating is given to the production estimate because it is based on reliable data from the U.S. 1987 Census of Manufactures. A "low" confidence rating is given to the emission estimates since they are based on direct measurements at only one U.S. copper smelter. Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (234 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 74 to 740 g TEQ/yr.

The secondary lead smelting industry produces elemental lead through the chemical reduction of lead compounds (obtained primarily from scrap motor vehicle lead-acid batteries) in a high temperature furnace (1,200 to 1,260 degrees C). Smelting is performed in reverberatory, blast, rotary, or electric furnaces. Blast and reverberatory furnaces are the most common types of smelting furnaces used by the 23 facilities that comprise the current secondary lead smelting industry in the United States. Of the 45 operating furnaces at these 23 facilities, 15 are reverberatory furnaces, 24 are blast furnaces, 5 are rotary furnaces, and 1 is an electric furnace. The one electric furnace and 11 of the 24 blast furnaces are co-located with reverberatory furnaces and most share a common exhaust and emissions control system (U.S. EPA, 1994a). Furnace charge materials consist of lead-bearing raw materials, lead-bearing slag and drosses, fluxing agents (blast and rotary furnaces only), and coke. Fluxing agents consist of iron, silica sand, and limestone or soda ash. Coke is used as fuel in blast furnaces and as a reducing agent in reverberatory and rotary furnaces. The PVC plastic seperators in the batteries are the primary source for HCl emissions from the smelters. However, the fluxing agents used at blast and rotary furnaces also react with chlorine to form calcium chloride or sodium chloride therby reducing HCl emissions from these furnaces relative to reverberatory furnaces. Organic emissions from co-located blast and reverbertory furnaces are more similar to the emissions of a reverberatory furnace than the emissions of a blast furnace (U.S. EPA, 1994a).

The total annual production capacity of the U.S. lead smelting industry is 1.36 million metric tons. Blast furnaces not co-located with reverberatory furnaces account for 21 percent of capacity (or 0.28 million metric tons). Reverberatory furnaces and blast and electric furnaces co-located with reverberatory furnaces account for 74 percent of capacity (or 1.01 million metric tons). Rotary furnaces account for the remaining 5 percent of capacity (or 0.07 million metric tons). Actual production volume statistics by furnace type are not available. However, if it is assumed that the total actual production volume of the industry (0.86 million metric tons in 1990) is reflective of the production capacity breakdown by furnace type, then the estimated actual production volumes of blast furnaces (not co-located), reverberatory and co-located blast/electric and reverberatory furnaces, and rotary furnaces are 180, 637, and 43 thousand metric tons, respectively (U.S. EPA, 1994a).

CDD/CDF and TEQ emission factors can be estimated for lead smelters based on the results of emission tests recently performed by EPA at three smelters (a blast furnace, a co-located blast/reverberatory furnace, and a rotary kiln furnace) (U.S. EPA, 1992i; 1993g; 1993h). The air pollution control systems at the three tested facilities consisted of both baghouses and scrubbers. Congener-specific measurements were made at the exit points of both APCD exit points at each facility. Although all 23 active smelters employ baghouses, only 9 employ scrubber technology. Facilities that employ scrubbers account for 14 percent of the blast furnace (not co-located) production capacity, 52 percent of the reverberatory and co-located furnace production capacity, and 57 percent of the rotary furnace production capacity. From the reported data, TEQ emission factors (ng TEQ/kg lead recovered) for each of the three furnace configurations are presented below as a range reflecting the presence or absence of a scrubber.

• Blast furnace: 0.63 to 8.31 ng TEQ/kg lead

• Reverberatory/co-located furnace: 0.10 to 0.77 ng TEQ/kg lead

• Rotary furnace: 0.28 to 0.21 ng TEQ/kg lead

If it is assumed that these emission rate ranges are representative of the range of emission rates at the non-tested facilities with the same basic furnace configuration and presence or absence of scrubbers, then combining these emission rate ranges with the estimates derived above for annual secondary lead production volumes yields a total industry-wide estimated emission to air of 1.6 g of TEQ. The estimated contributions to this total for each furnace configuration are:

• Blast furnaces with scrubbers: 0.02 g TEQ/yr

• Blast furnaces without scrubbers: 1.29 g TEQ/yr

• Reverberatory/co-located furnaces with scrubbers: 0.03 g TEQ/yr

• Reverberatory/co-located furnaces without scrubbers: 0.24 g TEQ/yr

• Rotary furnaces with scrubbers: 0.01 g TEQ/yr

• Rotary furnaces without scrubbers: 0.01 g TEQ/yr

A "medium" confidence rating is ascribed to the emission factors derived above because stack test data were available for 3 of the 23 active smelters in the United States and the stack test data used represent the three major furnace configurations. The "production" estimate has been assigned a "medium" confidence rating because, although it is based on a U.S. Department of Commerce estimate of total U.S. production, no production data were available on a furnace type or furnace configuration basis.

Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual emissions (1.6 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 0.7 to 3.5 g TEQ/yr.

Several European investigators have reported that iron ore sinter plants are sources of CDD/CDFs (Rappe et al., 1992b; Lexen et al., 1993; Lahl, 1993). However, insufficient information is available for evaluating CDD/CDF emissions from primary smelting/refining of ferrous metal in the United States.

Iron is manufactured from its ores (i.e., magnetic pyrites, magnetite, hematite, and carbonates of iron) in a blast furnace, and the iron obtained from this process is further refined in steel plants to make steel. During iron manufacturing, iron ores undergo sintering to enable better processing in the blast furnace. In the sintering process, iron ore fines are mixed with coke fines and the mixture is placed on a grate which is then heated to a temperature of 1093-1371^oC. The heat generated during combustion sinters the small particles. Also, iron-bearing dusts and slags from other processes in the steel plant are recycled as a feed mix for the sinter plant (Knepper, 1981; Capes, 1983). Lahl (1993) reported that the this management practice introduces traces of chlorine and organic compounds which are responsible for the generation of the CDD/CDFs found in these plants.

Sinter plants in Sweden and the Netherlands were reported to emit up to 3 ng TEQ/m³ stack gas or 2 to 4 g TEQ/yr per plant to the air (Rappe et al., 1992b; Lexen et al., 1993). Lahl (1993) report that emission data from plants in Germany indicate TEQ concentrations in stack gas after passage through mechanical filters and electrostatic precipitators ranging from 3 to 10 ng TEQ/nm³. Lahl (1993) estimated that, if all European sinter plants have stack concentrations of the same order of magnitude, then the total emission from sintering plants would be greater than 1 kg TEQ. This total is greater than the sum of all other identified European thermal sources of CDD/CDFs.

Tysklind et al. (1989) found scrap ferrous metal processing to be a source of CDDs and CDFs at a steel mill in Sweden. Analyses showed the presence of CDDs and CDFs in the range of 0.1 to 1.5 ng TEQ/Nm³ dry gas in a plant with a 10-ton electric furnace. The higher values reportedly were obtained during the melting of metal with chlorinated materials (e.g., PVC plastics). Raw gases collected over an open-furnace during batch jobs contained 110 ng TEQ/Nm³ dry gas when cutting oils containing chlorine were added to the scrap metal. The congener profiles of all flue gas samples showed that CDFs were predominant. The congener profiles also showed higher chlorine content when PVC was used.

Insufficient data exist to estimate emission factors for the U.S.

The objective of wire recovery is to remove the insulating material and reclaim the metal (e.g., copper, silver, and gold) comprising the electric wire. The reclaimed metal is then sold to a secondary metal smelter. Wire insulation commonly consists of a variety of plastics, asphalt-impregnated fabrics or burlap. In ground cables, chlorinated organics are used to preserve the cable casing.

In the past, scrap electric wire was thermally treated in the United States to burn off the insulating material. However, according to industry and trade association representatives, current recovery operations typically no longer involve thermal treatment but instead involve mechanical chopping into fine particles from which the insulating material is removed by air blowing and gravitational settling of the heavier metal fraction (telephone conversation between R. Garino, Institute of Scrap Recycling Industries, and T. Leighton, Versar, Inc. on March 2, 1993; telephone conversation between J. Sullivan, Triple F. Dynamics, and T. Leighton, Versar, Inc., on March 8, 1993). No independent confirmation of this technology switch could be obtained from EPA program office representatives.

The combustion of chlorinated organic compounds catalyzed by the presence of wire metals such as copper and iron can lead to the formation of CDDs and CDFs (Van Wijnen et al., 1992). CDDs and CDFs have been detected in fly ash and bottom residues from the open-air incineration of wire scraps, and in stack samples of a wire reclamation incinerator (Chen et al., 1986; Southerland et al., 1987). Huang et al. (1992b) detected CDDs and CDFs in soils collected near electronic wire scrap incinerators used for the recovery of metals. In these studies, the chlorinated compounds were considered to have been generated thermochemically from plastics covering the wires. Small-scale (and unpermitted) activities involving the incineration of scrap electrical wires have resulted in increased levels of CDDs and CDFs in soil samples collected from former scrap wire and car incineration sites within the vicinity of Amsterdam (Van Wijnen et al., 1992). Analysis of these soil samples showed CDD and CDF levels ranging between 60 and 98,000 ng/kg dry weight, with nine of fifteen soil samples having levels above 1,000 ng/kg dry weight.

Dioxin-like compounds emitted to the air from scrap electric wire incineration were measured from a facility during EPA’s National Dioxin Study of combustion sources (U.S. EPA, 1987). The tested facility was determined to be typical of this industrial source category at that time. Insulated wire and other metal-bearing scrap material were fed to the incinerator on a steel pallet. The incinerator was operated in a batch mode, with the combustion cycles for each batch of scrap feed lasting between 1 and 3 hours. Incineration of the material occurred by burning natural gas. Most of the wire had a tar-based insulation that was thermally removed; however, PVC coated wire was also fed to the incinerator. The estimated temperature during combustion was 650° C, and combustion preceded in a primary and secondary chamber. The tested facility was equipped with a high temperature afterburner to further destroy organic compounds entrained in the combustion gases prior to discharge to the air from the stack. Emission factors estimated for this one facility include an average emission factor for TEQ of 1.18E-02 µg/kg of scrap wire (range = 6.74E-03 to 1.69E-02 µg/kg), and an average emission for total CDD/CDF of 9.89E-01 µg/kg of scrap wire (range = 9.89E-01 to 3.28E+00 µg/kg). These emission factors are assigned a "low" confidence rating because the factors were derived from measurements at only one facility operating in the U.S. and it is not known how representative these test results are of other scrap electric wire incinerators. Although it is uncertain how many facilities still combust scrap wire, for purposes of this assessment, it is assumed that only minimal quantities of scrap wire are currently burned in the United States.

Hutzinger and Fiedler (1991b) reported that the CDDs and CDFs are emitted in stack gases from drum and barrel reclamation facilities and that the concentration of CDDs/Fs found in those emissions range from 5 to 27 ng/m³. Dioxin-like compounds were measured by EPA in the stack gas emissions of a drum and barrel reclamation furnace as part of the National Dioxin Study (U.S. EPA, 1987). These plants operate a burning furnace to prepare used steel 55-gallon drums for cleaning to base metal. The drums processed at these facilities come from a variety of sources in the petroleum and chemical industries. The cleaned drums are repaired, repainted, relined and sold for reuse. The drum burning process subjects used drums to an elevated temperature in a tunnel furnace for a sufficient time so that the paint, interior linings, and previous contents are burned or disintegrated. The furnace is fired by auxiliary fuel. Used drums are loaded onto a conveyor that moves at a fixed speed. As the drums pass through the preheat and ignition zone of the furnace, additional contents of the drums drain into the furnace ash trough. A drag conveyor moves these sludges and ashes to a collection pit. The drums are air cooled as they exit they furnace. Exhaust gases from the burning furnace are drawn through a breeching fan to a high-temperature afterburner.

Emission factors of dioxin-like compounds were developed from EPA stack tests of a prototypical operation (U.S. EPA, 1987) yielding the following emission factors in units of µg/kg: minimum TEQ = 1.12E-02; mean TEQ = 1.65E-02; maximum TEQ = 2.69E-02; mean CDD/CDF = 1.30E+00; minimum CDD/CDF = 1.16E+00; maximum CDD/CDF=1.85E+00.

Approximately 2.8 to 6.4 million 55-gallon drums are incinerated annually in the United States (telephone conversation between P. Rankin, Association of Container Reconditioners, and C. D'Ruiz, Versar, Inc., December 21, 1992). This estimate is based on the following assumptions: 1) 23 to 26 incinerators are currently in operation; 2) each incinerator, on average, handles 500 to 1,000 drums per day; and 3) on average, each incinerator operates 5 days per week with 14 days downtime per year for maintenance activities. The weight of 55-gallon drums varies considerably; however, on average, a drum weighs 38 lbs (or 17 kg). Therefore, an estimated 48 to 109 million kg of drums are estimated to be incinerated annually. Assuming that 109 million kg of drums are burned each year and applying the mean emission factors developed above, the best estimates of annual emissions are 140 grams per year of total CDD/CDF and 1.7 grams per year of TEQ.

A "low" confidence rating is assigned to the production estimate since it is based an expert judgement rather than a published reference. A "low" confidence rating is ascribed to the emission factor since it is developed from stack tests conducted by EPA on just one U.S. drum and barrel furnace and this one facility may not represent emissions from all current operations in the U.S. Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (1.7 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 0.5 to 5.4 g TEQ/yr.

Emissions of dioxin-like compounds from the incineration of automobile tires were measured from a tire incinerator stack tested by the State of California Air Resources Board (CARB, 1991). The facility consists of two excess air furnaces equipped with steam boilers to recovery the energy from the heat of combustion. Discarded whole tires are fed to the incineration units at a rate of 3000 kg/hr. The furnaces are equipped to burn natural gas as auxiliary fuel. The steam produced from the boilers is used to drive electrical turbine generators that produce 14.4 megawatts of electricity. The facility is equipped with a dry acid gas scrubber and fabric filter for the control of emissions prior to exiting the stack.

Emission factors for total CDD/CDF and TEQ in units of µg/kg of tires combusted were derived as average values from the one tested facility stack tested in California (CARB, 1991). From these data, an average emission factor of CDD/CDF was estimated to be 1.39E-02 µg/kg of tires incinerated (range: 4.28E-03 to 3.05E-02 µg/kg), and average emission factor of TEQ was estimated to be 5.42E-04 µg/kg (range: 1.91E-04 to 1.02E-03 µg/kg).

EPA's Office of Solid Waste estimates that approximately 0.50 million metric tons of tires are incinerated in the United States annually (U.S. EPA, 1992a). This production estimate is given a "high" confidence rating since it is based on detailed study specific to the United States. The use of scrap tires as a fuel increased significantly during the late 1980s. In 1990, 10.7 percent of the 242 million scrap tires generated were burned for fuel. This percentage is expected to continue to increase (U.S. EPA, 1992a).

If it is assumed that 500 million kilograms of discarded tires are incinerated annually in the United States, then, using the emission factors derived from stack data from the one tested facility, an average of 6.9 grams of total CDD/CDF per year and an average of 0.3 grams of TEQ per year are estimated to be emitted to the air. It must be noted that these may be underestimates of emissions from this source category because the one facility tested in California is equipped with a dry-scrubber combined with a fabric filter for air pollution control. These devices are capable of greater than 95 percent reduction and control of dioxin-like compounds prior to discharge from the stack. It is not know to what extent other tire incineration facilities operating in the U.S. are similarly controlled. If such facilities are not so equipped, then the uncontrolled emission of CDD/CDF and TEQ could be much greater than the estimates developed above. Therefore, the estimated emission factor of dioxin from tire incineration is given a "low" confidence rating. Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (0.3 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 0.1 to 1.0 g TEQ/yr.

Buser et al. (1991) indicated that PCDTs (polychlorinated dibenzothiophenes, possibly a dioxin-like compound) can be formed in situations where large amounts of sulfur and chlorine-containing compounds are incinerated or accidentally burned. Automobile tires are known to contain sulfurous (vulcanization) compounds and certain types of chloro compounds (e.g., chloroprene). Thus, it is possible that the burning of used automobile tires could result in the formation of PCDTs.

Some of the first evidence that CDD/CDFs might be created during the combustion processes in gasoline- and diesel-fueled engines came from Ballschmiter et al. (1986) who measured these compounds in used motor oil in Germany. Incomplete combustion and the presence of a chlorine source in the form of additives in the oil or the fuel (such as dichloroethane or pentachlorophenate) were speculated to lead to the formation of CDDs and CDFs. The isomeric patterns were characterized as typical of combustion processes.

Marklund et al. (1987) provided the first direct evidence of these compounds in car emissions based on tailpipe measurements on four Swedish cars running on leaded gasoline. They found 20 to 220 pg of TEQ per kilometer driven. For this study, a nonleaded gasoline was used to which was added tetramethyl lead (0.15 grams of lead per liter or 0.57 grams per gallon) and dichloroethane (0.1 g/L as a scavenger). The fuel used may not accurately represent commercial fuels which typically contain a mixture of chlorinated and brominated scavengers (Marklund et al., 1990). Also, the lead content of the fuel used (0.15 g lead/L), although the normal lead content for Swedish fuels (Marklund et al., 1990), is higher than the lead content of leaded gasoline in the United States during the late 1980s (lowered to 0.10 g lead/gallon or 0.026 g lead/L effective January 1, 1986). For two cars running on unleaded gasoline, CDD/CDF emissions were below the detection limit which corresponded to approximately 13 pg of TEQ per kilometer driven. Table 3-33 presents a summary description of this study and subsequent studies discussed below.

