EPA/600/P-03/002F

 November 2006 

An Inventory of Sources and Environmental Releases of
 
 
 

Dioxin-Like Compounds in the United States for
 
 
 

the Years 1987, 1995, and 2000
 
 
 

National Center for Environmental Assessment
 
 
 

Office of Research and Development
 
 
 

U.S. Environmental Protection Agency
 
 
 

Washington, DC 20460
 
 
 
DISCLAIMER
 

This document has been reviewed in accordance with U.S. Environmental Protection 

Agency policy and approved for publication. Mention of trade names or commercial products 

does not constitute endorsement or recommendation for use. 

ABSTRACT 

The purpose of this document is to present a comprehensive inventory and overview of 

sources and environmental releases of dioxin-like compounds in the United States. The major 

identified sources of environmental releases of dioxin-like compounds are grouped into six broad 

categories: combustion sources, metals smelting, refining and process sources, chemical 

manufacturing sources, natural sources, and environmental reservoirs. Estimates of annual 

releases to land, air, and water are presented for each source category and summarized for 

reference years 1987, 1995, and 2000. The quantitative results are expressed in terms of the 

toxicity equivalence (TEQ) of the mixture of polychlorinated dibenzo-p-dioxin (CDD) and 

polychlorinated dibenzofuran (CDF) compounds present in environmental releases using a 

procedure sanctioned by the World Health Organization (WHO) in 1998. This TEQ procedure 

translates the complex mixture of CDDs and CDFs characteristic of environmental releases into 

an equivalent toxicity concentration of 2,3,7,8-tetrachorodibenzo-p-dioxin (2,3,7,8-TCDD), the 

most toxic member of this class of compounds. Using this WHO procedure, the annual releases 

of TEQDF-WHO98 to the U.S. environment over the three reference years are 13,965 g in 1987, 

3,444 g in 1995, and 1,422 g in 2000. This analysis indicates that between reference years 1987 

and 2000, there was approximately a 90% reduction in the releases of dioxin-like compounds to 

the circulating environment of the United States from all known sources combined. In 1987 and 

1995, the leading source of dioxin emissions to the U.S. environment was municipal waste 

combustion; however, because of reductions in dioxin emissions from municipal waste 

combustors, it dropped to the fourth ranked source in 2000. Burning of domestic refuse in 

backyard burn barrels remained fairly constant over the years, but in 2000, it emerged as the 

largest source of dioxin emissions to the U.S. environment. 

Preferred Citation: 

U.S. EPA (Environmental Protection Agency). (2006) An inventory of sources and environmental releases of dioxinlike compounds in the United States for the years 1987, 1995, and 2000. National Center for Environmental 

Assessment, Washington, DC; EPA/600/P-03/002F. Available from: National Technical Information Service, 

Springfield, VA, and online at http://epa.gov/ncea. 

ii 
CONTENTS
 
 
 

LIST OF TABLES ....................................................... xii 
 
 
 

LIST OF FIGURES ..................................................... xxv 
 
 
 

LIST OF ABBREVIATIONS AND ACRONYMS ................................ xxx 
 
 
 

FOREWORD ......................................................... xxxiv
 
 
 

PREFACE ........................................................... xxxvi
 
 
 

AUTHORS, CONTRIBUTORS, AND REVIEWERS ........................... xxxvii
 
 
 

ACKNOWLEDGMENT ...................................................... xli 
 
 
 

EXECUTIVE SUMMARY .................................................... xlii 
 
 
 

1. 	 BACKGROUND, APPROACH, AND CONCLUSIONS ......................... 1-1 
 
 
 

1.1. 	 BACKGROUND ................................................... 1-1 
 
 
 

1.1.1. Reference Years .............................................. 1-3 
 
 

1.1.2. Regulatory Summary .......................................... 1-3 
 
 
 

1.1.3. Definition of Dioxin-Like Compound ............................. 1-4 
 
 
 

1.1.4. Toxicity Equivalence Factors .................................... 1-8 
 
 
 

1.1.5. Information Sources .......................................... 1-11 
 
 
 

1.2. 	 APPROACH ..................................................... 1-12 
 
 

1.2.1. Source Classes .............................................. 1-12 
 
 
 

1.2.2. Quantitative Method for Inventory of Sources ...................... 1-23 
 
 
 

1.2.3. Confidence Ratings ........................................... 1-27 
 
 
 

1.2.3.1. Rating Scheme ....................................... 1-31 
 
 
 

1.3. 	 CONCLUSIONS .................................................. 1-37 
 
 
 

1.3.1. Total Environmental Releases .................................. 1-37 
 
 
 

1.3.2. Time Trends ................................................ 1-42 
 
 
 

1.3.3. Sources Not Included in the Inventory ............................ 1-42 
 
 
 

1.3.4. Formation Theory ............................................ 1-43 
 
 
 

1.3.5. Congener Profiles of CDD/CDF Sources .......................... 1-44 
 
 
 

2. 	 MECHANISMS OF FORMATION OF DIOXIN-LIKE COMPOUNDS DURING
 
 

COMBUSTION OF ORGANIC MATERIALS ................................. 2-1 
 
 

2.1. 	 MECHANISM 1 (PASS THROUGH): CDD/CDF CONTAMINATION IN
 
 
 