Bingham et al. (1989) also analyzed 2,3,7,8-substituted CDD/CDFs in automobile exhausts. Four cars using leaded gasoline (0.45 g/L tetramethyllead, 0.22 g/L dichloroethane and 0.2 g/L dibromoethane) were tested and one car using lead free gasoline was tested. Only HpCDD and OCDD were detected in the exhaust from the vehicle using lead-free fuel. The total TEQ emission rate, based on these detected congeners, was 1 pg/km; the detection limit for the other 2,3,7,8-substituted CDD/CDFs was a combined 28 pg TEQ/km. 2,3,7,8-TCDF was detected in the exhaust of two of four cars using leaded fuel. OCDD was detected in the exhaust from three of the cars and PeCDF and HpCDD were each detected in the exhaust from one car. TEQ emission rates for the cars using leaded fuel, based on detected congeners, was 5 to 39 pg/km.

Haglund et al. (1988) sampled exhaust gases from three different vehicles (two cars fueled with leaded and unleaded gasoline, respectively, and a heavy diesel truck) for the presence of brominated dibenzo-p-dioxins (BDD) and brominated dibenzofurans (BDF). The authors concluded that the dibromoethane scavenger added to the tested gasoline probably acted as a halogen source. TBDF emissions measured 23 ng/km in the car with leaded gasoline and 0.24 ng/km in the car with unleaded gasoline. TBDD and PeBDF emissions measured 3.2 and 0.98 ng/km, respectively in the car with leaded gasoline. All BDD/Fs were below detection limits in the diesel truck emissions.

More recently, Marklund et al. (1990) tested gasoline- and diesel-fueled vehicles, measuring CDD/CDF emissions before and/or after the muffler of Swedish vehicles (including new and old vehicles). Marklund et al. (1990) reported the emission results in units of pg TEQ/L of fuel consumed and also in units of pg TEQ/km driven during the test. Based on the test driving cycle employed (i.e., 31.7 km/hr as a mean speed; 91.2 km/hr as a maximum speed; and 17.9 percent of time spent idling), Marklund et al. (1990) observed a fuel economy of approximately 9 to 10 km/L or 22 to 24 miles/gallon. The following measurements were reported:

n leaded gas/cars/before muffler: 2.4 to 6.3 pg TEQ/km (or 21 to 60 pg TEQ/L of fuel consumed)

n leaded gas/cars/in tailpipe: 1.1 to 2.6 pg TEQ/km (or 10 to 23 pg TEQ/L).

n lead-free gas/catalyst-equipped car/in tailpipe: 0.36 pg TEQ/km (or 3.5 pg TEQ/L)

n lead-free gas/cars/before muffler: 0.36 to 0.39 pg TEQ/km (or 3.5 pg TEQ/L)

n diesel fuel/heavy-duty truck/before muffler: not detected (i.e., less than 100 pg TEQ/L)

Regarding the diesel fuel measurement, the authors pointed out that the test fuel was a reference fuel and may not be representative of commercial diesel fuel. Also, due to analytical problems, a much higher detection limit (about 100 pg TEQ/L) was employed in the diesel fuel tests than in the gasoline tests (5 pg TEQ/L). Further uncertainty is introduced by the fact that diesel emission samples were only collected prior to the muffler. The TEQ levels in exhaust gases from older cars using leaded gasoline were up to six times greater when measured before the muffler than after the muffler. No muffler-related difference in new cars running on leaded gasoline or in old or new cars running on unleaded gasoline were observed.

Hagenmaier et al. (1990) ran a series of tests on gasoline engines for light duty vehicles in Germany. The following average TEQ emission rates per liter of fuel consumed were found:

n Leaded fuel: 1.083 ng TEQ/L

n Unleaded fuel (catalyst-equipped): 0.007 ng TEQ/L

n Unleaded fuel (not catalyst-equipped): 0.051 ng TEQ/L

n Diesel fuel: 0.024 ng TEQ/l

Several European studies have evaluated CDD/CDF emissions from vehicles by measuring the presence of CDD/CDFs in tunnel air. This approach has the advantage that it allows random sampling of large numbers of cars, including a range of ages and maintenance levels. The disadvantage of this approach is that it relies on indirect measurements (rather than tailpipe measurements) which may introduce unknown uncertainties and make interpretation of the findings difficult. Concerns have been raised that the tunnel monitors are detecting resuspended particulates which have accumulated over time, leading to overestimates of emissions. Also, the driving patterns encountered in these tunnel studies are more or less steady state driving conditions rather than the transient driving cycle and cold engine starts that are typical of urban driving conditions and which may affect emission levels. Wevers et al. (1992) found that CDD/CDF levels inside a Belgium tunnel were about twice the levels in ambient air and estimated the average level in vehicle emissions as 42 to 45 pg TEQ/Nm³. Rappe et al. (1988) conducted a similar tunnel study in Sweden and Oehme et al. (1991) conducted a similar study at a tunnel in Norway, the preliminary results of which were reported by Larssen et al. (1990). The Oehme et al. (1991) study estimated emissions for light duty and heavy duty vehicle classes. This was completed by counting the number of light duty vs. heavy duty vehicles passing through the tunnel during the study. The mean emission rate estimates from this study are:

n Light-duty vehicles using gasoline (approximately 30 percent using leaded gas): 0.28 ng TEQ/km

n Heavy-duty diesel trucks: 5.1 ng TEQ/km

These mean values are the averages of the emission rates corresponding to two operating modes: vehicles moving uphill on a 3.5 percent incline and vehicles moving downhill on a 3.5 percent decline. The TEQ emission rates for the two modes differ by an order of magnitude for both light and heavy duty vehicles. Although Oehme et al. (1991) reported results in units of Nordic TEQs rather than I-TEQs, the results in I-TEQ should be virtually identical because the only difference between the two TEQ schemes is the factor assigned to 1,2,3,7,8-PeCDF (0.1 in Nordic and 0.05 in I-TEQ), a minor component of the toxic CDD/CDFs measured in the tunnel air.

Virtually no testing of vehicle emissions in the United States for CDD/CDFs has been published. In 1987, the California Air Resources Board (CARB) produced a draft report on the testing of the exhausts of four gasoline-powered cars and three diesel fuel-powered vehicles (one truck, one bus, and one car) (CARB, 1987). However, CARB has indicated to EPA that the draft report should not be cited or quoted to support general conclusions about CDD/CDFs in motor vehicle exhausts because of the small sample size of the study and because the use of low rather than high resolution mass spectrometry in the study resulted in high detection limits and inadequate selectivity in the presence of interferences (Lew, 1993). CARB did state that the results of a single sample from the heavy-duty diesel truck could be reported because congeners from most of the homologue groups were present in the sample at levels that could be detected by the analytical method and there were no identified interferences in this sample. However, it should be noted that this test was conducted under steady state conditions and at low speeds which are not indicative of normal driving patterns. The TEQ content of this one sample was 218 pg per dry standard cubic meter (dscm) of exhaust. The CARB results suggest that diesel-fueled trucks do emit CDD/CDFs (Lew, 1993).

Jones (1993) estimated CDD/CDF emissions of the major vehicle types on the basis of the studies by Larssen et al. (1990), the 1987 draft report by CARB, and Hagenmaier et al. (1990). Using data on vehicle miles travelled in the U.S. and an assumed emission rate of 5.4 ng TEQ/km based on Larssen et al. (1990), Jones (1993) estimated that about 1000 g of TEQ were emitted from diesel vehicles nationwide in 1990. Jones (1993) also suggests that human exposure to diesel emissions are exacerbated relative to stack emissions from combustion sources on the basis that diesel emissions occur at ground level and that, unlike stack emissions, may not undergo much dilution in air before human contact occurs.

In 1973, EPA required refiners to meet a 0.5 gpg (gram per gallon) standard for the average lead content of all gasoline. EPA later replaced this standard with a standard for the lead content of leaded gasoline only. Effective November 1, 1982, large refineries were required to meet a standard of 1.10 grams per leaded gallon (gplg). Certain smaller refineries were subject to a 1.90 gplg standard until July 1, 1983, at which time they would also be subject to the 1.10 gplg standard (Federal Register, 1982). EPA further reduced the standard to 0.10 gplg effective January 1, 1986 with a[n] interim standard of 0.5 gplg effective July 1, 1985 (Federal Register, 1985). The Clean Air Act Amendments of 1990 imposed further restrictions as follows: "After December 31, 1995, it shall be unlawful for any person to sell, offer for sale, supply, offer for supply, dispense, transport, or introduce in commerce, for use as fuel in any motor vehicle any gasoline which contains lead or lead additives."

In 1985, the year before the phasedown of leaded gasoline from 1.10 gplg to 0.10 gplg, approximately 1,774 billion miles (2,855 billion km) were driven (U.S. DOC, 1992). Because leaded gasoline accounted for 35.5 percent of the gasoline supply that year (EIA, 1993) it can be estimated that 1,013 billion of these kilometers (i.e., 35.5 percent of 2,855 billion km) were driven by vehicles powered with leaded gasoline.

The U.S. Federal Highway Administration, as reported in U.S. DOC (1992), reports that 2,148 billion total vehicle miles (3,456 billion km) were driven in the U.S. during 1990. During 1990, automobiles and motorcycles accounted for 1,525 billion vehicle miles (2,454 billion km). Trucks accounted for 617 billion vehicle miles (993 billion km) and buses accounted for 5.7 billion vehicle miles (9.2 billion km) (U.S. DOC, 1992). In 1987, diesel-fueled trucks accounted for 17.2 percent of total truck vehicle km driven; gasoline-fueled trucks accounted for the remaining 82.8 percent (U.S. DOC, 1990b). Applying this factor (i.e., 17.2 percent) to the 1990 km truck mile estimate (i.e., 993 billion km) indicates that an estimated 171 billion km were driven by diesel-fueled trucks in 1990. It is assumed that all other vehicle km driven (3,285 billion km) were those of gasoline-powered vehicles. In 1990, leaded gasoline accounted for only 5.3 percent of total gasoline supplies (EIA, 1993). These mileage estimates are given a "high" confidence rating on the basis that they are based on U.S. Census of Transportation studies.

Using the above literature, separate emission estimates were developed for vehicles burning leaded gasoline, unleaded gasoline and diesel fuel:

n Leaded Gasoline: In general, the literature indicates that CDD/CDF emissions occur from vehicles using leaded gasoline and that considerable variation occurs depending, at least in part, on the types of scavengers used. Marklund et al. (1987) reported emissions ranging from 20 to 220 pg TEQ/km from four cars fueled with a reference fuel (0.5 gplg) to which lead and a chlorinated scavenger were added. Marklund et al. (1990) reported much lower emissions in the tailpipe exhaust of two cars (1.1 to 2.6 pg TEQ/km) using a commercial leaded fuel (0.57 gplg). Marklund et al. (1990) attribute the difference in the emission measurements to the different scavengers used in the two studies. Hagenmaier et al. (1990) reported TEQ emissions of 1,083 pg/L of fuel (or approximately 108 pg TEQ/km) from a car fueled with a commercial leaded fuel (lead content not reported). Bingham et al. (1989) reported emissions from four cars using gasoline with a lead content of 1.7 gplg in New Zealand to range from 5 to 39 pg/km. The tunnel study by Oehme et al. (1991) indicated that emissions from cars could be 38 to 520 pg TEQ/km. Since most of the vehicles passing through the tunnel studied by Oehme et al. (1991) used unleaded fuels (approximately 70 percent), the emissions from leaded fuel-powered cars were possibly even higher.

On the basis of the three studies performed using commercial leaded fuel, an emission factor range of 1.1 to 108 pg TEQ/km is recommended. A "low" confidence rating is assigned to this factor range because the range is based on European fuels and emission control technologies which may differ from U.S. fuels and technology and also because the factor range is based is based on tests with only seven cars. Combining this emission factor range with the estimates for km driven by leaded fuel-powered vehicles in 1985 (1,013 billion km or 35.5 percent of total km) and in 1990 (174 billion km or 5.3 percent of the 3,285 billion km driven by gasoline-powered vehicles) suggests that about 1.1 to 109 g TEQ/yr were emitted from vehicles using leaded fuels in 1985, the year immediately before the phasedown of leaded gasoline from 1.10 gplg to 0.10 gplg and the year when the interim standard of 0.5 gplg became effective. By comparison, the annual emission for 1990 from use of leaded gasoline is estimated to have ranged from 0.2 to 19 g TEQ.

n Unleaded Gasoline: The literature clearly indicates that CDD/CDF emissions are much less from vehicles burning unleaded fuels. The Marklund et al. (1990) study is the only one which provided an emission factor for this class of vehicles. On the basis of this study, an emission factor of 0.36 pg TEQ/km is recommended. A "low" confidence rating is assigned to this factor because the Swedish fuels and emission control technology used in the Marklund et al. (1990) study may differ from U.S. fuels and technology and also because the emission factor is based on tests with only one catalyst-equipped car. Combining this emission factor with the above estimates for vehicle km driven in 1990 by gasoline-powered vehicles (3,285 billion), suggests that about 1.3 g of TEQ/yr were emitted from vehicles using unleaded fuels in 1990. Based on the low confidence rating, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (1.3 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 0.4 to 4.1 g TEQ/yr.

n Diesel Fuel: Very few data are available upon which to base an evaluation of the extent of dioxin emissions resulting from diesel fuel combustion. The tunnel study by Oehme et al. (1991) generated an estimated emission factor of 5.1 ng TEQ/km. A "low" confidence rating is assigned to this factor because the factor is based on Norwegian fuels and emission control technology which may differ from U.S. fuels and technology. Also, although aggregate samples were collected representing hundreds of vehicles, the indirect method of analysis and the more or less steady state rather than transient driving conditions of the study introduce considerable uncertainty.

The results of only one tailpipe measurement (diesel fuel in a heavy-duty Swedish truck) have been published (Marklund et al., 1990) and that study reported no emissions at a detection limit of 100 pg TEQ/L. If it is assumed that the fuel economy of heavy-duty diesel vehicles is approximately 5 miles/gallon (or 2 km/L), then 100 pg TEQ/L converts to approximately 0.05 ng TEQ/km - a factor 100-fold lower than the emission rate reported by Oehme et al. (1991). Because the results of Marklund et al. (1990) are based on only one vehicle using a Swedish reference, not a commercial, diesel fuel this emission factor is also assigned a "low" confidence rating.

To obtain an estimate of the possible range of dioxin TEQ annual emissions resulting from diesel fuel use, the geometric mean of the emission factors derived from the Oehne et al. (1991) and Marklund et al. (1990) was calculated (0.5 ng TEQ/km). Combining this emission factor with the above estimate for vehicle kms driven in 1990 in the United States by diesel-fueled trucks (171 billion km) yields an annual emission estimate of 85 g TEQ/yr. Based on the low confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (85 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 27 to 270 g TEQ/yr.

Measurable levels of TCDDs have been found in chimney soot and bottom ash from wood-burning stoves and fireplaces (Clement et al., 1985b; Wenning et al., 1992). Chimney deposits from residential wood burning have been found to have CDD/CDF congener profiles similar to those in flue gases from municipal waste incineration (Bacher et al., 1992). Bacher et al. (1992) found concentrations of 2,3,7,8-substituted CDF and CDD congeners in soot from wood burning ranging from 40 to 930 ng/kg and from 30 to 150 ng/kg, respectively. Bacher et al. (1992) reported that the CDFs dominated the CDDs by a factor of 5 to 10 and that the lower chlorinated CDDs/CDFs (mono-through tri-) dominated the more chlorinated CDDs/CDFs. The TEQ content of the chimney soot was 720 ng/kg (Bacher et al., 1992).

Nestrick and Lamparski (1983) conducted a study of CDD formation in residential wood-burning chimneys in different areas of the United States. The results of their survey are presented in Table 3-34. As seen in Table 3-28, the eastern U.S. had overall higher estimated levels of TCDD generation than did the central or western United States. Levels of TCDDs in chimney soot ranged from 22 to 410 ng/kg for the eastern U.S., 21 to 294 ng/kg in the central U.S., and 2.9 to 28 ng/kg in the western U.S. Red oak and oak were the predominant fuels used in the eastern and western U.S., and ash, birch, and oak were the predominant fuels used in the central U.S.

Two studies have recently become available which provide direct measurement of CDD/CDF in emissions from wood stoves. These studies are summarized below.

Schatowitz et al. (1993) measured CDD/CDF in the emissions of a variety of residential wood burners in Switzerland. The study included three types of burners (household stoves, automatic chip burners, and wood boilers) and a variety of wood fuels (natural beech wood, natural wood chips, chipboard chips, waste wood chips, charcoal and household waste). The following emission factors were derived:

• household stove with open door burning natural beech wood: 0.77 ng TEQ/kg

• household stove with closed door burning natural beech wood: 1.25 ng TEQ/kg

The open door stove can be assumed to be representative of fireplaces since both have an uncontrolled draft. Also, Schatowitz et al. (1993) measured emissions from wood burning fireplaces and report the same flue gas concentration as found with the open door wood stove (i.e., 0.064 ng TEQ/m³). All of the toxic congeners of CDD/CDF were found at levels above the detection limit.