FUEL AS A SOURCE OF COMBUSTION STACK EMISSIONS ............ 2-3 
 
 
 

2.2. 	 MECHANISM 2 (PRECURSOR): FORMATION OF CDDs/CDFs FROM
 
 
 

PRECURSOR COMPOUNDS ........................................ 2-4 
 
 
 

2.3. 	 MECHANISM 3 (DE NOVO SYNTHESIS): SYNTHESIS OF CDDs/CDFs
 
 
 

DURING COMBUSTION OF ORGANIC MATERIALS .................. 2-12 
 
 
 

2.4. 	 THE ROLE OF CHLORINE IN THE FORMATION OF CDDs/CDFs IN
 
 
 

COMBUSTION SYSTEMS ......................................... 2-20 
 
 
 

2.4.1. Review of Laboratory-Scale Studies ............................. 2-21 
 
 
 

2.4.2. Review of Full-Scale Combustion Systems ........................ 2-26 
 
 
 

2.5. 	 POTENTIAL PREVENTION OF CDD/CDF FORMATION IN
 
 
 

COMBUSTION SYSTEMS ......................................... 2-28 
 
 
 

2.6. 	 THEORY ON THE EMISSION OF PCBs .............................. 2-29 
 
 
 

iii 
CONTENTS (continued) 

2.7. SUMMARY AND CONCLUSIONS .................................. 2-31 
 
 
 

2.7.1. Mechanisms of Formation of Dioxin-Like Compounds ............... 2-31 
 
 
 

2.7.2. Role of Chlorine ............................................. 2-33 
 
 
 

2.7.3. General Conclusion ........................................... 2-34 
 
 
 

3. COMBUSTION SOURCES OF CDDs/CDFs: WASTE INCINERATION ........... 3-1 
 
 
 

3.1. MUNICIPAL WASTE COMBUSTION .................................. 3-1 
 
 

3.1.1. Description of Municipal Waste Combustion Technologies ............ 3-2 
 
 
 

3.1.1.1. Furnace Types ........................................ 3-2 
 
 
 

3.1.1.2. Air Pollution Control Devices ............................ 3-9 
 
 
 

3.1.1.3. Classification Scheme ................................. 3-10 
 
 
 

3.1.2. Characterization of MWCs in Reference Years 2000, 1995, and 1987 . . . 3-11
 
 
 

3.1.3. Estimation of CDD/CDF Emissions from MWCs ................... 3-19 
 
 
 

3.1.3.1. Estimating CDD/CDF Emissions from MWCs in Reference

 Year 2000 ........................................... 3-19 
 
 
 

3.1.3.2. Estimating CDD/CDF Emissions from MWCs in Reference

 Years 1995 and 1987 .................................. 3-21 
 
 
 

3.1.4. Summary of CDD/CDF (TEQ) Emissions from MWCs for 2000, 1995,
 
 
 

and 1987 ................................................... 3-23 
 
 
 

3.1.5. Congener Profiles of Municipal Waste Combustion Facilities ......... 3-31 
 
 
 

3.1.6. Estimated CDDs/CDFs in MWC Ash ............................ 3-31 
 
 
 

3.1.7. Recent EPA Regulatory Activities ............................... 3-40 
 
 
 

3.2. HAZARDOUS WASTE INCINERATION .............................. 3-41 
 
 
 

3.2.1. Furnace Designs for HWIs ..................................... 3-42 
 
 
 

3.2.2. APCDs for HWIs ............................................ 3-44 
 
 
 

3.2.3. Estimation of CDD/CDF Emission Factors for HWIs ................ 3-45 
 
 
 

3.2.4. Emission Estimates for HWIs ................................... 3-47 
 
 
 

3.2.5. Recent EPA Regulatory Activities ............................... 3-51 
 
 
 

3.2.6. Industrial Boilers and Furnaces Burning Hazardous Waste ............ 3-51 
 
 
 

3.2.7. Halogen Acid Furnaces Burning Hazardous Waste .................. 3-53 
 
 
 

3.2.8. Solid Waste from Hazardous Waste Combustion .................... 3-55 
 
 
 

3.3. MEDICAL WASTE INCINERATION ................................. 3-55 
 
 
 

3.3.1. Design Types of MWIs Operating in the United States ............... 3-55 
 
 
 

3.3.2. Characterization of MWIs for Reference Years 1987, 1995, and 2000 . . . 3-57
 
 
 

3.3.3. Estimation of CDD/CDF Emissions from MWIs .................... 3-59 
 
 
 

3.3.4. Summary of CDD/CDF Emissions from MWIs ..................... 3-64 
 
 
 

3.3.5. Recent EPA Regulatory Activities ............................... 3-65 
 
 
 

3.4. CREMATORIA ................................................... 3-65 
 
 
 

3.4.1. Human Crematoria ........................................... 3-65 
 
 
 

3.4.1.1. Emissions Data ....................................... 3-65 
 
 
 

3.4.1.2. Activity Level Information.............................. 3-73 
 
 
 

3.4.1.3. Emission Estimates ................................... 3-76 
 
 
 

3.4.2. Animal Crematoria ........................................... 3-76 
 
 
 

iv
 
CONTENTS (continued) 

3.4.2.1. Emissions Data ....................................... 3-76 
 
 
 

3.4.2.2. Activity Level Information .............................. 3-76 
 
 
 

3.4.2.3. Emission Estimates ................................... 3-76 
 
 
 

3.5. SEWAGE SLUDGE INCINERATION ................................. 3-77 
 
 
 

3.5.1. Emissions Estimates from Sewage Sludge Incinerators ............... 3-79 
 
 
 

3.5.2. Solid Waste from Sewage Sludge Incinerators ...................... 3-83 
 
 
 

3.6. 	 TIRE COMBUSTION .............................................. 3-83 
 
 
 

3.7. 	 COMBUSTION OF WASTEWATER SLUDGE AT BLEACHED
 
 
 

CHEMICAL PULP MILLS .......................................... 3-87 
 
 
 

3.8. 	 BIOGAS COMBUSTION ........................................... 3-87 
 
 
 

4. 	 COMBUSTION SOURCES OF CDDs/CDFs: POWER/ENERGY GENERATION . . . 4-1
 
 

4.1. MOTOR VEHICLE FUEL COMBUSTION .............................. 4-1 
 
 
 

4.1.1. Tailpipe Emission Studies ...................................... 4-1 
 
 
 

4.1.2. Tunnel Emission Studies ...................................... 4-20 
 
 
 

4.1.3. National Emission Estimates ................................... 4-28 
 
 
 

4.1.3.1. Activity Information for On-Road Vehicles ................ 4-28 
 
 
 

4.1.3.2. Activity Information for Off-Road Uses ................... 4-31 
 
 
 

4.1.3.3 Emission Estimates .................................... 4-32 
 
 
 

4.2. WOOD COMBUSTION ............................................ 4-38 
 
 
 

4.2.1. Flue Emissions from Wood Combustion (Residential) ............... 4-38 
 
 
 

4.2.1.1. Emissions Data ....................................... 4-38 
 
 
 

4.2.1.2. Activity Level Information .............................. 4-41 
 
 
 

4.2.1.3. Emission Estimates ................................... 4-44 
 
 
 

4.2.2. Stack Emissions from Wood Combustion (Industrial) ............... 4-44 
 
 
 

4.2.2.1. Emissions Data ....................................... 4-44 
 
 
 

4.2.2.2. Activity Level Information .............................. 4-50 
 
 
 

4.2.2.3. Emission Estimates ................................... 4-51 
 
 
 

4.2.3. Solid Waste from Wood Combustion (Residential and Industrial) ...... 4-52 
 
 
 

4.3. OIL COMBUSTION ............................................... 4-60 
 
 
 

4.3.1. Institutional/Commercial and Residential Oil Combustion ............ 4-61 
 
 
 

4.3.2. Utility Sector and Industrial Oil Combustion ...................... 4-64 
 
 
 

4.3.3. Used Oil Combustion ......................................... 4-67 
 
 
 

4.4. COAL COMBUSTION ............................................. 4-67 
 
 
 

4.4.1. Utilities and Industrial Boilers .................................. 4-68 
 
 
 

4.4.2. Residential Coal Combustion .................................. 4-74 
 
 
 

4.4.3. Solid Wastes from Coal Combustion ............................. 4-77 
 
 
 

5. 	 COMBUSTION SOURCES OF CDDs/CDFs: OTHER HIGH-TEMPERATURE
 
 

SOURCES ............................................................. 5-1 
 
 

5.1. CEMENT KILNS .................................................. 5-1 
 
 
 

5.1.1. Process Description of Portland Cement Kilns ....................... 5-1 
 
 
 

5.1.2. Cement Kilns That Burn Hazardous Waste ......................... 5-2 
 
 
 

v 
CONTENTS (continued) 

5.1.3. Air Pollution Control Devices ................................... 5-3 
 
 
 

5.1.4. CDD/CDF Emissions Data ...................................... 5-4 
 
 
 

5.1.4.1. Emissions Data for 1989 through 1996 ..................... 5-4 
 
 
 

5.1.4.2. Emissions Data for 1999 and 2000 ......................... 5-7 
 
 
 

5.1.4.3. Emission Factor Estimates for Cement Kilns Burning Hazardous
 
 
 

Waste ............................................... 5-7 
 
 
 

5.1.4.4. Emission Factor Estimates for Cement Kilns Burning
 
 
 

Nonhazardous Waste ................................... 5-9 
 
 
 

5.1.4.5. Confidence Ratings of Emission Factor Estimates ........... 5-10 
 
 
 

5.1.5. Activity Level Information ..................................... 5-10 
 
 
 

5.1.6. National CDD/CDF Emission Estimates .......................... 5-14 
 
 
 

5.1.6.1. Estimates for Reference Years 1987 and 1995 ............... 5-14 
 
 
 

5.1.6.2. Estimates for Reference Year 2000 ....................... 5-15 
 
 
 

5.1.7. EPA Regulatory Activities ..................................... 5-16 
 
 
 

5.1.8. Solid Waste from Cement Manufacturing: Cement Kiln Dust ......... 5-16 
 
 
 

5.2. LIGHTWEIGHT AGGREGATE KILNS ............................... 5-21 
 
 
 

5.3. ASPHALT MIXING PLANTS ....................................... 5-22 
 
 
 

5.4. PETROLEUM REFINING CATALYST REGENERATION ................ 5-24 
 
 
 