Vikelsoe et al. (1993) studied emissions of CDD/CDFs from residential wood stoves in Denmark. The wood fuels used in the experiments were seasoned birch, beech and spruce harvested in Denmark. Four different types of stoves were evaluated under normal and optimal operating conditions. Widely varying emissions were found for different fuel/stove combinations. The emissions from spruce were about twice as high as the emissions from birch and beech. In many of the experiments the CDD/CDF emissions were higher when the stove was operated under normal conditions vs optimal conditions (minimum CO). The weighted average (considering wood and stove types) emission factor for wood stoves in Denmark was estimated to be 1.9 ng Nordic-TEQ/kg.

Based on the above studies, 1 ng TEQ/kg appears to be a reasonable average emission factor for residential wood burning. A "medium" confidence rating was assigned to this estimate on the basis that: (1) it is derived from only two studies; (2) both studies used direct measurement; and (3) although the studies were conducted in Europe, residential wood burning practices are probably sufficiently similar to apply to the United States.

In 1990, wood provided about 2.8 percent of the total primary energy consumed in the United States (EIA, 1991). Total wood energy consumption during 1990 is estimated at 2,359 trillion BTU. Assuming that 1 kg of oven-dried wood (i.e., 2.15 kg of green wood) provides approximately 19,000 BTU, then an estimated 124.2 million metric tons of oven-dried wood equivalents were burned for energy purposes in 1990 (EIA, 1991). Residential wood consumption in 1990 was estimated at 786 trillion BTU (41.4 million metric tons), or 33 percent of total U.S. consumption. Industrial fuel wood consumption in 1990 totaled 1,562 trillion BTU (82.2 million metric tons), or 66 percent of total U.S. consumption, with the majority (1,232 trillion BTU) of this fuel wood being consumed by the Paper and Allied Products Industry. 1990 consumption of fuel wood by the utility sector was approximately 11.9 trillion BTU (0.6 million metric tons) (EIA, 1991). These production estimates are given a "high" confidence rating since they are based on a detailed published study specific to the United States.

Combining the best estimate of the emission factor (1 ng TEQ/kg wood) with the mass of wood consumed annually by residences, indicates that the annual TEQ emissions from this source are about 40 grams. Based on the "medium" confidence rating assigned to the emission factor, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual emissions (40 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 13 to 63 g TEQ/yr.

Emissions of dioxin-like compounds have been measured in stack emissions from an industrial wood-burning furnace by EPA (U.S. EPA, 1987). The tested facility was located at a lumber products plant that manufactures overlay panels and other lumber wood products. The wood-fired boiler tested was a three-cell dutch oven equipped with a waste heat boiler. During normal operation, the furnace is 100 percent fired with scrap wood from the lumber plant. The feed wood is a mixture of bark, hogged wood, and green and dry planar shavings. The composition of the feed was estimated to be wood from fir and hemlock. Nearly all the wood fed to the lumber plant had been stored in sea water adjacent to the facility, and therefore had a significant concentration of inorganic chloride. The scrap wood fed to the boiler had not been treated with chemical preservatives, e.g., pentachlorophenol. The wood was fed to the boiler by a screw conveyor that dumps the feed into a pile in the primary combustion chamber. The furnace is operated at air in 50 percent excess of stoichiometric requirements. Boilers capture the heat of combustion and transfer the heat into steam for co-generation of energy at the plant. The exhausted gases from the boiler pass through a cyclone and fabric filter prior to discharge from the stack. From this study, an average emission factor for CDD/CDF of 1.02 µg/kg of wood burned (range: 5.52E-01 to 1.41E+00 µg/kg), and an average emission factor for TEQ of 1.71E-02 µg/kg of wood burned (range: 7.34E-03 to 2.28E-02 µg/kg) are estimated. Emissions testing at this facility demonstrated that the fabric filter was reducing dioxin emissions by about 90 percent (U.S. EPA, 1987).

In a second study, CDD/CDF was measured in the emissions from a quad-cell wood-fired boiler used to generate electricity (CARB, 1990b). The fuel consisted of coarse wood waste and sawdust from nonindustrial logging operations. The exhaust gas passed through a multicyclone before entering the stack. This study suggests an emission factor of 5.4E-02 m g/kg for CDD/CDF. If the same TEQ to total CDD/CDF ratio is assumed as in the first industrial burner study, then an emission factor of 9E-04 m g TEQ/kg can be estimated.

To obtain an estimate of the possible range of dioxin TEQ annual emissions resulting from industrial wood-burning facilities, the geometric mean of the emission factors derived from the U.S. EPA (1987) and CARB (1990b) studies was calculated (3.9 ng TEQ/kg wood). Because test data are available for only two facilities and because the emission factors measured at these two facilities vary greatly, this emission factor was given a "low" confidence rating.

In 1990, it was estimated that 82.2 million metric tons of wood were burned in industrial furnaces ("high" confidence rating, see discussion in Section 3.6.14). Applying the above emission factor to the estimated annual mass of wood burned by industrial facilities gives an estimated TEQ emission of 320 g TEQ/yr. Based on the low confidence rating given to the emission factor, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (320 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 100 to 1,000 g TEQ/yr.

Based on the findings of Wenning et al. (1992), Bacher et al. (1992), and Nestrick and Lamparski (1983), indicating generation of CDDs/CDFs in ash and soot during residential wood burning, it is reasonable to hypothesize that wood burned in forest fires may also be a source of CDDs/CDFs. Support for this hypothesis is provided by Bumb et al. (1980) who, in their study on trace chemistry of fire, have shown that combustion of hydrocarbons in the presence of chlorine compounds (which are naturally found at low levels in wood) can generate CDDs and CDFs in small amounts. Also, the pre-industrial existence of CDDs and CDFs, presumably due to combustion sources, has been demonstrated in analyses of ancient human tissues and ancient aquatic sediment deposits (ECETOC, 1992).

Only one study could be found that made direct measurements of CDD/CDFs in the actual emissions from forest fires. This study by Tashiro et al. (1990) detected levels ranging from about 15 to 400 pg/m³ for total CDD/CDFs. The samples were collected from fixed collectors 10 m above the ground and from aircraft flying through the smoke. Background samples collected before and after the tests indicated negligible levels in the atmosphere. These results were presented in the form of a preliminary report and no firm conclusions were drawn about whether forest fires are a CDD/CDF source. Coauthor Dr. Ray Clement presented the final report on this study at Dioxin '91. Clement and Tashiro (1991) reported total CDD/CDF levels in the smoke of about 20 pg/m³. The authors concluded that CDD/CDFs are emitted during forest fires but recognized that some portion of these emissions could represent resuspension from residues deposited on leaves rather than newly formed CDD/CDFs.

The concentrations presented by Clement and Tashiro (1991) cannot accurately be converted to an emission factor since the corresponding rates of combustion gas production and wood consumption are not known. As a result, three alternative approaches were considered to develop these emission factors:

Soot-Based Approach: This approach assumes that the level of CDD/CDFs in chimney soot are representative of the CDD/CDFs in emissions, and estimates the CDD/CDF emission rate as the product of the soot level and the total particulate emission rate. This involves first assuming that the CDD/CDF levels measured by Bacher et al. (1992) in chimney soot (720 ng TEQ/kg) are representative of the CDD/CDF concentrations of particles emitted during forest fires. Second, the total particulate generation rate must be estimated. Ward et al.(1976) estimated the national average particulate emission factor for wildfires as 150 lb/ton biomass dry weight based primarily on data for head fires. Ward et al. (1993) estimated the national average particulate emission factor for prescribed burning as 50 lb/ton biomass dry weight. Combining the total particulate generation rates with the CDD/CDF levels in soot yields emission factor estimates of 54 µg of TEQ and 18 µg of TEQ/metric ton of biomass burned in wildfires and prescribed burning, respectively. This corresponds to a range of 54 to 18 ng TEQ/kg of biomass. This estimate is likely to be an overestimate since the levels of CDD/CDF measured in chimney soot by Bacher et al. (1992) may represent accumulation/enrichment of CDD/CDFs measured in chimney soot over time, leading to much higher levels than what is actually on emitted particles.

Carbon Monoxide (CO) Approach: CO is a general indicator of the efficiency of combustion and the emission rate of many emission products can be correlated to the CO emission rate. The Schatowitz et al. (1993) data for emissions during natural wood burning in open stoves suggests an emission rate of 10 ug TEQ/kg of CO. Combining this factor with the CO production rate during forest fires (roughly 0.1 kg CO/kg of biomass - Ward et al. (1993)) yields an emission factor of 1000 ng TEQ/kg biomass. This factor appears unreasonably high since it is even higher than the soot-based factor discussed above. Although the formation kinetics of CDD/CDF during combustion are not well understood, it appears that CDD/CDF emissions do not correlate well with CO emissions.

Wood Stove Approach: This approach assumes that the emission factor for residential wood burning (using natural wood and open door, i.e., uncontrolled draft) applies to forest fires. As discussed in Section 3.6.14, this approach suggests an emission factor of about 1 ng TEQ/kg. This value appears more reasonable than the factors suggested by the soot and CO approaches. However, forest fire conditions differ significantly from combustion conditions in wood stoves. For example, forest fire combustion does not occur in an enclosed chamber and the biomass consumed in forest fires is usually green and includes underbrush, leaves and grass. Given these differences and the uncertainties about the formation kinetics of CDD/CDF during combustion, it is difficult to determine whether CDD/CDF emissions would be higher or lower from forest fires than from wood stoves. Thus, although an emission factor of 1 ng TEQ/kg appears to be the best estimate that can be made currently, it must be considered highly uncertain and a "low" confidence rating was assigned to this estimate.

According to the Council on Environmental Quality's 21st Annual Report (CEQ, 1990), an average of 5.1 million acres of biomass have been burned in wildfires every year from 1950 to 1990. This value also corresponds well to the data provided by the USDA Forest Service for 1975 in which 4.4 million acres of biomass were burned in wildfires (Ward et al., 1976). Yearly estimates cited in the CEQ report (CEQ, 1990) ranged from a high of 15.5 million acres burned to an annual low of 1.8 million acres burned over the forty year time period. Additionally, 5.1 million acres of biomass were burned in 1989 during prescribed burns (Ward et al., 1993). Prescribed burning is also known as managed or controlled burning and is used as a forest management tool under exacting weather and fuel conditions. These acreage estimates can be combined with biomass consumption rates of 10.4 tons/acre in areas consumed by wildfires (Ward et al., 1976) and 8.2 tons/acre in areas consumed in prescribed burns (Ward et al., 1993). This combination suggests a total of 53 million tons (or 48 million metric tons) of biomass are consumed annually in wildfires while a total of 42 million tons (or 38 million metric tons) of biomass are consumed annually in prescribed burns. These production estimates were assigned a "medium" confidence rating since they are based on a combination of estimates involving detailed historical data specific to the United States on acres burned but less accurate estimates of biomass burned/acre.

Combining the emission factor developed using the "wood stove" approach (i.e., 1 ng TEQ/kg biomass) with the amount of biomass consumed annually in wildfires and prescribed fires (total of 86 million metric tons) indicates that the best estimate of annual TEQ emissions from this source is 86 g. Based on the low confidence rating given to the emission factor, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (86 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 27 to 270 g TEQ/yr.

Whether releases from this source result in significant human exposure is questionable. If wood burning today is a major source of human exposure to CDD/CDFs, then the tissues of ancient humans (who relied on wood as a fuel source more so than do humans in industrialized settings today) should have CDD/CDF levels that are a substantial fraction of the levels found in humans today. The 1987 NHATS study indicates that the U.S. average CDD/CDF adipose tissue concentration is currently about 1000 ppt (U.S. EPA 1990d) and Schecter (1991) reports that the total CDD/CDF concentration in liver tissues today is about 400 ppt. However, with the exception of OCDD in one sample, Ligon et al. (1989) could not detect CDD/CDFs that exceeded the analytical background (detection limit = 0.3 to 5 ppt) in the mummified muscle tissues of nine 2,800 year old Chilean Indians. Similarly, Schecter (1991) examined the livers of two frozen 400 year old Eskimo women and found only HpCDD, OCDD, and HpCDF at levels only 15 percent of current levels.

Fiedler and Hutzinger (1992) estimate that 1.1 g of dioxin TEQ may be released to the atmosphere in Germany annually from residential combustion of coal. In the United Kingdom, combustion of coal by residential, industrial, and utility sources is estimated to account for 38 percent (1,489 g TEQ/yr) of all dioxin TEQ releases to the atmosphere (ECETOC, 1992). The Clean Air Act requires an assessment of the emissions of toxic air contaminants (including CDDs and CDFs) discharged from the stacks of coal-fired power plants. The EPA is collaborating with the U.S. Department of Energy to conduct this study. Stack testing at seven plants is currently underway and the results will be incorporated to the extent possible in the final version of this report.

In the United States, the consumption of coal accounts for approximately 25 percent of the energy consumed from all sources (U.S. DOC, 1992). In 1991, 806 million metric tons of coal were consumed in the United States (EIA, 1993). Of this total, 87 percent (or 701 million metric tons ) was consumed by electric utilities, 12.3 percent (or 99 million metric tons) was consumed by the industrial sector, and 0.7 percent (or 5.5 million metric tons) was consumed by residential and commercial sources (EIA, 1993). These production estimates are assigned a "high" confidence rating since they are based on detailed studies specific to the U.S.

Derivation of emission factors is difficult due to the extremely limited data on emission of dioxin-like compounds from coal-fired utility boilers (NATO,1988). Most investigations of emissions from facilities in the United States have not reported the detection of dioxin congeners at the exit to the stack (NATO, 1988). Therefore in the development of emission factors representative of coal-fired utility boilers (coal-fired power plants), the reported limit of detection of the analytical method was applied as an upper-bound to the plausibility of emissions of dioxin from the source. If it is assumed that burning one kg of coal in a modern power plant produces an estimated 6.2 dry standard cubic meters (dscm) of combustion gas, then the limit of detection from the study of U.S. power plants can be used to estimate an emission factor (NATO,1988). From the reported limit of detection of dioxin at the stack, an upper-bound emission factor for total CDD/CDF of 3.11E-02 µg/kg of coal combusted and an upper-bound emission factor for TEQ of 4.22E-04 µg/kg of coal combusted can be derived. If it is assumed that 700 billion kg of coal is combusted each year by power plants in the United States, then the upper-bound emission factors indicate an annual emission to the air of less than 2.2E+04 grams of total CDD/CDF and less than 3.0E+02 grams of TEQ/yr. The emission factors are assigned a "low" confidence rating because the stack emission of dioxin-like compounds from coal-fired utility boilers operating in the United States has yet to be determined. Emissions tests reported to date have not detected these compounds. The estimated emissions must, therefore, be considered the upper-bound of possible emissions from the source category based on available data.

The accidental or intentional combustion of PCBs in incinerators and boilers not approved for PCB burning (40 CFR 761) may produce CDDs and CDFs. At elevated temperature, such as those in transformer fires, PCBs can undergo reactions to form CDF and other by-products. Several accidental fires in the U.S. and Sweden which involved the combustion of PCBs and the generation of CDDs and CDFs are discussed in Hutzinger and Fiedler (1991b). For example, analyses of soot samples from a Binghamton, New York office building fire detected 20 µg/g of total CDDs (0.6 to 2.8 µg/g of 2,3,7,8-TCDD) and 765 to 2,160 µg/g of total CDFs with 12 to 270 µg/g of 2,3,7,8-TCDF. At that site, the fire involved the combustion of a mixture containing PCBs (65 percent) and chlorobenzene (35 percent). Hutzinger and Fiedler (1991b) also reported that laboratory analyses of soot samples from a PCB transformer fire which occurred in Reims, France indicated total CDD and CDF levels in the range of 4 to 58,000 ng/g and 45 to 81,000 ng/g, respectively.

The use of PCBs in new transformers in the United States has been banned and their use in existing transformers is being phased out. Because of the accidental nature of transformer fires it is not possible to accurately estimate annual emissions from this source.

The pyrolysis and photolysis of brominated phenolic derivatives and polybrominated biphenyl ethers used as flame retardants can generate polybrominated dibenzo-p-dioxins (BDDs) and dibenzofurans (BDFs) (Hutzinger and Fiedler, 1991a; Luijk et al., 1992; Watanabe and Tatsukawa, 1987). Watanabe and Tatsukawa (1987) observed the formation of BDFs from the photolysis of decabromobiphenyl ether. Approximately 20 percent of the decabromobiphenyl ether was converted to BDFs in samples that were irradiated with ultraviolet light for 16 hours (Watanabe and Tatsukawa, 1987). Decabromobiphenyl ether is used as a flame retardant in resins, textiles, and paints.

Luijk et al.(1990) studied the formation of BDD/Fs during the compounding/ extrusion of decabromodiphenyl ether into high-impact polystyrene polymer at 275° C. HpBDF and OBDF were formed during repeated extrusion cycles, and the yield of BDFs increased as a function of the number of extrusion cycles (Luijk et al., 1990). HpBDF increased from 1.5 to 9 ppm (in the polymer matrix) and OBDF increased from 4.5 to 45 ppm after four extrusion cycles.

Thoma and Hutzinger (1989) observed the formation of BDFs during combustion experiments with polybutylene-terephthalate polymers containing 9 to 11 percent decabromodiphenyl ether. Maximum formation of BDFs occurred at 400 to 600° C with a BDF yield of 16 percent. Although Thoma and Hutzinger (1989) did not provide specific quantitative results for similar experiments conducted with octabromodiphenyl ether and 1,2-bis(tri-bromophenoxy)ethane, they did report that BDDs and BDFs were formed.

Insufficient data are available upon which to derive annual BDD/BDF emission estimates from this source.

Granular activated carbon (GAC) is an adsorbent that is widely used to remove organic pollutants from wastewater and in the treatment of finished drinking water at water treatment plants. Activated carbon is manufactured from the heat treatment of nut shells and coal under pyrolytic conditions (Buonicore, 1992a). The properties of GAC make it ideal for adsorbing and controlling vaporous organic and inorganic chemicals entrained in combustion plasmas, as well as soluble organic contaminants in industrial effluents and drinking water. The high ratio of surface area to particle weight (e.g., 600 - 1600 m²/g), combined with the extremely small pore diameter of the particles (e.g., 15-25 Angstroms) increases the adsorption characteristics (Buonicore, 1992a). GAC will eventually become saturated and the adsorption properties will significantly degrade. When saturation occurs, the GAC usually must be replaced and discarded, which significantly increases the costs of pollution control. The introduction of carbon reactivation furnace technology in the mid-1980s created a method involving the thermal treatment of used GAC to thermolytically desorb the synthetic compounds and restore the adsorption properties for reuse (Lykins et al., 1987).

The used GAC can contain compounds that are precursors to the formation of CDDs/Fs during the thermal treatment process. The U.S. EPA measured precursor compounds in spent GAC used as a feed material to a carbon reactivation furnace tested during the National Dioxin Study (U.S. EPA, 1987). The total chlorobenzene content of the GAC ranged from 150 ppb to 6,630 ppb. Trichlorobenzene was the most prevalent species present, with smaller quantities of di- and tetra-chlorobenzenes detected. Total halogenated organics were measured to be about 150 ppm.

The U.S. EPA has stack tested two GAC reactivation furnaces for the emission of dioxin (U.S. EPA, 1987; Lykins et al., 1987). One facility was an industrial carbon reactivation plant, and the second facility was used to restore GAC at a municipal drinking water plant. The industrial carbon regeneration plant processed 36,000 kg/day of spent GAC used in the treatment of industrial wastewater effluents. Spent carbon was reactivated in a multiple-hearth furnace, cooled in a water quench and, stored and shipped back to primary chemical manufacturing facilities for reuse. The furnace fired natural gas, and consisted of seven hearths arranged vertically in series. The hearth temperatures ranged from 480° C to 1000° C. The spent GAC contained about 40 percent weight moisture. The used GAC was fed to the top hearth. In the furnace, the spent carbon was dried and the organics adsorbed onto the carbon particles were volatilized and burned in the heated combustion atmosphere. The regenerated carbon dropped from the bottom hearth of the furnace to a quench tank to reduce the temperature. Air pollutant emissions were controlled by an afterburner, a sodium spray cooler, and a fabric filter. Temperatures in the afterburner were about 930° C.

The second GAC reactivation facility tested by U.S. EPA consisted of a fluidized-bed furnace located at a municipal drinking water treatment plant (Lykins et al., 1987). The furnace was divided into three sections: a combustion chamber, a reactivation section and a dryer section. The combustion section was fired by natural gas, and consisted of a stoichiometrically balanced stream of fuel and oxygen. These expanding gases of combustion provided heat and suspended and fluidized the carbon. Temperatures of combustion were about 1,038° C. The reactivation section outside the combustion chamber allowed for the complete volatilization of the heated GAC. Off-gasses from the reactivation/combustion section were directed through an acid gas scrubber and high-temperature afterburner prior to discharge from a stack.

The industrial GAC reaction furnace test data indicate that an average of 5.87E-02 µg of CDD/CDF per kg of GAC incinerated may be emitted from the stack during operation (U.S. EPA, 1987). An average of 2.98E-03 µg TEQ per kg of GAC may be released to the air during operation. A "medium" confidence rating is given to these emission factors, because only one industrial GAC reactivation furnace operating in the United States has been stack tested. In the second GAC reactivation furnace tested by EPA (Lykins et al., 1987), measurable concentrations of dioxin-like compounds were detected in the stack emissions. When chlorine was used in pretreatment of the surface water for preliminary disinfection prior to filtration with GAC, the 2,3,7,8-TCDD congener was seen in the particulate stack emission discharges to the incinerator afterburner in low concentration [0.001-0.02 parts per trillion by volume (ppt/v)]. 1,2,6,7-TCDF was detected in two out of four stack tests in a concentration range of 0.004-0.02 ppt/v. When no chlorine was used to disinfect the surface water prior to filtration with GAC, no 2,3,7,8-TCDD was detected (<0.001 ppt/v). With the afterburner operating, no CDD congeners below HpCDD were detected in the stack emissions. Concentrations of HpCDDs and OCDD ranged from 0.001 to 0.05 ppt/v and 0.006 to 0.28 ppt/v, respectively. All congener groups of CDFs were detected in the stack emissions even with the afterburner operating. Total CDFs emitted from the stack averaged 0.023 ppt/v. Measurements of the individual CDD/CDF congeners were not performed, therefore it was not possible to derive emission factors for this facility.

The mass of GAC that is reactivated annually in carbon reactivation furnaces is not known. However, a crude estimate, which is given a "low" confidence rating, is the mass of virgin GAC shipped each year by GAC manufacturers. According to U.S. DOC (1990c), 48 thousand metric tons of GAC were shipped in 1987. Applying the emission factors developed above to this crude estimate of potential GAC reactivation volume, annual releases of 0.14 grams of TEQ and 2.8 grams of total CDD/CDF are estimated. Based on the "medium" confidence rating assigned to the emission factor, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual emissions (0.14 g TEQ/yr) is the geometric mean of this range, the range is calculated to be 0.06 to 0.3 g TEQ/yr.

Portland cement is a fine, grayish powder consisting of a mixture of four basic materials: lime (calcareous), silica (siliceous), alumina (argillaceous), and iron (ferriferous). Pyroprocessing in a rotary-type kiln plays a central role in fusing the basic raw materials into cement. The raw materials are ground into fine particles and are then either suspended in water to form a pumpable slurry (i.e., wet process) or are fed directly (i.e., dry process) into a rotary kiln for processing at elevated temperatures in an oxygen-enriched atmosphere. In the rotary kiln, evaporation of the water, calcination of the carbonate constituents, and fusion of the minerals occurs to form clinker. Clinker is a gray-colored, glass-hard material comprised of the cement minerals, dicalcium silicate, tricalcium silicate, calcium aluminate, and tetracalcium aluminoferrite. The clinker is then ground into a fine powder and mixed with gypsum to form portland cement. Approximately 1,575 kg of dry raw materials are needed to produce about 1000 kg of cement clinker (Greer et al., 1992). In 1991, the last year in which data is available, about 66 billion kg of cement clinker was produced in the United States by 212 portland cement kilns requiring about 103 billion kg of raw materials (U.S. EPA, 1993f; Greer et al.,1992).

Because of the relatively high combustion temperature required to produce cement clinker (1400° to 1510° C), coal or petroleum coke are typically used as the primary fuel to sustain combustion in the kiln. However, some cement kilns do burn hazardous liquid and solid waste as supplemental fuel to reduce the amount of coal that is purchased. It is estimated that 34 of the 212 existing cement kilns (i.e., 16 percent) burn hazardous waste as supplemental fuel (U.S. EPA, 1993f). Other types of non-hazardous liquid and solid wastes used as supplemental fuels include tires, waste oil, and wood chips. The most common air pollution control devices (APCD) employed on rotary kilns are those intended to control dust and particulate matter (i.e., fabric filters and/or electrostatic precipitators). Dioxins were first detected in stack emissions from portland cement kilns in the early 1980s (U.S. EPA, 1987; Peters, 1983; Branscome et al. 1984, 1985). Dioxin was detected only in low amounts and was thought to be caused by the co-firing of liquid hazardous waste with conventional fossil fuels (Peters, 1983). The EPA gave this source category a low priority for follow-up testing in EPA's National Dioxin Study conducted in 1985-1986 (U.S. EPA, 1987). Since then, the thermolytic reactions and the conditions favoring the formation of CDDs and CDFs in combustion processes have become better understood. (See Section 3.5). Some aspects of this theory warrant investigation into the formation of dioxin in portland cement kilns, including:

• Some primary combustion fuels (i.e., coal and petroleum coke) and fuel supplements (wood chips and tires) used to sustain elevated temperatures in the kiln to form clinker may also produce aromatic hydrocarbon compounds (e.g., benzene, phenol) that can later become chlorinated ring structures. The oxidation of HCl gas has been shown to provide chlorine available for ring substitution. In addition, chlorine has been measured directly in the combustion fuels to cement kilns (EER, 1993).

• The chlorinated aromatic compounds may act as precursor molecules to the thermalytic formation of CDD/CDFs on the active surface of carbonaceous particulates;

• De novo synthesis of CDD/CDFs on the active surface of carbonaceous particulates in the presence of a catalytic agent (e.g., a metal ion such as copper chloride);

• Post-kiln temperatures of the combustion gases in the APCD system are within the range of temperatures observed in laboratory studies that promote the continued formation of CDD/CDFs (i.e., 250° to 350° C); and

• Co-firing of liquid hazardous organic wastes with coal and petroleum coke may lead to an increase in the amount of CDD/CDFs formed in the post-combustion zone.

Currently, cement kilns that accept and burn hazardous waste as an auxiliary fuel are required under RCRA to characterize pollutant stack emissions, including emissions of CDD/CDFs. EPA's Office of Solid Waste is in the process of collecting and analyzing these emission reports to determine the extent and magnitude of CDD/CDF releases and the need for further regulation. Preliminary stack test data are available from 14 of the 34 cement kilns burning hazardous waste and from 3 of the 178 kilns not burning hazardous waste (EER, 1993; RTI, 1993). Table 3-35 is a summary of the available emissions data. For kilns accepting and burning hazardous waste as supplemental fuel, it appears that the concentration of dioxin in the stack gas (grams/dscm at 7 percent O₂) is highly variable. For example, the average stack emissions of total CDD/CDF for individual kilns range from 2 to 2000 ng/dscm, a thousand-fold difference. There appears to be no consistent pattern to the relationship of total CDD/CDFs to the estimated dioxin TEQ, indicating a wide variability in the distribution of toxic congeners of CDDs and CDFs in the emissions. For example, the ratio of total CDD/CDFs to the TEQ ranges from about a factor of 5:1 to a factor of 1000:1, indicating that some kilns have a congener distribution skewed toward the lower chlorinated more toxic congeners and others are skewed toward the higher chlorinated less toxic congeners. Cement kilns which do not burn hazardous waste as supplemental fuel appear to be less variable. However, this observation must be tempered by the fact that fewer of these kilns have been stack tested.

The limited emission data suggests that the average stack concentrations of CDD/CDFs are about eight times higher among the kilns burning hazardous waste than those that do not. As discussed below, a similar relationship was seen in the cement kiln dust samples from these two categories of kilns. On this basis, it was decided that separate emission factors should be developed for the kilns burning hazardous waste and those that do not. Given the limited emission test data, especially among the kilns that do not burn hazardous waste, clearly more testing is needed to confirm this difference in emission factors.

National estimates of air emissions of dioxin TEQ/yr from all operating cement kilns were made using two different methods:

1. Dioxin emissions correlate with the total mass of materials processed and burned at the kiln to form clinker (i.e., related to the kiln throughput); and

2. Dioxin emissions correlate with the total energy content of the fuel (including hazardous waste) used to sustain combustion in the kiln.

Although these two methods would likely generate different emission estimates if site-specific data (i.e., throughput and energy consumption data) were available, such data were not available for this report and, therefore, the use of generic industry average data resulted in identical estimates.

For the first method, three values must be known in order to calculate annual dioxin TEQ emissions: (a) the average concentration of dioxin TEQ in the stack gas (g TEQ/dscm); (b) the average volume of combustion gas evolved per kg of material fed to the kiln; and (c) an estimate of the total dry weight of materials processed by all operating cement kilns (kg/yr). Average dioxin TEQ stack emission concentrations are presented in Table 3-35 for kilns burning and not burning hazardous waste as supplemental fuel. The averaging gave equal weighting to all tested kilns. The average dioxin TEQ stack concentrations are 7.1 ng/dscm and 0.9 ng/dscm for kilns burning hazardous waste and not burning hazardous waste, respectively. A reasonable estimate of combustion gas volume/kg of materials processed in the kiln is 1.75 dscm/kg (RTI, 1993). This approach suggests an emission factor of 12.4 ng TEQ/kg and 1.6 ng TEQ/kg for kilns burning and not burning hazardous waste, respectively.

Greer et. al (1992) estimated that the ratio of the dry weight of materials charged into the kiln to the weight of dry clinker produced is 1.575:1. Therefore, if 66 billion kg of clinker were produced by 212 cement kilns in 1991 (U.S. EPA, 1993f), then 104 billion kg of raw materials were consumed in that year. A final assumption is that the annual throughput of raw materials processed is roughly proportional to the number of kilns in the class of cement kilns (i.e., the number burning hazardous waste versus the number of facilities not burning hazardous waste). Multiplying the emission factors by the raw material throughput yields annual emission estimates of 210 g TEQ for kilns burning hazardous waste and 140 g TEQ for kilns not burning hazardous waste.

For the second method, it is assumed that the dioxin emissions are better calculated on the basis of annual fuel consumption rather than the annual amount of materials processed by the kiln. Given the diversity and mixtures of fuel types typically used (coal, coke, liquid hazardous waste, natural gas, oil, wood chips, tires), a good measurement of total fuel consumption is the amount of total energy consumed to produce the clinker (U.S. EPA, 1993f). Johnson (1992) estimated that 71 trillion kcals were consumed in the year 1991 by all operating portland cement kilns. Dividing this value by the total kg of raw material processed (is 104 billion kg, as reported above) yields an average energy usage of 680 kcal/kg of raw material. Dividing this value into the emission factors derived above (in units of ng of dioxin TEQ/kg of raw material) yields emission factors in terms of ng of dioxin per kcal. This approach suggests an emission factor of 2.3 pg of TEQ/kcal for kilns not burning hazardous waste and 18 pg of TEQ/kcal for kilns which do burn hazardous waste. Finally, it is assumed that the total energy usage can be apportioned between kilns burning hazardous waste (34 of 212 kilns) and those that do not (178 of 212 kilns) on the basis of the number of kilns in each group. Multiplying these energy use estimates by the energy based emission factors yields the following estimates of annual emissions: 210 g TEQ for kilns burning hazardous waste and 140 g TEQ for kilns not burning hazardous waste.

Both of these methods suggest that annual dioxin emissions to air from all cement kilns combined is about 350 grams TEQ. The estimated TEQ emission factors used to derive this best estimate of annual TEQ emissions are given a "low" confidence rating because the mechanisms giving rise to dioxin emissions from cement kilns are largely unknown; very few of the existing facilities have been stack tested for emissions; and because of the apparent high variability in the ratio of total CDD/CDFs to dioxin TEQ and the apparent high variability in emissions between tested facilities as shown by the large standard deviation in Table 3-35. The "production" estimate of annual raw material throughput is given a "high" confidence rating because it is based on recent survey data. Based on these confidence ratings, the estimated range of potential emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the lowest estimate of annual emissions (350 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 110 to 1,100 g TEQ/yr.

In a recent Report to Congress (U.S. EPA, 1993f), EPA's Office of Solid Waste establishes the factual basis for its decision making regarding the appropriate regulatory status, under RCRA, of cement kiln dust (CKD) waste. To aid EPA in their study, the Portland Cement Association (PCA) conducted a survey in 1991 of cement manufacturers. Survey responses were received from 64 percent of the active cement kilns in the United States. Based on the survey responses, EPA estimated that the U.S. cement industry generated about 12.9 million metric tons of gross CKD and 4.6 million metric tons of "net CKD", of which 4.2 million metric tons was land disposed, in 1990. The material collected by the APCD system is called "gross CKD" (or "as generated" CKD). The gross CKD is either recycled back into the kiln system or is removed from the system for disposal (i.e., "net CKD" or "as managed" CKD) (U.S. EPA, 1993f).

Also in support of the Report to Congress, EPA conducted sampling and analysis during 1992 and 1993 of CKD and clinker. The purposes of the sampling and analysis efforts were: (1) to characterize the CDD/CDF content of clinker and CKD ; (2) to determine the relationship, if any, between the CDD/CDF content of CKD and the use of hazardous waste as fuel; and (3) to determine the relationship, if any, between the CDD/CDF content of CKD and the use of wet versus dry process cement kilns. Clinker samples from 9 kilns and CKD samples from 11 kilns (six of which burn hazardous waste) were analyzed (U.S. EPA, 1993f).