5.5. CIGARETTE SMOKING ........................................... 5-28 
 
 
 

5.6. PYROLYSIS OF BROMINATED FLAME RETARDANTS ............... 5-39 
 
 
 

5.7. CARBON REACTIVATION FURNACES ............................. 5-39 
 
 
 

5.8. KRAFT BLACK LIQUOR RECOVERY BOILERS ...................... 5-42 
 
 
 

5.9. OTHER IDENTIFIED SOURCES .................................... 5-46 
 
 
 

6. 	 COMBUSTION SOURCES OF CDDs/CDFs: MINIMALLY CONTROLLED
 
 

AND UNCONTROLLED COMBUSTION SOURCES .......................... 6-1 
 
 

6.1. COMBUSTION OF LANDFILL GAS .................................. 6-1 
 
 
 

6.1.1. Emissions Data ............................................... 6-1 
 
 
 

6.1.2. Activity Level Information ...................................... 6-1 
 
 
 

6.1.2.1. Activity Levels for 1987 and 1995 ......................... 6-3 
 
 
 

6.1.2.2. Activity Level for 2000 ................................. 6-4 
 
 
 

6.1.3. Emission Estimates ........................................... 6-4 
 
 
 

6.2. ACCIDENTAL FIRES .............................................. 6-5 
 
 
 

6.2.1. Soot and Ash Studies .......................................... 6-5 
 
 
 

6.2.2. Fume and Smoke Studies ....................................... 6-7 
 
 
 

6.2.3. Data Evaluation ............................................... 6-8 
 
 
 

6.2.3.1. Structural Fires ........................................ 6-8 
 
 
 

6.2.3.1.1. Emissions data ................................ 6-8 
 
 
 

6.2.3.1.2. Activity level information ....................... 6-9 
 
 
 

6.2.3.1.3. Emission estimates ............................ 6-9 
 
 
 

6.2.3.2. Vehicle Fires ......................................... 6-10 
 
 
 

6.3. LANDFILL FIRES ................................................. 6-11 
 
 
 

6.3.1. Emissions Data .............................................. 6-11 
 
 
 

vi 
CONTENTS (continued) 

6.3.2. Activity Level Information and Emission Estimates ................. 6-12 
 
 
 

6.4. FOREST AND BRUSH FIRES ....................................... 6-12 
 
 
 

6.4.1. Emissions Data .............................................. 6-12 
 
 
 

6.4.2. Activity Level Information ..................................... 6-18 
 
 
 

6.4.2.1. Approach for Reference Year 2000 (Office of Air Quality
 
 
 

Planning and Standards [OAQPS]) ....................... 6-18 
 
 
 

6.4.2.2. Approach for Reference Years 1987 and 1995 ............... 6-18 
 
 
 

6.4.3. Emission Estimates ........................................... 6-19 
 
 
 

6.5. BACKYARD BARREL BURNING ................................... 6-20 
 
 
 

6.5.1. Emissions Data .............................................. 6-20 
 
 
 

6.5.2. Activity Level Information ..................................... 6-21 
 
 
 

6.5.2.1. Summary of Barrel Burn Surveys ......................... 6-23 
 
 
 

6.5.2.2. Estimates of Activity Level ............................. 6-24 
 
 
 

6.5.2.3. Alternative Approach to Estimating Activity Level ........... 6-25 
 
 
 

6.5.3. Emission Estimates ........................................... 6-26 
 
 
 

6.5.4. Composition of Ash from Barrel Burning ......................... 6-26 
 
 
 

6.6. RESIDENTIAL YARD WASTE BURNING ............................ 6-26 
 
 
 

6.6.1. Emissions Data .............................................. 6-26 
 
 
 

6.6.2. Activity Level Information ..................................... 6-28 
 
 
 

6.6.3. Emission Estimates ........................................... 6-29 
 
 
 

6.7. LAND-CLEARING DEBRIS BURNING .............................. 6-30 
 
 
 

6.7.1. Emissions Data .............................................. 6-30 
 
 
 

6.7.2. Activity Level Information ..................................... 6-30 
 
 
 

6.7.2.1. Residential Construction ............................... 6-30 
 
 
 

6.7.2.2. Nonresidential Construction ............................. 6-30 
 
 
 

6.7.2.3. Roadway Construction ................................. 6-31 
 
 
 

6.7.2.4. Fuel Loading Factors .................................. 6-31 
 
 
 

6.7.3. Emission Estimates ........................................... 6-32 
 
 
 

6.8. 	 UNCONTROLLED COMBUSTION OF POLYCHLORINATED 
 
 


BIPHENYLS ..................................................... 6-32 
 
 
 

6.9. 	 VOLCANOES .................................................... 6-33 
 
 
 

6.10. FIREWORKS .................................................... 6-34 
 
 
 

6.11. OPEN BURNING AND OPEN DETONATION OF ENERGETIC 
 
 


MATERIALS ..................................................... 6-36 
 
 
 

7. METAL SMELTING AND REFINING SOURCES OF CDDs/CDFs .............. 7-1 
 
 
 

7.1. PRIMARY NONFERROUS METAL SMELTING/REFINING ............... 7-1 
 
 
 

7.1.1. Primary Copper Smelting and Refining .......................... 7-1 
 
 
 

7.1.2. Primary Magnesium Smelting and Refining ....................... 7-4 
 
 
 