CDD/CDFs were not detected in any of the clinker samples. Tetra- through octa-chlorinated CDDs and CDFs were detected in the "gross CKD" samples obtained from 10 of the 11 kilns and in the "net CKD" samples obtained from 8 of the 11 kilns. The CDD/CDF content of "gross CKD" ranged from 0.008 to 247 ng TEQ/kg and for "net CKD" the content ranged from 0.045 to 195 ng TEQ/kg. Analyses for seven PCB congeners were also conducted but no congeners were detected in any clinker or CKD sample. TCLP leachate testing of the CKD samples from six kilns showed no leaching of CDD/CDFs (detection limits ranged from 3 to 37 pg/L) except for OCDD in two samples (110 and 170 pg/L). Statistical analysis of the results indicated that mean CDD/CDF concentrations in "net CKD" generated by the sampled kilns burning hazardous waste are higher (35 ng/kg) than in "net CKD" generated by the sampled facilities not burning hazardous waste (3.0E-02 ng/kg). These calculations of mean values treated not detected values as zero. If the not detected values had been excluded from the calculation of the means, then the mean value for "net CKD" from kilns burning hazardous waste would increase by a factor of 1.2 and the mean value for "net CKD" from kilns not burning hazardous waste would increase by a factor of 1.7. One sampled kiln had CDD/CDF concentrations more than two orders of magnitude greater than the TEQ levels found in samples from any other kiln. If this kiln is considered to be atypical of the industry (U.S. EPA, 1993f) and is not included in the calculation, then the mean "net CKD" concentration for hazardous waste burning kilns decreases to 2.9 ng/kg.

From these data, an estimate of dioxin TEQ emissions to land in the form of land-disposed "net CKD" can be made. The estimate of land-disposed CKD from the 1991 PCA Survey, 4.2 million metric tons per year (basis year is 1990), was divided among kilns burning hazardous waste (34 kilns) and those which do not (178 kilns) on the basis of the number of kilns in each category. The average TEQ concentration in the net CKD from kilns burning hazardous waste, including the potentially non-typical kiln, was 35 ng TEQ/kg. For kilns which do not have hazardous waste the average concentration in the "net CKD" was 3.0E-02. Multiplying these average concentrations by the annual "net CKD" production, yields estimates of 24 g TEQ/yr for kilns burning hazardous waste and 0.1 g TEQ/yr for kilns not burning hazardous waste, yielding a total of 24.1 g TEQ/yr for all kilns. The "production" estimate was assigned a "high" confidence rating because it is based on recent EPA survey data. The "emission factor" estimates are assigned a "low" confidence rating because the sampling data upon which they are based showed high variability among the 11 kilns sampled (out of 212 kilns in the United States). Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (24.1 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 7.6 to 76 g TEQ/yr.

Although discussed as potential sources in the preceding sections, insufficient data are available upon which to develop emission factors for the primary ferrous and primary nonferrous metal refining/smelting industries. In addition, although emission estimates were developed in Section 3.6.13 for diesel-powered on-highway vehicles (i.e., the largest use of diesel fuel in the United States), because of insufficient data no estimates were developed for off-highway transportation diesel engine fuel use (including railroad engine fuel and fuel for agricultural machinery) or for diesel fuel use by the commercial and industrial sectors of the economy. Also, although no estimated emissions could be developed for residential oil, gas, and charcoal combustion because of lack of emission rate factors, these have been identified as potential sources by Harrad et al. (1992a, 1992b) and Fiedler and Hutzinger (1991b).

It is very difficult to estimate CDD/CDF releases that may be occurring from reservoir sources. However, some idea of the potential magnitude of these emissions can be gained by estimating the size of the overall reservoir. Equation 1 calculates the concentration of a contaminant in a reservoir given the deposition rate of the contaminant into the reservoir and the rate of dissipation from that reservoir:

where C is the concentration after time t, DEP is the deposition rate (in units of mass/area-time), k is the first order dissipation rate (time^-1), and MIX is the mass of the reservoir into which DEP mixes (mass units, corresponding to area of DEP). Consider the case where DEP has been occurring for a number of years at a steady rate. A question that might be asked is, what is the contribution of a year's worth of deposition to the amount that is already there. This can be estimated with a ratio of estimated concentrations C₁/C₂ using Equation (1), where C₁ is the concentration after the number of years, and C₂ is the year's worth of deposition. Assuming DEP, MIX, and k are constant, the ratio of C₁/C₂ reduces to:

where t₁ is the number of years that DEP has been occurring, and t₂ is equal to 1 for the one year's worth of deposition. For the sake of this discussion, if one assumes that DEP for dioxin-like compounds has been steady since the 1940s, and one wants to evaluate a year's worth of deposition in the 1990s, then t₁ equals about 50 years. A dissipation rate of 0.0693, corresponding to a 10-year half-life was used for atmospheric deposition onto soils for the methodologies described in Volume III. A t₁ of 50 and k of 0.0693 applied to Equation (2) yields a ratio of about 14.5. This means that there would be about 14.5 times more contaminant in the reservoir than one year's contribution.

However, the half-life of 10 years is probably too low for this exercise. This half-life was generated from data on 2,3,7,8-TCDD applied to experimental plots as 2,4,5-T in Agent Orange testing (Young, 1983). This might be appropriate for dissipation from a bounded area of high soil contamination, where dissipation mechanisms such as soil erosion, dust resuspension, or volatilization might be occurring. However, if MIX is considered as the total reservoir of soil and surface vegetation, then losses from a bounded area are unlikely to translate to losses from the larger system. In other words, a more appropriate half-life for this discussion might be more like 50 years than 10 years. If 50 years is assumed in the above exercise, than the ratio increases to 36.

This analysis suggests that dioxin-like compounds already in the reservoir source may exceed annual contributions to the reservoir source by 15 or more times. The potential for emissions from this large source is uncertain. Dioxin-like compounds that accumulate in deep sediments or become buried in the soil are not likely to contribute to current emissions. However, those located near the surface could become re-entrained into the air or water bodies.

As discussed in Section 3.1, several investigators have attempted to conduct "mass balance" checks on the estimates of national dioxin releases to the atmosphere. Basically, this procedure involves comparing estimates of the emissions to estimates of aerial deposition. Such studies in Sweden and Great Britain have suggested that the deposition exceeds the estimated emissions by about 10-fold. These studies are acknowledged to be quite speculative due to the strong potential for inaccuracies in the emission and/or deposition estimates. In addition, the apparent discrepancies could be explained by long range transport from outside the country, resuspension and deposition of reservoir sources, atmospheric transformations, or unidentified sources. Bearing these limitations in mind, this procedure has been used below to compare the estimated emissions and deposition in the U.S.

Koester and Hites (1992) measured both wet and dry deposition of CDD/CDFs at two locations in Indiana. These measurements indicate a range of 370 to 540 ng/m²-yr on a total CDD/CDF basis. If perfect congener distribution is assumed for the Koester and Hites measurements, these total deposition rates correspond to about 1 to 2 ng TEQ/m²-yr. Based on studies of temporal trends in CDD/CDF concentrations in sediments from Green Lake, a non-industrially impacted lake near Syracuse, New York, Smith et al. (1993) have calculated the total CDD/CDF atmospheric deposition rate to be 375 ng/m²-yr for the period 1986 to 1990 which is very similar to that reported by Koester and Hites (1992). Andersson et al. (1992) estimated a deposition rate of 1 ng TEQ/m²-yr on the basis of snow measurements in Northern Sweden.

Fernandez et al. (1992) measured the wet and dry deposition over a period of one month at an urban/semiurban location in Great Britain. Assuming that the deposition measured over the one month period is representative of an entire year, then the rate of deposition is 13 ng TEQ/m²-yr (setting nondetects equal zero) or 17 ng TEQ/m²-yr (setting nondetects at half the detection limit). Van Jaarsveld and Schutter (1992), using long range transport modelling, estimate that national average deposition rates in Northern European countries are in the range of 1 to 10 ng TEQ/m²-yr.

In 1992, Hiester et al. (1993) collected two month duration deposition samples in seven urban and one rural location in Germany. The total CDD/CDF deposition rates ranged, on an annualized basis, from 246 to 1,687 ng/m²-yr at the urban locations and 420 ng/m²-yr at the rural location. The TEQ deposition rates ranged from 3.6 to 30.3 ng TEQ/m²-day at the urban locations and 4.4 ng TEQ/m²-day at the rural location. For these calculations, Hiester et al. (1993) set nondetects equal to zero.

Liebl et al. (1993) reported the results of long term deposition measurements at three locations in Germany: a rural region with some industry, a rural "background" site, and an industrial/living area site. The deposition rates, based on 11 months of measurement in 1992, were reported as 1.1 ng TEQ/m²-yr for the rural background site, 1.5 ng TEQ/m²-yr for the rural/industrial site (estimated from figure in the report), and 7.6 ng TEQ/m²-yr for the industrial/living area site (estimated from figure in the report).

Fiedler (1993) reports that the average deposition rate for rural areas in Germany (defined as agriculture, forest, and water) is 4.4 TEQ/m²-yr with a range of 1.8 to 7.3 ng TEQ/m²-yr. The estimated deposition rate in industrialized areas typically ranges from 7.3 to 36 ng TEQ/m²-yr with even higher deposition rates (up to 1,000 ng TEQ/m²-yr) in a small number of areas.

Broman et al. (1991) collected a two-month duration deposition sample at a remote open coastal area in Sweden and reported a deposition rate for total CDD/CDFs of 38 ng/m²-yr. Näf et al. (1992) measured the flux of CDD/CDFs to aquatic sediments at two remote background sites; one in the Bothnian Sea and the other in the Baltic Sea. The measured rates for total CDD/CDFs ranged from 240 to 1,200 ng/m²-yr and for TEQ ranged from 3 to 14 ng TEQ/m²-yr. Atmospheric deposition was presumed by the authors to be the principal source of the measured flux because the sampling sites were located far from coastal areas, freshwater input sources, and industrial input.

For purposes of generating a preliminary estimate of CDD/CDF deposition in the United States, it is assumed that the deposition rate of 1 ng TEQ/m²-yr measured in Sweden applies to Alaska (land area = 1.5 x 10¹² m²). The average deposition rate for the continental United States (land area = 7.8 x 10¹²m²) could be as low as 2 ng TEQ/m²-yr (based on limited U.S. data) or perhaps about 6 ng TEQ/m²-yr (based on European data). Based on these assumptions, total U.S. deposition can be estimated as about 20,000 to 50,000 g TEQ/yr. This range can be compared to the range of estimated annual air emissions in the United States, 3,300 - 26,000 g TEQ/yr, as presented in Table 3-2. As noted above however, making and interpreting such comparisons is highly speculative considering the very limited data on emissions and deposition.

A and D International (1991) Letter from A and D International to USEPA regarding information on the enclosed protocols and data on polychlorinated dibenzo-p-dioxins/dibenzofurans with attachments, April 13, 1990.

Addink, R.; Van Bavel, B.; Visser, R.; Wever, H.; Slot, P.; Olie, K. (1990) Surface catalyzed formation of polychlorinated dibenzo-p-dioxins/dibenzofurans during municipal waste incineration. Chemosphere 20(10-12):1929-1934.

Addink, R.; Drijver, D.J.; Olie, K. (1991) Formation of polychlorinated dibenzo-p-dioxins/dibenzofurans in the carbon/fly ash system. Chemosphere 23(8-1):1205-1211.

Ahling, B.; Bjorseth, A.; Lunde, G. (1978) Formation of chlorinated hydrocarbons during combustion of poly vinyl chloride. Chemosphere 7:799-806.

Ahling, B.; Lindskog, A. (1982) Emission of chlorinated organic substances from combustion. In: Hutzinger, O.; Safe, S., eds, Chlorinated dioxins and related compounds, impact on the environment. New York, NY: Pergamon Press.

Andersson, P.; Marklund, S.; Rappe, C. (1992) Levels and profiles of PCDDs and PCDFs in environmental samples as determined in snow deposited in northern Sweden. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Aittola, J.; Paasivirta, J.; Vattulainen, A. (1992) Measurements of organochloro compounds at a metal reclamation plant. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Bacher, R.; Swerev, M.; Ballschmiter, K. (1992) Profile and pattern of monochloro- through octachlorodibenzodioxins and -dibenzofurans in chimney deposits from wood burning. Environ. Sci. Technol. 26(8):1649-1655.

Ballschmiter, K.; Buchert, H.; Niemczyk, R.; Munder, A.; Swerev, M. (1986) Automobile exhausts versus municipal waste incineration as sources of the polychloro-dibenzodioxins (PCDD) and -furans (PCDF) found in the environment. Chemosphere 15(7):(901-915).

Battelle (1992a) Determination of polybrominated dibenzo-p-dioxins and polybrominated dibenzofurans in octabromodiphenyloxide (sponsored by Ethyl Corporation). Amended final report.

Battelle (1992b) Determination of polybrominated dibenzo-p-dioxins and polybrominated dibenzofurans by HRGC/MRMs in octabromodiphenyloxide (sponsored by Great Lakes Chemial Corporation). Final Report.

Battelle (1993) Determination of polybrominated dibenzo-p-dioxins and polybrominated dibenzofurans by high resolution gas chromatography/medium/high resolution mass spectrometry in octabromodiphenyloxide (sponsored by Ameribrom, Inc.). Amended Final Report.

Beard, A.; Krishnat, P.N.; Karasek, F.W. (1993) Formation of polychlorinated dibenzofurans by chlorination and de novo reactions with FeCl₃ in petroleum refining processes. Environ. Sci. Technol. 27(8):1505-1511.

Berenyi, E.B.; Gould, R.N.(1993) Municipal waste combustion in 1993. Waste Age 24:51-56.

Berenyi, E.B. (1993) A decade of municipal waste combustion in the United States: prospects and problems. In: Municipal waste combustion, proceedings of an international specialty conference of the Air and Waste Management Association. Williamsburg, Va; March 1993. pp. 51-66.

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.

Bingham, A.G.; Edmunds, C.J.; Graham, B.W.; Jones, M.T. (1989) Determination of PCDDs and PCDFs in car exhaust. Chemosphere 19(1-6):669-673.

Boschi, G.; Cocheo, V.; Giannandrea, G.; Magagni, A. (1992) Hospital and municipal solid waste incinerators: Correlation between PVC present in waste feed and emitted organic micropollutants. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Branscome, M., et al. 1984. Evaluation of waste combustion in a dry-process cement kiln at Lone Star Industries, Oglesby, Illinois. EPA Contract No. 69-02-3149., December 1984.

Branscome, M, et al. 1985. Evaluation of waste combustion in a wet-process cement kiln at General Portland, Inc., Paulding, OH. EPA Contract No 68-02-3149. February 1985.

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.

Bruce, K.R.; Beach, L.O.; Gullet, R.K. (1991) The role of gas-phase Cl₂ in the formation of PCDD/PCDF during waste combustion. Waste Management 11:97-102.

Brunner, C.R. (1984) Incineration systems, selection and design. New York, NY: Van Nostrand Reinhold Co.

Brunner, C.R. (1992) Sewage sludge. In: Buonicore, A.J., Davis, W.T., eds., Air pollution engineering manual. Air and Waste Management Association. New York, NY: Van Nostrand Reinhold. pp. 296-311.

Bumb, R.R.; Crummett, W.B.; Cutie, S.S.; Gledhill, J.R.; Hummel, R.H.; Kagel, R.O.; Lamparski, L.L.; Luoma, E.V.; Miller, D.L.; Nestrick, T.J.; Shadoff, L.A.; Stehl, R.H.; Woods,J.S. (1980) Trace chemistries of fire: a source of chlorinated dioxins. Science 210(4468):385-390.

Buonicore, A.J. (1992a) Control of gaseous pollutants. In: Buonicore, A.J., Davis, W.T., eds., Air pollution engineering manual. Air and Waste Management Association. New York, NY: Van Nostrand Reinhold.

Buonicore, A.J. (1992b) Medical waste incineration. In: Buonicore, A.J., Davis, W.T., eds., Air pollution engineering manual, Air and Waste Management Association. New York, NY: Von Nostrand Reinhold.

Burton, B.K.; Kiser, J.V.L. (1993) Status and directions of the U.S. municipal waste combustion industry. In: Municipal waste combustion, proceedings of an international specialty conference of the Air and Waste Management Association. Williamsburg, Va.; March 1993. pp. 3-11.

Buser, H.R. (1979) Formation of polychlorinated dibenzofurans (PCDFs), and dibenzo-p-dioxins (PCDDs) from the pyrolysis of chlorobenzenes. Chemosphere 8:415-424.

Buser, H.R. (1992) Identification and sources of dioxin-like compounds: I. polychloro-dibenzothiophenes and polychlorothianthrenes, the sulfur-analogues of the polychlorodibenzofurans and polychlorodibenzodioxins. Chemosphere 25(1-2):45-48.

Buser, H.R.; Dolezal, I.S.; Wolfensberger, M.; Rappe, C. (1991) Polychloro- dibenzothiophenes, the sulfur analogues of the polychlorodibenzofurans identified in incineration samples. Environ. Sci. Technol. 25(9):1637-1643.

Capes, C.E. (1983) Size enlargement. In: Kirk-Othmer Encyclopedia of Chemical Technology. 3rd ed. 21:77-105. New York, NY: John Wiley and Sons.

CARB (1987) California Air Resources Board. Determination of PCDD and PCDF emissions from motor vehicles. Draft report. October 1987.

CARB (1990a) California Air Resources Board. Technical support document to proposed dioxin control measures for medical waste incinerators. May 25, 1990.

CARB (1990b) California Air Resources Board. Evaluation on a woodwaste fired incinerator at Koppers Company, Oroville,California. Test report No. C-88-065. Engineering Evaluation Branch Monitoring and Laboratory Division. May 29, 1990.