7.1.3. Primary Nickel Smelting and Refining ........................... 7-6 
 
 
 

7.1.4. Primary Aluminum Smelting and Refining ........................ 7-6 
 
 
 

7.1.5. Primary Titanium Smelting and Refining ......................... 7-7 
 
 
 

7.2. SECONDARY NONFERROUS METAL SMELTING .................... 7-8 
 
 
 

vii 
CONTENTS (continued) 

7.2.1. Secondary Aluminum Smelters ................................ 7-8 
 
 
 

7.2.2. Secondary Copper Smelters ................................. 7-14 
 
 
 

7.2.2.1. Emissions Data ................................... 7-15 
 
 
 

7.2.2.2. Activity Level Information ............................ 7-20 
 
 
 

7.2.2.3. Emission Estimates ................................ 7-20 
 
 
 

7.2.3. Secondary Lead Smelters ................................... 7-22 
 
 
 

7.3. PRIMARY FERROUS METAL SMELTING/REFINING ................. 7-27 
 
 
 

7.3.1. Sinter Production ........................................ 7-27 
 
 
 

7.3.2. Coke Production ......................................... 7-32 
 
 
 

7.4. SECONDARY FERROUS METAL SMELTING/REFINING .............. 7-33 
 
 
 

7.5. FERROUS FOUNDRIES ........................................ 7-36 
 
 
 

7.6. 	 SCRAP ELECTRIC WIRE RECOVERY ............................. 7-40 
 
 
 

7.7. 	 DRUM AND BARREL RECLAMATION FURNACES ................... 7-45 
 
 
 

7.8. 	 SOLID WASTE FROM PRIMARY/SECONDARY IRON/STEEL
 
 
 

MILLS/FOUNDRIES ........................................... 7-48 
 
 
 

8. CHEMICAL MANUFACTURING AND PROCESSING SOURCES ............... 8-1 
 
 
 

8.1. BLEACHED CHEMICAL WOOD PULP AND PAPER MILLS .............. 8-1 
 
 
 

8.1.1. Estimates of National Emissions in 1987 and 1995 ................... 8-8 
 
 
 

8.1.2. Estimates of National Emissions in 2000 ........................... 8-8 
 
 
 

8.2. 	 MANUFACTURE OF CHLORINE, CHLORINE DERIVATIVES, AND METAL
 
 
 

CHLORIDES ...................................................... 8-9 
 
 
 

8.2.1. Manufacture of Chlorine ........................................ 8-9 
 
 
 

8.2.2. Manufacture of Chlorine Derivatives and Metal Chlorides ............ 8-12 
 
 
 

8.3. MANUFACTURE OF HALOGENATED ORGANIC CHEMICALS ......... 8-14 
 
 
 

8.3.1. Chlorophenols ............................................... 8-15 
 
 
 

8.3.1.1. Regulatory Actions for Chlorophenols ..................... 8-16 
 
 
 

8.3.2. Chlorobenzenes .............................................. 8-24 
 
 
 

8.3.2.1. Regulatory Actions for Chlorobenzenes .................... 8-26 
 
 
 

8.3.3. Chlorobiphenyls ............................................. 8-27 
 
 
 

8.3.4. Ethylene Dichloride/Vinyl Chloride Monomer/Polyvinyl Chloride
 
 
 

Manufacturing .............................................. 8-30 
 
 
 

8.3.4.1. Water Releases ....................................... 8-33 
 
 
 

8.3.4.2. Land Releases ........................................ 8-36 
 
 
 

8.3.4.3. Air Releases ......................................... 8-36 
 
 
 

8.3.4.4. Transfers to Secure Landfills ............................ 8-38 
 
 
 

8.3.4.5. Products ............................................ 8-41 
 
 
 

8.3.5. Other Aliphatic Chlorine Compounds ............................ 8-44 
 
 
 

8.3.6. Dyes, Pigments, and Printing Inks ............................... 8-45 
 
 
 

8.3.6.1. Dioxazine Dyes and Pigments ........................... 8-45 
 
 
 

8.3.6.2. Phthalocyanine Dyes and Printing Inks .................... 8-48 
 
 
 

8.3.7. TSCA Dioxin/Furan Test Rule .................................. 8-50 
 
 
 

8.3.8. Halogenated Pesticides and FIFRA Pesticides Data Call-In ............ 8-52 
 
 
 

viii 
CONTENTS (continued) 

8.4. 	 OTHER CHEMICAL MANUFACTURING AND PROCESSING 
 
 


SOURCES ....................................................... 8-73 
 
 
 

8.4.1. Municipal Wastewater Treatment Plants .......................... 8-73 
 
 
 

8.4.1.1. Sources ............................................. 8-73 
 
 
 

8.4.1.2. Releases to Water ..................................... 8-77 
 
 
 

8.4.1.2.1 Emissions data .............................. 8-77 
 
 
 

8.4.1.2.2 Activity level information ..................... 8-78 
 
 
 

8.4.1.2.3 Emission Estimates .......................... 8-81 
 
 
 

8.4.1.3. Sewage Sludge Land Disposal ........................... 8-81 
 
 
 

8.4.1.3.1 Emissions data .............................. 8-81 
 
 
 

8.4.1.3.2 Activity level information .................... 8-88 
 
 
 

8.4.1.3.3 Emission estimates .......................... 8-89 
 
 
 

8.4.2. Drinking Water Treatment Plants ............................... 8-91 
 
 
 

8.4.3. Soaps and Detergents ......................................... 8-93 
 
 
 

8.4.4. Textile Manufacturing and Dry Cleaning ......................... 8-94 
 
 
 

9. INDICATIONS OF POSSIBLE NATURAL SOURCES OF CDDs/CDFs ............. 9-1 
 
 
 

9.1. BIOLOGICAL SOURCES OF CDDs/CDFs ............................... 9-1 
 
 
 

9.1.1. Biotransformation of Chlorophenols ............................... 9-1 
 
 
 

9.1.2. Biotransformation of Higher CDDs/CDFs .......................... 9-4 
 
 
 

9.1.3. Dioxin-Like Compounds in Animal Manure ......................... 9-6 
 
 
 

9.2. PHOTOTRANSFORMATION OF CHLOROPHENOLS .................... 9-8 
 
 
 

9.2.1. Photolysis of Higher CDDs/CDFs ................................ 9-10 
 
 
 

9.2.2. Photolysis in Water ........................................... 9-11 
 
 
 

9.2.3. Photolysis on Soil ............................................ 9-11 
 
 
 

9.2.4. Photolysis on Vegetation ....................................... 9-12 
 
 
 

9.2.5. Photolysis in Air ............................................. 9-13 
 
 
 

9.3. CDDs/CDFs IN BALL CLAY ......................................... 9-14 
 
 
 

9.3.1. Initial Discovery of CDD/CDF Contamination of Ball Clay ............ 9-14 
 
 
 

9.3.2. Characteristics of Mississippi Embayment Ball Clays ................ 9-15 
 
 
 

9.3.3. Levels of Dioxin-Like Compounds in Ball Clay .................... 9-15 
 
 
 

9.3.4. Evidence for Ball Clay as a Natural Source ........................ 9-16 
 
 
 

9.3.5. Environmental Releases of Dioxin-Like Compounds from the Mining
 
 
 

and Processing of Ball Clay .................................... 9-19 
 
 
 

10. SOURCES OF DIOXIN-LIKE POLYCHLORINATED BIPHENYLS (PCBs) ........ 10-1 
 
 
 

10.1. GENERAL FINDINGS OF THE EMISSIONS INVENTORY ................ 10-1 
 
 
 

10.2. RELEASES OF COMMERCIAL PCBs ................................. 10-4 
 
 
 

10.2.1. Approved PCB Disposal/Destruction Methods .................... 10-11 
 
 
 

10.2.1.1. Approved Incinerators/High-Efficiency Boilers ............ 10-14 
 
 
 

10.2.1.2. Approved Chemical Waste Landfills ..................... 10-14 
 
 
 

10.2.1.3. Other Approved Disposal Methods ...................... 10-15 
 
 
 

10.2.2. Emission Estimates .......................................... 10-15 
 
 
 

ix 
CONTENTS (continued) 

10.2.3. Accidental Releases of Still-in-Service PCBs ..................... 10-15 
 
 
 

10.2.3.1. Leaks and Spills ..................................... 10-16 
 
 
 

10.2.3.2. Accidental Fires ..................................... 10-18 
 
 
 

10.2.4. Municipal Wastewater Treatment ............................... 10-19 
 
 
 

10.3. CHEMICAL MANUFACTURING AND PROCESSING SOURCES ......... 10-26 
 
 
 

10.4. COMBUSTION SOURCES .......................................... 10-26 
 
 
 

10.4.1. Municipal Waste Combustors .................................. 10-26 
 
 
 

10.4.2. Industrial Wood Combustion .................................. 10-27 
 
 
 

10.4.3. Medical Waste Incineration ................................... 10-29 
 
 
 

10.4.4. Tire Combustion ............................................ 10-30 
 
 
 

10.4.5. Cigarette Smoking .......................................... 10-31 
 
 
 

10.4.6. Sewage Sludge Incineration ................................... 10-31 
 
 
 

10.4.7. Backyard Barrel Burning ..................................... 10-34 
 
 
 

10.4.8. Petroleum Refining Catalyst Regeneration ........................ 10-34 
 
 
 

10.5. NATURAL SOURCES ............................................. 10-36 
 
 
 

10.5.1. Biotransformation of Other PCBs ............................... 10-36 
 
 
 

10.5.2. Photochemical Transformation of Other PCBs .................... 10-39 
 
 
 

10.6. PAST USE OF COMMERCIAL PCBs ................................. 10-41 
 
 
 

11. RESERVOIR SOURCES OF CDDs/CDFs AND DIOXIN-LIKE PCBs .............. 11-1 
 
 
 

11.1. POTENTIAL RESERVOIRS .......................................... 11-1 
 
 
 

11.2. CHARACTERIZATION OF RESERVOIR SOURCES ..................... 11-3 
 
 
 

11.2.1. Soil ....................................................... 11-3 
 
 
 

11.2.1.1. Potential Mass of Dioxin-Like Compounds Present ......... 11-3 
 
 
 

11.2.1.2. Mechanisms Responsible for Releases from Surface Soils .... 11-9 
 
 
 

11.2.1.3. Estimated Annual Releases from Soil to Water ............ 11-12 
 
 
 