CARB (1991) California Air Resources Board. Emissions test of the Modesto Energy Project tires-to-energy facility. State of California Air Resources Board (CARB), Engineering Evaluation Branch, Monitoring and Laboratory Division. Project # C-87-072, May 24, 1991.

Cash, G. (1993) Document on dioxin/furan data call-ins [memorandum to John Schaum, U.S. EPA, Office of Research and Development]. Washington, D.C.: U.S. Environmental Protection Agency, Office of Pollution Prevention and Toxics. May 4, 1993.

CEQ (1988) Environmental Quality 1987-1988. Municipal solid waste. Washington, DC: Council on Environmental Quality. Executive Office of the President of the United States.

CEQ (1990) Environmental Quality 1990. Forest fire damage and reforestation, 1930-1990. 21st Annual Report. Washington, DC: Council on Environmental Quality. Executive Office of the President of the United States.

ChemRisk (1993) Critique of dioxin factories. Prepared for: The Vinyl Institute, September 1993.

Chen, Y.R.; Chen, C.J.; Lee, C.C.; Chuong, C.Y. (1986) Reproductive effects of emissions from waste electrical products combustion. Presented at: Dioxin '86, 6th International Symposium on Chlorinated Dioxins and Related Compounds; Fukuoka, Japan.

Choudhry, G.G.; Hutzinger, O. (1983) Mechanistic aspects of the thermal formation of halogenated organic compounds including polychlorinated dibenzo-p-dioxins. New York, NY: Gordon and Breach.

Christmann, W.; Klöppel, K.D.; Partscht, H.; Rotard, W. (1989) Tetrachlorobenzoquinones, a source of PCDD/PCDF. Chemosphere 18(1-6):789-792.

Clement, R.E.; Tosine, H.M; Osborne, J.; Ozvacic, V.; Wong, G. (1985a) Levels of chlorinated organics in a municipal incinerator. In: Keith, L.H.; Rappe, C.; Choudhary, G., eds., Chlorinated dioxins and dibenzofurans in the total environment II. Boston, MA: Butterworth Publishers.

Clement, R.E.; Tosine, H.M.; Ali, B. (1985b) Levels of polychlorinated dibenzo-p-dioxins and dibenzofurans in wood burning stoves and fireplaces. Chemosphere, 14:815-819.

Clement, R.E.; Tosine, H.M.; Osborne, J.; Ozvacic, V.; Wong, G. (1988) Gas chromatographic/mass spectrometric determination of chlorinated dibenzo-p-dioxins and dibenzofurans in incinerator stack emissions and fly-ash: a 13-test study. Biomedical and Environmental Mass Spectrometry 17:81-96.

Clement, R.E.; Boomer, D.; Nalkwadi, K.P.; Karasek, F.W. (1990) Study of the relationship between trace elements and the formation of chlorinated dioxin on flyash. In: Clement, R.E.; Kagel, R.O., eds, Emissions From Combustion: Origin, Measurement, Control. Lewis Publishers, Inc., Chelsea, MI., pp 57-64.

Clement, R.E.; Tashiro, C. (1991) Forest fires as a source of PCDD and PCDF. Presented at: Dioxin '91, 11th International Symposium on Chlorinated Dioxins and Related Compounds; Research Triangle Park, NC; September 1991.

Cleverly, D.H. (1984) Chlorinated dibenzo-p-dioxins and furans in incineration of municipal solid waste. Proceedings of workshop on energy from municipal waste research: A technical review of thermochemical systems, February 22-24,1984, Kissimmee, FL. Argonne National lab, U.S. Department of Energy, pp: 295-319.

Cleverly, D.H.; Morrison, R.M.; Riddle, B.L.; Kellam, R.G. (1991) Regulatory analysis of pollutant emissions, including polychlorinated dibenzo-p-dioxins (CDDs) and dibenzofurans (CDFs), from municipal waste combustors. In: Hattemer-Frey, H.A.; Travis,C., eds., Health effects of municipal waste incineration. Boca Raton,FL: CRC Press, pp.47-65.

Commoner, B; McNamara, M; Shapiro, K; Webster, T. (1984). Environmental and economic analysis of alternative municipal solid waste disposal technologies, part II: The origins of chlorinated dioxins and dibenzofurans emitted by incinerators that burn unseparated municipal solid waste, and an assessment of methods of controlling them. Center for the Biology of Natural Systems, Queens College, CUNY, Flushing, New York. December 1, 1984.

Commoner, B.; Webster, T., Shapiro, K; McNamara, M. (1985) The origins and methods of controlling polychlorinated dibenzo-p-dioxin and dibenzofuran emissions from MSW incinerators. Presented at: 78th Annual Meeting of the Air Pollution Control Assoc.; Detroit, MI; June 1985.

Commoner, B.; Shapiro, K.; Webster, T. (1987) The origin and health risks of PCDD and PCDF. Waste Management and Research 5: 327-346.

Commoner, B. (1990) Making peace with the planet. New York, NY: Random House, Inc.

Crummett, W.B. (1982) Environmental chlorinated dioxins from combustion - the trace chemistries of fire hypothesis. In: Hutzinger, O.; Frei, R.W.; Merian, E.; Pocchiari, F., eds. Chlorinated dioxins and related compounds: impact on the environment. New York NY: Pergamon Press.

Curlin, L.C.; Bommaraju, T.V. (1991) Alkali & chlorine products. In: Kirk-Othmer Encyclopedia of Chemical Technology. 4th ed. pp. 938-1025.

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.

Czuczwa, J.M; Hites, R.A. (1985). Historical record of polychlorinated dioxins and furans in Lake Huron sediments. In: Keith, L.H.; Rappe. C.; Choudhary, G, eds. Chlorinated dioxins and dibenzofurans in the total environment II. Boston, MA: Butterworth Publishers.

Czuczwa, J.M.; Hites, R.A. (1986). Airborne dioxins and dibenzofurnas: Sources and fates. Environ. Sci. Technol. 20: 195-200.

Day, D.R.; Cox, L.A.; Mournighan, R.E. (1984) Evaluation of hazardous waste incineration in a lime kiln: Rockwell Lime Co. EPA-600/S2-84-132.

Dempsey, C.R.; Oppelt, E.T. (1993) Incineration of hazardous waste: A critical review update. Air & Waste 43: 25-73.

Dickson, L.C.; Karasek, F.W. (1987) Mechanism of formation of polychlorinated dibenzo-p-dioxins produced on municipal incinerator flyash from reactions of chlorinated phenols. J. Chromatography 389:127-137.

Dickson, L.C.; Lenoir, D.; Hutzinger, O. (1992) Quantitative comparison of de novo and precursor formation of polychlorinated dibenzo-p-dioxins under simulated municipal solid waste incinerator post combustion conditions. Environ. Sci. Technol. 26:1822-1828.

Donnelly, J.R. (1992) Waste incineration sources: refuse. In: Buonicore, A.J.; Davis, W.T., eds, Air pollution engineering manual. Air and Waste Management Association. New York, NY: Van Nostrand Reinhold. pp. 263-275.

Domalski, E.S.; Churney, K.L.; Ledford, A.E.; Bruce, S.S. (1986) Monitoring the fate of chlorine from MSW sampling through combustion. Part 1: Analysis of the waste stream for chlorine. Chemosphere 15(9-12):1339-1354.

Environment Canada. (1985) The national incineration testing and evaluation program: Two-stage combustion (Prince Edward Island). Environment Canada, Ottawa, Ontario, Canada. Report EPS 3/UP/1. September 1985.

ECETOC (1992) European Centre for Ecotoxicology and Toxicology of Chemicals. Exposure of man to dioxins: a perspective on industrial waste incineration. Technical Report No. 49. Brussels, Belgium.

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

EER. Energy and Environmental Research Corp. (1993) Engineering analysis and technical support to control metal and organic emissions from combustion of hazardous waste in incinerators and BIFs. Washington, DC: Office of Solid Waste. EPA Contract No. 68-C0-0094, Work Assignment No. 2-6.

Eitzer, B.D.; Hites, R.A. (1987) Reply to comment on "Airborne dioxins and dibenzofurans: sources and fate." Environ. Sci. Technol. 21:924.

EIA (1991) Energy Information Agency. Estimate of U.S. biofuels consumption 1990. Washington, D.C.: U.S. Department of Energy, Office of Coal, Nuclear, Electric, and Alternative Fuels. DOE/EIA-0548(90).

EIA (1992) Energy Information Agency. Fuel oil and kerosene sales 1991. Washington, D.C.: U.S. Department of Energy, Office of Oil and Gas. DOE/EIA-0535(91).

EIA (1993) Energy Information Agency. Monthly energy review - January 1993. Washington, D.C.: U.S. Department of Energy, Office of Energy Markets and End Use. DOE/EIA-0035(93/01).

Esposito, M.P.; Tiernan, T.O.; Drydent, F.E. (1980). Dioxins. Prepared for the U.S. Environmental Protection Agency, Cincinnati, OH. November, 1980. EPA-600/2-9-80-197.

Federal Register (1982) F.R. (October 29) 47:49322.

Federal Register (1985) Regulation of fuels and fuel additives; Gasoline lead content: final rule and supplemental notice of proposed rulemaking. F.R. (March 7) 50:9386-9408.

Federal Register (1987a) Pentachlorophenol; amendment of notice of intent to cancel registrations. F.R. (January 2) 52:140-148.

Federal Register (1987b) Final determination and intent to cancel and deny applications for registrations of pesticide products containing pentachlorophenol (including but not limited to its salts and esters) for non-wood uses. F.R. (January 21) 52:2282-2293.

Federal Register (1987c) Appendix A to part 60 (Amended). F.R. (December 16) 52:47853-47856.

Federal Register (1990) National sewage sludge survey: availability of information and data, and anticipated impacts on proposed regulation. F.R. (November 9) 55:47210-47283.

Federal Register (1991) Standards of performance for new stationary sources and final guidelines for municipal waste combustors. F.R. (February 11) 56:5488-5527.

Federal Register (1993a) Effluent limitations guidelines, pretreatment standards, and new source performance standards: pulp, paper, and paperboard category; national emission standards for hazardous air pollutants for source category: pulp and paper production; proposed rule. F.R. (December 17) 58:66078-66216.

Federal Register (1993b) Standards for the use or disposal of sewage sludge; final rules. F.R. (February 19) 58:9248-9404.

Fernandez, A.R.; Bushby, B.R.; Faulkner, J.E.; Wallace, D.S.; Clayton, P.; Davis, O.J. (1992) The analysis of toxic organic micropollutants in ambient air and atmospheric deposition. Chemosphere 25:1311-1316.

Fiedler, H.; Hutzinger, O. (1992) Sources and sinks of dioxins: Germany. Chemosphere, 25:1487-1491.

Fiedler, H. (1993) Formation and sources of PCDD/PCDF. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Freiter, E.R. (1979) Chlorophenols. In: Kirk-Othmer Encyclopedia of Chemical Technology, 3rd ed. pp. 864-872.

Fries, G.F.; Paustenbach, D.J. (1990) Evaluation of potential transmission of 2,3,7,8-tetrachlorodibenzo-p-dioxin-contaminated incinerator emissions to humans via foods. Journal of Toxicology and Environmental Health 29:1-43.

Funk, S. (1994) Summary of OPP reviews performed (1993 through FEbruary 1994) of dioxin data call-in analytical studies. Washington, DC: U.S. Enviornmental Protection Agency, Office of Pesticide Programs. March 14, 1994.

GAO (1990) U.S. General Accounting Office. Medical waste regulation - health and environmental risks need to be fully assessed. Washington, DC: U.S. General Accounting Office. Report No. GAO/RCED-90-86.

Gould, R.N.(1991) 1991 Resource recovery yearbook, directory and guide. New York, NY: Government Advisory Associates.

Greenpeace (1993) Dioxin factories: a study of the creation and discharge of dioxins and other organochlorines from the production of PVC. Amsterdam, The Netherlands: Greenpeace International.

Greer, W.L.; Johnson, M.D.; Morton, E.L; Raught, E.C.; Steuch, H.E.; Trusty, C.G. (1992) Portland cement. In: Buonicore, A.J.; Davis, W.T., eds. Air Pollution Engineering Manual, Air & Waste Management Association, Von Nostrand Reinhold Publishers, New York, NY, pp: 746-766.

Gullett, B.K., Bruce, K.R., Beach., L.O. (1990a) Formation of chlorinated organics during solid waste combustion. Waste Management and Research 8:203-214.

Gullett, B.K.; Bruce, K.R., Beach, L.O. (1990b) The effect of metal catalysts on the formation of polychlorinated dibenzo-p-dioxin and polychlorinated dibenzofuran precursors. Chemosphere 20: 1945-1952.

Gullett, B.K.; Bruce, K.R.; Beach. L.O. (1991a) The effect of sulfur compounds on the formation mechanism of PCDD and PCDF in municipal waste combustors.U.S. EPA, Research Triangle Park,NC; Acurex Corp., Research Triangle Park,NC. In: Conference papers from the second international conference on municipal waste combustion, April 15-19, 1991, Tampa, FL. Air and Waste Management Association, Pittsburgh, PA. pp.16-34.

Gullett, B.K.; Bruce, K.R.; Beach. L.O; Drago, A. (1991b) Mechanistic steps in the production of PCDD and PCDF during waste combustion. Presented at: Dioxin '91, 11th International Symposium on Chlorinated Dioxins and Related Compounds; Research Triangle Park, NC; September 1991.

Gullett, B; Bruce, K.; Beach, L. (1992) Mechanistic steps in the production of PCDD and PCDF during waste combustion. Chemosphere 25(7-10):1387-1392.

Gullett, B.K.; Lemieux, P.M. (1993) Role of combustion and sorbent parameters in prevention of polychlorinated dibenzo-p-dioxin and polychlorinated dibenzofuran formation during waste combustion. Environ. Sci. Technol. 28(1):107-118.

Hagenmaier, H.; Kraft, M.; Jager, W.; Mayer, U.; Lutzke, K.; Siegel, D. (1986) Comparison of various sampling methods for PCDDs and PCDFs in stack gas. Chemosphere 15:1187-1192.

Hagenmaier, H. (1987) Belastung der umwelt mit dioxinen. Institut für Organische Chemie der Universität Tübingen.

Hagenmaier, H.; Brunner, H. (1987) Isomer-specific analysis of pentachlorophenol and sodium pentachlorophenate for 2,3,7,8-substituted PCDD and PCDF at sub-ppb levels. Chemosphere 16:1759-1764.

Hagenmaier, H.; Dawidowsky, V.; Weber, U.B.; Hutzinger, O.; Schwind, K.H.; Thoma, H.; Essers, U.; Buhler, B.; Greiner, R. (1990) Emission of polyhalogenated dibenzodioxins and dibenzofurans from combustion-engines. Short Papers, Volume 2. Presented at: Dioxin '90, 10th International Symposium on Chlorinated Dioxins and Related Compounds; Bayreuth, Federal Republic of Germany; September 1990.

Haglund, P.; Egeback, K.-E.; Jansson, B. (1988) Analysis of PBDD/F in vehicle exhaust. Presented at: Dioxin '88, 8th International Symposium on Chlorinated Dioxins and Related Compounds; Umea, Sweden; August 1988.

Harrad, S.J.,; Malloy, T.A. (1991) Levels and sources of PCDDs, PCDFs, chlorophenols (CPs) and chlorobenzenes (CBZs) in composts from a municipal yard waste composting facility. Chemosphere 23(2):181-191.

Harrad, S.J.; Stewart, A.P.; Jones, K.C. (1992a) PCDD/CDFs in the British environment: sinks, sources and temporal trends. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Harrad, S.J.; Jones, K.C. (1992b) A source inventory and budget for chlorinated dioxins and furans in the United Kingdom environment. The Science of the Total Environment 126:89-107.

Hay, D.J.; Finkelstein, A.; Klicius, R. (1986) The national incineration testing and evaluation program: An assessment of two-stage incineration and pilot scale emission control. Paper presented to U.S. EPA Science Advisory Board, Washington, DC. Environment Canada, Ottawa, Ontario, Canada. June 1986.

Helbe, J.J. (1993) Analysis of dioxin emissions from the incineration of hazardous waste. Washington, DC: U.S. Environmental Protection Agency, Office of Solid Waste, Permits and State Programs Division.

Hiester, E.; Böhm, R.; Eynck, P.; Gerlack, A.; Mülder, W.; Ristow, H. (1993) Long-term monitoring of PCDD, PCDF and PCB in bulk deposition samples. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Hites, R.A. (1991) Atmospheric transport and deposition of polychlorinated dibenzo-p-dioxins and dibenzofurnas. Prepared for the U.S. Environmental Protection Agency, Methods Research Branch, Atmospheric Research and Exposure Assessment Laboratory, Office of Research and Development, Research Triangle Park, NC. EPA/600/3-91/002.

Horstmann, M.; McLachlan, M.; Reissinger, M. (1992) Investigation of the origin of PCDD/CDF in municipal sewage. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Horstmann, M.; McLachlan, M.; Reissinger, M. (1993a) Further investigations of the sources of PCDD/CDF in municipal sewage sludge. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Horstmann, M.; McLachlan, M.; Reissinger, M.; Morgenroth, M. (1993b) An investigation of PCDD/CDF formation during textile production and finishing. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Horstmann, M.; McLachlan, M.S. (1994) Textiles as a source of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/CDF) in human skin and sewage sludge. Environ. Sci. & Pollut. Res. 1(1):15-20.