11.2.1.4. Estimated Annual Releases from Soil to Air .............. 11-14 
 
 
 

11.2.2. Water ..................................................... 11-17 
 
 
 

11.2.2.1. Potential Mass of Dioxin-Like Compounds Present ........ 11-17 
 
 
 

11.2.2.2. Mechanisms Responsible for Supply to and Releases 
 
 


from Water ........................................ 11-17 
 
 
 

11.2.3. Sediment .................................................. 11-20 
 
 
 

11.2.3.1. Potential Mass of Dioxin-Like Compounds Present ........ 11-20 
 
 
 

11.2.3.2. Mechanisms Responsible for Supply to and Releases 
 
 


from Sediment ..................................... 11-20 
 
 
 

11.2.3.3. Releases from Sediment to Water ...................... 11-21 
 
 
 

11.2.4. Biota ..................................................... 11-22 
 
 
 

11.2.4.1. Potential Mass of Dioxin-Like Compounds Present ........ 11-22 
 
 
 

11.2.4.2. Mechanisms Responsible for Supply to and Releases 
 
 


from Biota ........................................ 11-22 
 
 
 

11.2.4.3. Approaches for Measuring and Estimating Releases 
 
 


from Biota ........................................ 11-24 
 
 
 

11.3. SUMMARY AND CONCLUSIONS ................................... 11-26 
 
 
 

x 
CONTENTS (continued) 

11.3.1. Reservoir Sources ........................................... 11-26 
 
 
 

11.3.2. Implications for Human Exposure .............................. 11-27 
 
 
 

REFERENCES ............................................................. R-1 
 
 
 

xi
 
LIST OF TABLES
 
 
 

Table 1-1. Nomenclature for dioxin-like compounds .............................. 1-2 
 

Table 1-2. Municipal waste combustorsa ....................................... 1-4 
 

Table 1-3. Hazardous waste incinerators and cement kilns and lightweight aggregate kilns
 

burning hazardous waste ........................................... 1-5 
 

Table 1-4. Cement kilns not burning hazardous waste ............................. 1-5 
 

Table 1-5. Secondary aluminum smelters ....................................... 1-6 
 

Table 1-6. Medical waste incinerators .......................................... 1-6 
 

Table 1-7. Pulp and paper mills ............................................... 1-6 
 

Table 1-8. The TEF scheme for I-TEQDF ....................................... 1-9 
 

Table 1-9. The TEF scheme for dioxin-like PCBs, as determined by the World Health
 

Organization in 1994 .............................................. 1-9 
 

Table 1-10. The TEF scheme for TEQDFP-WHO98 ................................ 1-10 
 

Table 1-11. Known and suspected sources of CDDs/CDFs ......................... 1-14 
 

Table 1-12. Inventory of contemporary releases (g/yr) of dioxin-like compounds from known
 

sources in the United States for reference years 2000, 1995, and 1987 and
 

preliminary release estimates for 2000 ............................... 1-17 
 

Table 1-13. Products containing CDDs/CDFs (g TEQDF-WHO98 /yr) .................. 1-23 
 

Table 1-14. I-TEQDF emission factors used to develop national emission inventory estimates of
 

releases to air ................................................... 1-32 
 

Table 1-15. TEQDF-WHO98 emission factors used to develop national emission inventory
 

estimates of releases to air ......................................... 1-34 
 

Table 1-16. Confidence rating scheme for U.S. emission estimates ................... 1-36 
 

Table 1-17. Ranking of sources of dioxin-like compounds based on environmental releases
 

(from high to low) for reference years 2000, 1995, and 1987 .............. 1-38 
 

Table 2-1. Concentration of CDDs/CDFs on municipal incinerator fly ash at varying
 

temperatures ..................................................... 2-7 
 

xii
 
LIST OF TABLES (continued) 

Table 2-2. 	 	 	 CDDs/CDFs formed from the thermolytic reaction of 690 mg benzene + FeCl3
 

silica complex .................................................. 2-12 
 

Table 2-3. 	 	 	 De novo formation of CDDs/CDFs after heating Mg-Al silicate, 4% charcoal, 7%
 

Cl, 1% CuCl2 in H2O at 300°C ..................................... 2-14 
 

Table 3-1. 	 	 	 Inventory of municipal waste combustors (MWCs) in 2000 by technology, air
 

pollution control device (APCD), sizea, and annual activity level (kg/yr) ..... 3-14 
 

Table 3-2. 	 	 	 Inventory of municipal waste combustors (MWCs) in 1995 by technology, air
 

pollution control device (APCD), and annual activity level (kg/yr) ......... 3-16 
 

Table 3-3. 	 	 	 Inventory of municipal waste combustors (MWCs) in 1987 by technology, air
 

pollution control device (APCD), and annual activity level (kg/yr) ......... 3-18 
 

Table 3-4. 	 	 	 National average CDD/CDF congener concentrations for
 

large municipal waste combustors (ng/dscm @ 7% O2) .................. 3-20 
 

Table 3-6. 	 	 	 CDD/CDF TEQ emission factors (ng TEQ/kg waste) for municipal solid waste
 

incineration .................................................... 3-24 
 

Table 3-7a. 	 	 	Annual I-TEQDF emissions from municipal waste combustors (MWCs) operating
 

in 1995 ........................................................ 3-25 
 

Table 3-7b. 	 	 Annual TEQDF-WHO98 emissions from municipal waste combustors (MWCs)
 

operating in 1995 ................................................ 3-27 
 

Table 3-8a. 	 	 	Annual I-TEQDF emissions to the air from municipal waste combustors (MWCs)
 

operating in 1987 ................................................ 3-29 
 

Table 3-8b. 	 	 Annual TEQDF-WHO98 emissions to the air from municipal waste combustors
 