Hryhorczuk, O.D.; Withrow, W.A.; Hesse, C.S.; Beasly, V.R. (1981) A wire reclamation incinerator as a source of environmental contamination with TCDDs and TCDFs. Archives of Environmental Health 36(5):228-234.

Huang, L.Q.; Tong, H.; Donnelly, J.R. (1992a) Characterization of dibromopolychloro-dibenzo-p-dioxins and dibromopolychlorodibenzofurans in municipal waste incinerator fly ash using gas chromatography/mass spectrometry. Anal. Chem. 64:1034-1040.

Huang, C.W.; Miyata, H.; Lu, J.; Ohta, S.; Chang, T.; Kashimoto, T. (1992b) Levels of PCBs, PCDDs and PCDFs in soil samples from incineration sites for metal reclamation in Taiwan. Chemosphere 24(11):1669-1676.

Hutzinger, O.; Fiedler, H. (1991a) Formation of dioxins and related compounds in industrial processes. In: Bretthauer, E.W.; Kraus, H.W.; di Domenico, A., eds. Dioxin perspectives. A pilot study on international information exchange on dioxins and related compounds. New York, NY: Plenum Press.

Hutzinger, O.; Fiedler, H. (1991b) Formation of dioxins and related compounds from combustion and incineration processes. In: Bretthauer, E.W.; Kraus, H.W.; di Domenico, A., eds. Dioxin perspectives. A pilot study on international information exchange on dioxins and related compounds. New York, NY: Plenum Press.

Hutzinger, O.; Fiedler, H. (1992) From source to exposure: some open questions. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Jansson, B.; Sundstrom, G; Ahling, B. (1977). Formation of polychlorinated dibenzo-p-dioxins during combustion of chlorophenol formulations. Science in the Total Environment 10: 209-217.

Johnson, W. (1992) Cement annual report 1990. Washington, DC: U.S. Bureau of Mines. April 1992.

Jones, K.H.; Walsh, J.; Alston, D. (1987) The satisfied properties of available worldwide MSW combustion/furan emissions data as they apply to the conduct of risk assessments. Chemosphere 16:2183-2186.

Jones, K. (1993) Diesel truck emissions, an unrecognized source of PCDD/CDF exposure in the United States. J. Risk Analysis 13(3):245-252.

Kiser, J.V.L; Bridges, J.P. (1993) The IWSA municipal waste combustion directory: 1993 update of U.S. plants. Washington, DC: Integrated Waste Services Association.

Kirsbaum, I.Z.; Domburg, G.E.; Sergeyeva, V.N. (1972) Extension of application range of gas chromatography to analysis of lignin pyrolysis products. Latv. PSR Zinat. Akad. Vestis, Khim. Ser., pp 700-707. Cited in: Choudhry,G.G.; Hutzinger, O. (1983). Mechanistic aspects of the thermal formation of halogenated organic compounds including polychlorinated dibenzo-p-dioxins. New York, NY: Gordon and Breach.

Knepper, W.A. (1981) Iron. In: Kirk-Othmer Encyclopedia of Chemical Technology. 3rd ed. 13:735-753. New York, NY: John Wiley and Sons.

Koester, C.J.; Hites, R.A. (1992) Wet and dry deposition of chlorinated dioxins and furans.

Environ. Sci. Technol. 26:1375-1382.

Koning, J.; Sein, A.A.; Troost, L.M.; Bremmer, H.J. (1993) Sources of dioxin emissions in the Netherlands. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Lahl, U. (1993) Sintering plants of steel industry - the most important thermical PCDD/CDF source in industrialized regions. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds: Vienna, Austria; September 1993.

Larssen, S.; Brevik, E.M.; Oehme, M. (1990) Emission factors of PCDD and PCDF for road vehicles obtained by a tunnel experiment. In: Hutzinger, O.; Fiedler, H. eds. Dioxin '90, EPRI Serminar. Bayreuth, Federal Republic of Germany: Eco-Informa Press.

Lew, G.; Jenkins, A.(1988) Evaluation retest on a hospital refuse incinerator at Sutter General Hospital, Sacramento, CA. California Air Resources Board, Sacramento, CA. April 1988.

Lew, G.; McCormack, J.; Ouchida, P. (1989) Evaluation test on a small hospital refuse incinerator at Saint Bernardines Hospital, San Bernardino, CA., California Air Resources Board, Sacramento, CA. July 1989.

Lew, G. (1993) [Letter to Mr. John Schaum concerning CARB (1987) draft report on CDD/CDFs in vehicle exhausts]. Sacramento, CA: California Air Resources Board, Monitoring and Laboratory Division. Available for inspection at: U.S. Environmental Protection Agency, Office of Health and Environmental Assessment, Washington, DC.

Lexen, K.; De Wit, C; Jansson, B.; Kjeller, L.O.; Kulp, S.E.; Ljung, K.; Söderstrom, G.; Rappe, C. (1993) Polychlorinated dibenzo-p-dioxin and dibenzofuran levels and patterns in samples from different Swedish industries analyzed within the Swedish dioxin survey. Chemosphere 27(1-3):163-170.

Liberti, A.; Brocco, D. (1982) Formation of polychlorodibenzodioxins and polychlorodibenzofurans in urban incinerator emissions. In: Hutzinger, O.; Frei, R.W.; Merian, E.; Pocchiari, F., eds., Chlorinated dioxins and related compounds. New York, NY: Pergamon Press.

Liberti, A.; Goretti, G.; Russo, M.V. (1983) PCDD and PCDF formation in the combustion of vegetable wastes. Chemosphere 12(4/5):661-663.

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.

Ligon, W.V.; Dorn, S.B.; May, R.J.; Allison, M.J. (1989) Chlorodibenzofuran and

chlorodibenzo-p-dioxin levels in Chilean mummies dated to about 2800 years before the present. Environ. Sci. Technol. 23:1286-1290.

Luijk, R.; Govers, H.A.; Neilssen, L. (1992) Formation of polybrominated dibenzofurans during extrusion of high impact polystyrene/decabromodiphenyl ether/antimony (III) oxide. Environ. Sci. Technol. 26(11):2191-2198.

Lustenhouwer, J.A.; Olie, K; Hutzinger, O. (1980) Chlorinated dibenzo-p-dioxins and related compounds in incinerator effluents. A review of measurements and mechanisms of formation. Chemosphere 9:501-522.

Lykins, B.W.; Clark, R.; Cleverly, D.H. (1987) Polychlorinated dioxin and furan discharge during carbon reactivation. Journal of Environmental Engineering 114(2): 330-316.

McCormack, J.E. (1990) ARB evaluation test conducted on a hospital waste incinerator at Los Angeles County, USC Medical Center, Los Angeles, CA. California Air Resources Board, Sacramento, CA. January 1990.

Mahle, N.H.; Whitting, L.F. (1980) The formation of chlorodibenzo-p-dioxins by air oxidation and chlorination of bituminous coal. Chemosphere 9:693-699.

Maniff, K.; Lewis, M. (1988) MISA monitoring discovers dioxins and furans in wastewater stream of petroleum refinery. Ministry of the Environment Communique. December 5, 1988.

Marklund, S. (1990) Dioxin emissions and environmental emissions. Ph.D. thesis, University of Umea, Sweden.

Marklund, S.; Kjeller, L.O.; Hasson, M.; Tysklund, M.; Rappe, C.; Ryon, C.; Collazo, H.; Dougherty, R. (1986) Determination of PCDDs and PCDFs in incineration samples and pyrolytic products. In: Rappe, C.; Choudhary, G.; Kerth, L.H., eds. Chlorinated dioxins and dibenzofurans in perspective. Lewis Publishers Inc., Michigan. pp. 79-94.

Marklund, S.; Rappe, C.; Tysklind, M.; Egeback, K.E. (1987) Identification of polychlorinated dibenzofurans and dioxins in exhausts from cars run on leaded gasoine. Chemosphere 16(1):29-36.

Marklund, S.; Andersson, R.; Tysklind, M.; Rappe, C.; Egeback, K.E.; Bjorkman, E.; Grigoriadis, V. (1990) Emissions of PCDDs and PCDFs in gasoline and diesel fueled cars. Chemosphere 20(5):553-561

McKee, ZP.; Burt, A.; McCurvin, D. (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.

Mattila, H.; Virtanen, T.; Vartiainen, T.; Ruuskanen, J. (1992) Emissions from combustion of waste plastic material in fixed bed boiler. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Miller, A. (1993) Dioxin emissions from EDC/VCM plants. Environ. Sci. Technol. 27:1014-1015.

Muller, M.D.; Buser, H.R. (1986) Halogenated aromatic compounds in automotive emissions Environ. Sci. Technol. 20:1151.

Muto, H.; Saito, K.; Shinada, M.; Takizawa, Y. (1991) Concentrations of polychlorinated dibenzo-p-dioxins and dibenzofurans from chemical manufacturers and waste disposal facilities. Environmental Research 54:170-182.

NATO (1988) Pilot study on international information exchange on dioxins and related compounds: Emissions of dioxins and related compounds from combustion and incineration sources. North Atlantic Treaty Organization, Committee on the Challenges of Modern Society. Report #172. August 1988.

NCASI. National Council of the Paper Industry for Air and Stream Improvement. (1993). Summary of data reflective of pulp and paper industry progress in reducing the TCDD/TCDF content of effluents, pulps and wastewater treatment sludges. New York, NY. NCASI. June 1993.

Näf, C.; Broman, D.; Pettersen, H.; Rolff, C.; Zebühr, Y. (1992) Flux estimates and pattern recognition of particulate polycyclic aromatic hydrocarbons, polychlorinated dibenzo-p-dioxins, and dibenzofurans in the waters outside various emission sources on the Swedish Baltic coast. Environ. Sci. Technol. 26(7):1444-1457.

Nestrick, T.J., Lamparski, L.L. (1983) Assessment of chlorinated dibenzo-p-dioxin formation and potential emission to the environment from wood combustion. Chemosphere 12(4/5):617-626.

Nestrick, T.J., Lamparski, L.L., Crummett, W.B. (1987) Thermolytic surface-reaction of benzene and iron (III) chloride to form chlorinated dibenzo-p-dioxins and dibenzofurnas. Chemosphere 16(4):777-790.

Oberg, T.; Warman, K.; Berstrom, J. (1989). Production of chlorinated aromatics in the post-combustion zone and boiler. Chemosphere 19(1-6):317-322.

Oberg L.G.; Andersson, R., Rappe, C. (1992) De novo formation of hepta- and octa-chlorodibenzo-p-dioxins from pentachlorophenol in municipal sewage sludge. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Oehme, M.; Mano, S.; Bjerke, B. (1989) Formation of polychlorinated dibenzofurans and dibenzo-p-dioxins by production processes for magnesium and refined nickel. Chemosphere 18(7-8):1379-1389.

Oehme, M.; Larssen,S.; Brevik, E.M. (1991) Emission factors of PCDD/CDF for road vehicles obtained by a tunnel experiment. Chemosphere 23:1699-1708.

Olie, K.; Vermeulen, P.L.; Hutzinger, O. (1977) Chlorodibenzo-p-dioxins and chlorodibenzofurans are trace components of fly ash and flue gas of some municipal incinerators in the Netherlands. Chemosphere 8:455-459.

Olie, K.; Berg, M.v.d.; Hutzinger, O. (1983) Formation and fate of PCDD and PCDF from combustion processes. Chemosphere 12(4/5):627-636.

Ozvacic, V. (1985). A review of stack sampling methodology for PCDDs/PCDFs. Chemosphere 15(9-12):1173-1178.

Paciorek, K.L.; Kratzer, R.H.; Kaufman, J.; Nakahara, J.; Hartstein, A.M. (1974) Oxidative thermal decomposition of poly (vinyl chloride) compositions. J. Appl Polym. Sci. 18:3723-3729.

Peters, J.A. (1983) Evaluation of hazardous waste incineration in cement kilns at San Juan Cement Company. Prepared for Incineration Research Branch, Industrial Research Laboratory, U.S. Environmental Protection Agency, Cincinnati, OH.

Peters, J.S., Hughes, T.W., and Mournighan, E.E. (1984). Evaluation of hazardous waste incineration in a cement kiln at San Juan Cement. In: Incineration and Treatment of Hazardous Waste - Proceedings of the 9th annual Research Symposium (EPA-600/9-84-015). July 1984; pp: 210-24.

Pitea, D.; Lasagni, M.; Bonati, L.; Moro, G.; Todeschini, R.; Chiesa, G. (1989a) The combustion of municipal solid wastes and PCDD and PCDF emissions. Part 1. PCDD and PCDF in MSW. Chemosphere 18(7-8):1457-1464.

Pitea, D.; Lasagni, M.; Bonati, L.; Moro, G.; Todeschini, R.; Chiesa, G. (1989b) The combustion of municipal solid wastes and PCDD and PCDF emissions. Part 2. PCDD and PCDF in stack gases. Chemosphere 18(7-8):1465-1474.

Radian Corporation (1991a) Medical waste incineration emission test report, Lenoir memeorial hospital, Kinsston, NC. EMB Report No. 90-MWI-3, Volume I. Prepared for the U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC. November 1991.

Radian Corporation (1991b) Medical waste incineration emission test report, AMI Central Carolina Hospital, Sanford, NC, EMB Report # 90-MWI-5, Volume I. Prepared for the U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC. December 1991.

Radian Corporation (1991c) Medical waste incineration emission test report,Cape Fear Memorial Hospital, Wilmington, NC, EMB Report # 90-MWI-4, Volume I. Prepared for the U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards, Research Triangle Park, NC. December 1991.

Radian Corporation (1994) Memorandum from R. Harrison, C. Blackley to W. Stevenson, U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards, Research Triangle Park, N.C. Radian Corp., Research Triangle Park, N.C. February 16, 1994.

Rappe, C. (1987) Global distribution of polychlorinated dioxins and dibenzofurans. In: Exner, J.H., ed. Solving hazardous waste problems learning from dioxins. ACS Symposium Series 338. Division of Environmental Chemistry. 191st meeting the American Chemical Society New York, New York, April 13-18, 1986.

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

Rappe, C. (1992b) Sources of exposure, environmental levels and exposure assessment of PCDDs and PCDFs. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

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

Rappe, C.; Andersson, R. (1992) Analyses of PCDDs and PCDFs in wastewater from dish washers and washing machines. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Rappe, C.; Andersson, R.; Berggvist, P.A.; Brohede, C.; Hansson, M.; Kjeller, L.O.; Lindstrom, G.; Marklund, S.; Nygren, M.; Swanson, S.E.; Tysklind, M.; Wiberg, K. (1987) Overview of the environmental fate of chlorinated dioxins and dibenzofurans. Sources, levels and isomeric pattern in various matrices. Chemosphere 16(8-9):1603-1618.

Rappe, C.; Kjeller, L.O.; Bruckmann, P.; Hackhle, K.H. (1988) Identification and quantification of PCDD/CDFs in urban air. Chemosphere 17(1):3-20.

Rappe, C.; Swanson, S.; Gals, B.; Kringstad, K.P.; de Sousa, P.; Zensaku, A. (1989) Formation of PCDDs and PCDFs by the chlorination of water. Chemosphere 19(12):1875-1880.

Rappe, C.; Glas, B.; Kjeller, L.O.; Kulp, S.E. (1990) Levels of PCDDs and PCDFs in products and effluent from the Swedish pulp and paper industry and chloralkali process. Chemosphere 20(10-12):1701-1706.

Ree, K.C.; Evers, E.H.G.; Van Der Berg, M. (1988) Mechanism of formation of polychlorinated dibenzo(p)dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) from potential industrial sources. Toxicological and Environmental Chemistry 17:171-195.

Rieger, R.; Ballschmiter, K. (1992) Search for sources of CLxDD/CLxDF in sewage sludge of mixed industrial/domestic origin. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Riss, A.; Aichinger, H. (1993) Reduction of dioxin emissions and regulatory measures in Austria. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Robbins, J.A.; Edgington, D.N. (1973) Geochim Cosmochim Acta 39, p.285. As cited in: Czuczwa, J.M; Hites, R.A. (1985). Historical record of polychlorinated dioxins and furans in Lake Huron sediments. In: Keith, L.H.; Rappe. C.; Choudhary, G, eds. Chlorinated dioxins and dibenzofurans in the total environment II. Boston, MA: Butterworth Publishers.

RTI. Research Triangle Institute (1993a) Dioxin/furan summary. Information gathering for NESHAP for the portland cement manufacturing industry. Memorandum to Joseph P. Wood (U.S. EPA, Office of Air Quality Planning and Standards, RTP, NC) from Jeffrey W. Portzer (RTI, RTP, NC) dated October 1, 1993. EPA Contract No. 68-D1-0118.

RTI. Research Triangle Institute (1993b) Submission of third revision of model plant definition information gathering for NESHAP for the portland cement manufacturing industry. Memorandum to Mary K. Johnson (U.S. EPA, Office of Air Quality Planning and Standards, RTP, NC) from Jeffrey W. Portzer (RTI, RTP, NC) dated July 23, 1993. EPA Contract No. 68-D1-0118.

Rubin, A.; White, C. (1992) Computer printout from National Sewage Sludge Survey Data Base. USEPA, Office of Science and Technology, Health and Ecological Criteria Division, Sludge Risk Assessment Branch. December 21, 1992.

Schatowitz, B.; Brandt, G.; Gafner, F.; Schlumpf, E.; Biihler, R.; Hasler, P.; Nussbaumer, T. (1993) Dioxin emissions from wood combustion. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Schaub, W.M.; Tsang, W. (1983) Physical and chemical properties of dioxins in relation to their disposal. In: Tucker, R.E.; Young, A.L., Gray, A.P., eds. Human and Environmental Risks of Chlorinated Dioxins and Related Compounds. New York, NY: Plenum Press.

Schecter, A. (1991) Dioxins and related chemicals in human and in the environment. In: Gallo, M.; Scheuplein, R.; Van der Heijden, K. eds., Banbury Report #35: Biological basis for risk assessment of dioxins and related compounds. Plainview, NY: Cold Spring Harbor Laboratory Press.

Schmitt, R. (1993) Comments of pesticide data call-in [memorandum to John Schaum, U.S. EPA, Office of Research and Development]. Washington, D.C.: U.S. Environmental Protection Agency, Office of Pesticide Programs. April 27, 1993.

Smith, R.M.; O'Keefe P.; Briggs, R.; Hilker, D.; Connor, S. (1992) Measurement of PCDFs and PCDDs in air samples and lake sediments at several locations in upstate New York. Chemosphere 25(1-2):95-98.

Smith, R.M.; O'Keefe, P.W.; Hilker, D.R.; Bask, B.; Connor, S.; Donnelly, R.; Storm, R.; Liddle, M. (1993) The historical record of PCDDs, PCDFs, PAHs, PCBs, and lead in Green Lake, New York - 1860 to 1990. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Someshwar, A.V.; Pinkerton, J.E. (1992) Wood processing industry. In: Buonicore, A.J., Davis, W.T., eds., Air pollution engineering manual. Air and Waste Management Association. New York, NY: Van Nostrand Reinhold. pp. 835-847.

Southerland, J.H.; Kuykendal, W.B.; Lamasan, W.H.; Miles, A.; Oberacker, D.A. (1987) Assessment of combustion sources as emitters of chlorinated dioxin compounds: A report on the results of tier 4 of the National Dioxin Strategy. Chemosphere 16(8/9):2161-2168.

Stehl, R.H., Papenfuss, R.R., Bredeweg, R.A., Roberts, R.W. (1973) The stability of pentachlorophenol and chlorinated dioxins to sunlight, heat, and combustion. In: E.H. Blair ed., Chlorodioxins - origin and fate. Washington, DC: American Chemical Society.

Stehl, R.H.; Lamparski, L.L. (1977) Combustion of several 2,4,5-trichlorophenoxy compounds: formation of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Science 197:1008-1009.

Stieglitz, L.; Zwick, G.; Beck, J.; Bautz, H.; Roth, W. (1989a) Carbonaceous particles in fly ash--a source for the de novo synthesis of organochlorocompounds. Chemosphere 19(1-6):283-290.

Stieglitz, L.; Zwick, G.; Beck, J.; Roth, W.; Vogg, H. (1989b) On the de-novo synthesis of PCCCD/PCDF on fly ash of municipal waste incinerators. Chemosphere 18:1219-1226.

Stieglitz, L; Vogg, H.; Zwick, G.; Beck, J.; Bautz, H. (1991) On formation conditions of organohalogen compounds from particulate carbon of fly ash. Chemosphere 23(8-10):1255-1264.

SRI International (1992) 1992 Directory of chemical producers: United States of America. SRI International, Menlo Park, California. pp. 838.

Suter-Hofmann, M.; Schlatter, C.H. (1986) Toxicity of particulate emissions from a municipal incinerator: critique of the concept of TCDD-equivalents. Chemosphere 15:1733-1743.

Svensson, B.G.; Barregard, L.; Sallsten, G.; Nilsson, A.; Hansson, M.; Rappe,C. (1992) Exposure to polychlorinated dioxins (PCDD) and dibenzofurans (PCDF) from graphite electrodes in a chloralkali plant. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Swerev, M.; Ballschmiter, K. (1989) Pattern analysis of PCDDs and PCDFs in environmental samples as an approach to an occurrence/source correlation. Chemosphere 18(1-6):609-616.

Tashiro, C.; Clement, R.E.; Stocks, B.J.; Radke, L.; Cofer, W.R.; Ward, P. 1990. Preliminary report: dioxins and furans in prescribed burns. Chemosphere 20:1533-1536.

Thoma, H.; Hutzinger, O. (1989) Content and formation of toxic products in flame retardants. Presented at: Workshop on Brominated Aromatic Flame Retardants. Slokloster, Sweden; October 24-26, 1989.

Thompson, T.S.; Clement, R.E.; Thornton, N.; Luty, J. (1990) Formation and emission of PCDDs/PCDFs in the petroleum refining industry. Chemosphere 20(10-12):1525-1532.

Tong, H.G.; Monson, S.J.; Gross, M.L.; Bopp, R.F., Simpson, H.J.; Deck, B.L.; Moser, F.C. (1990) Analysis of dated sediment samples from the Newark Bay area for selected PCDD/CDFs. Chemosphere 20(10-12):1497-1502.

Tosine, H.; Clement, R.; Osvacic, V.; Osborne, J.; Wong, C. (1983). Levels of chlorinated organics in a municipal incinerator. Presented before the Division of Environmental Chemistry, American Chemical Society 186th National Meeting, September, 1, 1983.

Travis, C.C.; Hattemer-Frey, H.A. (1991) Human exposure to dioxin. Science of the Total Environment 104:97-127.

Tysklind, M.; Söderström; Rappe, C.; Hägerstedt, L-E.; Burström, E. (1989) PCDD and PCDF emissions from scrap metal melting processes at a steel mill. Chemosphere 19(1-6):705-710.

U.S. Department of Commerce (1990a) 1987 Census of manufactures. Smelting and refining of nonferrous metals and alloys. Washington, DC: Bureau of the Census. Report No. 33C-18.

U.S. Department of Commerce (1990b) 1987 Census of transportation. Truck inventory and use survey. Washington, DC: Bureau of the Census. Report No. TC87-T-52.

U.S. Department of Commerce (1990c) 1987 Census of manufactures. Industrial inorganic chemicals. Washington, DC: Bureau of the Census. Report No. MC87-1-28A.

U.S. Department of Commerce (1992) Statistical abstract of the United States, 1992. 112th ed.

U.S. Environmental Protection Agency; Environment Canada. (1991) The environmental characterization of RDF combustion technology. Mid-Connecticut facility, Hartford,CT. Test program and results, volume II., June 1991.

U.S. Environmental Protection Agency (1982) Development document for effluent limitations guidelines and standards for the inorganic chemicals manufacturing point source category. Washington, DC: Office of Water Regulations and Standards. EPA 440/1-82/007.

U.S. Environmental Protection Agency. (1984) Thermal degradation products from dielectric fluids. Washington, DC: Office of Toxic Substances. EPA-560/5-84-009.

U.S. Environmental Protection Agency (1987) National dioxin study Tier 4 - combustion sources. Engineering analysis report. Research Triangle Park, NC: Office of Air Quality Planning and Standards. EPA-450/4-84-01h.

U.S. Environmental Protection Agency (1988) Pesticide assessment guidelines, subdivision D-product chemistry. Series 61-4, addendum 1 on data reporting. Washington, DC: Office of Pesticide Programs. PB88-191705.

U.S. Environmental Protection Agency (1990a) Summary report on USEPA/Industry Cooperative Dioxin Study. "The 104 Mill Study." Washington, DC: Office of Water Regulations and Standards. July 1990.

U.S. Environmental Protection Agency (1990b) Background document to the integrated risk assessment for dioxins and furans from chlorine bleaching in pulp and paper mills. Washington, DC: Office of Toxic Substances. EPA 560/5-90-014.

U.S. Environmental Protection Agency (1990c) Characterization of municipal waste combustion ash, ash extracts, and leachates. Washington, DC: Office of Solid Waste and Emergency Response. EPA 530-SW-90-029A.

U.S. Environmental Protection Agency (1990d) Chlorinated dioxins and furans in the general US population: NHATS FY 1987 results. Washington, DC: Office of Toxic Substances. EPA-560/5-91/003.

U.S. Environmental Protection Agency (1991a) Regulatory determination-landfills and surface impoundments receiving pulp and paper mill sludge. Washington, DC: Office of Solid Waste and Emergency Response. EPA/530-SW-91-077.

U.S. Environmental Protection Agency (1991b) Methodology for assessing environmental releases of and exposure to municipal solid waste combustor residuals. Washington, DC: Office of Research and Development. EPA/600/8-91/031.

U.S. Environmental Protection Agency (1991c) Compilation of air pollutant emission factors. Report No. AP-42. Residential fireplaces and wood stoves. Chapter 1.10. September 1991.

U.S. Environmental Protection Agency (1991d) Medical waste incinerators background information for proposed standards and guidelines: Industry profile report for new and existing facilities. Public Docket No. A-91-61 II-C, Office of Air and Radiation, Washington, DC, September 23, 1991 Draft.

U.S. Environmental Protection Agency (1991e) Medical waste incinerators-background information for proposed standards and guidelines: Environmental impacts report for new and existing facilities. Public Docket No. A-91-61 II-C-1, Office of Air and Radiation, Washington, DC, September 30, 1991 Draft.

U.S. Environmental Protection Agency (1991f) Medical waste incinerators-background information for proposed standards and guidelines: Process description/baseline emissions report for new and existing facilities. Public Docket No. A-91-61II-C-4, Office of Air and Radiation, Washington, DC, September 30, 1991 Draft.

U.S. Environmental Protection Agency (1992a) Summary of markets for scrap tires. Washington, DC: Office of Solid Waste. EPA/530-SW-90-074B.

U.S. Environmental Protection Agency (1992b) Industry agrees to switch to low-dioxin chloranil from contaminated chloranil. Chemicals in Progress Bulletin 12(2):23.

U.S. Environmental Protection Agency (1992c) Characterization of municipal solid waste in the United States: 1992 update, executive summary. Washington, DC: Office of Solid Waste. EPA/530-S-92-019.

U.S. Environmental Protection Agency (1992d) Effluent guidelines and MACT standards for the pulp, paper, and paperboard industry. Presentation for the Effluent Guidelines Task Force, February 10, 1993. Washington, DC: Office of Science and Technology.

U.S. Environmental Protection Agency (1992e) Preliminary risk assessment of inhalation exposures to stack emissions from the WTI incinerator. Prepared by A.T. Kearney, Inc. and Environ. Corp for the U.S. EPA Region 5. July 1992.

U.S. Environmental Protection Agency (1992f) Hospital waste incinerators: Background information document for proposed existing source guidelines. July 1992. Draft. Research Triangle Park, NC: Office of Air Quality Planning and Standards.

U.S. Environmental Protection Agency (1992g) 1990 National census of pulp, paper, and paperboard manufacturing facilities. Response to 308 questionnaire. Part A: Technical Information. Washington, DC: Office of Water.

U.S. Environmental Protection Agency (1992h) Economic impact and preliminary regulatory impact analysis for proposed MACT-based emission standards and guidelines for municipal waste combustors. Research Triangle Park, NC: Office of Air Quality Planning and Standards. EPA-450/3-91-029.

U.S. Environmental Protection Agency (1992i) Emission test report, HAP emission testing on selected sources at a secondary lead smelter, Tejas Resources, Inc., Terrell, Texas. Research Triangle Park, NC: Office of Air Quality Planning and Standards, Emission Measurement Branch. EPA Contract No. 68-D1-0104, Work Assignments No. 23 and 1-30.

U.S. Environmental Protection Agency (1993a) Locating and estimating air emissions from sources of dioxins and furans. Draft Final Report. Prepared by Radian Corp. for the Office of Air Quality Planning and Standards, Research Triangle Park, NC. September 30, 1993.

U.S. Environmental Protection Agency (1993b) States support efforts on forming voluntary agreement on pulp and paper mill sludge. Chemicals in Progress Bulletin 14(3):20.

U.S. Environmental Protection Agency (1993c) EPA publishes proposed rule for chloranil. Chemicals in Progress Bulletin 14(2):26.

U.S. Environmental Protection Agency (1993d) Development document for proposed effluent limitations guidelines and standards for the pulp, paper and paperboard point source category. Washington, DC: Office of Water. EPA-821-R-93-019.

U.S. Environmental Protection Agency (1993e) Economic analysis of impacts of integrated air/water regulations for the pulp and paper industry on disposal of wastewater sludge. Washington, DC: Office of Pollution Prevention and Toxics, Regulatory Impacts Branch.

U.S. Environmental Protection Agency (1993f) Report to Congress on cement kiln dust. Washington, DC: Office of Solid Waste. December 1993.

U.S. Environmental Protection Agency (1993g) Summary of results, draft data tables, HAP emission testing on selected sources at a secondary lead smelter, East Penn Manufacturing Company, Lyon Station, Pennsylvania. Research Triangle Park, NC: Office of Air Quality Planning and Standards, Emission Measurement Branch. EPA Contract Nos. 68-D1-0104 and 68-D2-0029, Work Assignments No. 30 and 1-12.

U.S. Environmental Protection Agency (1993h) Emission test report, HAP emission testing on selected sources at a secondary lead smelter, Schuylkill Metals Corporation, Forest City, Missouri. Research Triangle Park, NC: Office of Air Quality Planning and Standards, Emission Measurement Branch. EPA Contract No. 68-D1-0104. Work Assignments No. 23 and 1-30.

U.S. Environmental Protection Agency (1994a) Secondary lead smelting industry: background information document for proposed emission standards. Draft report. Research Triangle Park, NC: Office of Air Quality Planning and Standards.

U.S. International Trade Commission (1991) Synthetic organic chemicals - United States production, 1990. Washington, DC: U.S. International Trade Commission. USITC Public. No. 2470.

U.S. International Trade Commission (1993) Synthetic organic chemicals - United States production and sales, 1991. Washington, DC: U.S. International Trade Commission. USITC Public. No. 2607.

Van Jaarsveld, JA; Schutter, M.A.A. (1992) Modeling the long range transport and deposition of dioxins; first results for the North Sea and surrounding countries. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Van Wijnen, J.H.; Liem, A.K.D.; Olie, K.; van Zorge, J.A. (1992) Soil contamination with PCDDs and PCDFs of small (illegal) scrap wire and scrap car incineration sites. Chemosphere 24(2):127-134.

Versar Inc. (1985) List of chemicals contaminated or precursors to contamination with incidentally generated polychlorinated and polybrominated dibenzodioxins and dibenzofurans. EPA Contract No. 68-02-3968, Task No. 48.

Vikelsoe, J.; Madsen, H.; Hansen, K. (1993) Emission of dioxins from Danish wood-stoves. Presented at: Dioxin '93, 13th International Symposium on Chlorinated Dioxins and Related Compounds; Vienna, Austria; September 1993.

Vogg, H.; Metzger, M.; Steiglitz, L. (1987) Recent findings on the formation and decomposition of PCDD/PCDF in municipal solid waste incineration. Waste Management and Research 5(3):285-294.

Vogg, H; Nunsinger, H., Merz, A; Stieglitz, L. (1992). Influencing the production of dioxin/furan in solid waste incineration plants by measures affecting the combustion as well as the flue gas cleaning systems. Chemosphere 25(1-2):149-152.

Ward, D.E.; McMahon, C.K.; Johansen, R.W. (1976) An update on particulate emissions from forest fires. Presented at: 69th Annual Meeting of the Air Pollution Control Association. Portland, OR. June 27-July 1, 1976.

Ward, D.E.; Peterson, J.; Hao, W.M. (1993) An inventory of particulate matter and air toxic emissions from prescribed fires in the USA for 1989. Presented at: 86th Annual Meeting & Exhibition, Air & Waste Management Association; Denver, Colorado; June 1993.

Watanabe, I; Tatsukawa, R. (1987) Formation of brominated dibenzofurans from the photolysis of flame retardant decabromobiphenyl ether in hexane solution by UV and sun light. Bull. Environ. Contam. Toxicol. 39:953-959.

Wenning, R.J.; Harris, M.A.; Paustenbach, D.J.; Bedbury, H. (1992) Potential sources of 1,2,8,9-tetrachlorodibenzo-p-dioxin in the aquatic environment. Ecotoxicology and Environmental Safety 23:133-146.

Wevers,M., R. De Fre, T. Rymen. (1992) Dioxins and dibenzofurans in tunnel air. Volume 9 (Sources of Exposure) of Extended Abstracts. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Wilken, M.; Cornelsen, B.; Zeschmar-Lahl, B.; Jager, J. (1992) Distributrion of PCDD/PCDF and other organochlorine compounds in different municipal solid waste fractions. Chemosphere 25:1517-1523.

Williams, D.T.; LeBel, G.L., Benoit, F.M. (1992) Polychlorodibenzodioxins and polychlorodibenzofurans in dioxazine dyes and pigments. Chemosphere 24(2):169-180.

Yamamoto, T.; Fukushima, M. (1992) Modeling study on contribution of combustion source complex to PCDD/CDF levels in urban air. Presented at: Dioxin '92, 12th International Symposium on Chlorinated Dioxins and Related Compounds; Tampere, Finland; August 1992.

Yasuhara, A.; Ito, H.; Morita, M. (1987) Isomer-specific determination of polychlorinated dibenzo-p-dioxins and dibenzofurans in incinerator-related environmental samples. Environ. Sci. Technol. 21:971-979.