(MWCs) operating in 1987 ........................................ 3-30 
 

Table 3-9. 	 	 	 Average and range of CDD/CDF congener groups in fly ash from a municipal
 

incinerator (:g/kg) .............................................. 3-36 
 

Table 3-10. 	 	 Comparison of the amount of TEQs generated annually in municipal waste
 

combustor ash .................................................. 3-38 
 

Table 3-11. 	 	 Concentration of CDD/CDF congener groups (ng/kg) in fly ash samples from
 

combustion of municipal solid waste in eight Japanese incinerators ......... 3-39 
 

xiii 
LIST OF TABLES (continued) 

Table 3-12. Concentration of CDD/CDF congener groups in fly ash samples from municipal 

solid waste ..................................................... 3-39 
 

Table 3-13. CDD/CDF concentrations in municipal solid waste ash from a newly constructed
 

municipal waste combustor in Japan ................................. 3-40 
 

Table 3-14a. CDD/CDF emission factors for hazardous waste incinerators and boilers tested
 

from 1993 to 1996 ............................................... 3-48 
 

Table 3-14b. CDD/CDF emission factors for hazardous waste incinerators and boilers tested in
 

2000 .......................................................... 3-49 
 

Table 3-15. CDD/CDF emission factors for halogen acid furnaces tested in 2000 ....... 3-54 
 

Table 3-16. Estimated breakdown of facilities by air pollution control device (APCD) . . . 3-60
 

Table 3-17. Summary of annual operating hours for each medical waste incinerator (MWI) 
 

type ........................................................... 3-61 
 

Table 3-18. TEQ emissions from medical waste incinerators (MWIs) for reference 
 

year 1987 ...................................................... 3-62 
 

Table 3-19. TEQ emissions from medical waste incinerators (MWIs) for reference 
 

year 1995 ...................................................... 3-62 
 

Table 3-20. TEQ emissions from medical waste incinerators (MWIs) for reference 
 

year 2000 ...................................................... 3-63 
 

Table 3-21. Congener-specific profile for Camellia Memorial Lawn crematorium ....... 3-70 
 

Table 3-22. Congener-specific profile for the Woodlawn Cemetery crematorium ........ 3-71 
 

Table 3-23. Operational data for the Woodlawn Cemetery crematorium, scrubber inlet . . . 3-73
 

Table 3-24. Congener-specific profile for the Camellia Memorial Lawn crematorium and
 

the Woodlawn Cemetery crematorium ............................... 3-74 
 

Table 3-25. Congener-specific profile for the University of Georgia Veterinary School . . . 3-77
 

Table 3-26. CDD/CDF emission factors for sewage sludge incinerators ............... 3-82 
 

Table 3-27. CDD/CDF air emission factors for a tire combustion facility .............. 3-84 
 

xiv 
LIST OF TABLES (continued) 

Table 3-28. CDD/CDF emission factors for combustion of bleached-kraft mill sludge in 

wood residue boilers ............................................. 3-88 
 

Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste
 

combustors (MWCs) operating in 2000 .............................. 3-90 
 

Table 4-1. Description and results of vehicle emission testing studies for CDDs 
 

and CDFs ....................................................... 4-3 
 

Table 4-2. CDD/CDF congener emission factors (pg/L) for diesel-fueled automobiles .... 4-5 
 

Table 4-3. CDD/CDF congener emission factors (pg/L) for diesel-fueled trucks ........ 4-7 
 

Table 4-4. CDD/CDF congener emission factors (pg/L) for leaded gasoline-fueled
 

automobiles ..................................................... 4-9 
 

Table 4-5. CDD/CDF congener emission factors (pg/L) for unleaded gasoline-fueled
 

automobiles without catalytic converters .............................. 4-11 
 

Table 4-6. CDD/CDF congener emission factors (pg/L) for unleaded gasoline-fueled
 

automobiles with catalytic converters ................................ 4-13 
 

Table 4-7. Total dioxin emission concentrations from heavy-duty diesel engines 
 

in Germany ..................................................... 4-21 
 

Table 4-8. Levels of 2,3,7,8-chlorine-substituted congeners and total CDDs/CDFs in
 

vehicle exhaust particles (pg/g) for gasoline engines and suspended 

particulate matter (SPM) .......................................... 4-22 
 

Table 4-9. Levels of 2,3,7,8-chlorine-substituted congeners and total CDDs/CDFs in
 

vehicle exhaust particles (pg/g) for diesel engines ...................... 4-23 
 

Table 4-10. CDD/CDF concentrations (pg/m3) from European tunnel air studies ........ 4-24 
 

Table 4-11. Baltimore Harbor tunnel study: estimated emission factors (pg/km) for heavy-
 

duty diesel trucks ................................................ 4-26 
 

Table 4-12. Average CDD/CDF concentration in flue gas while burning wood and 
 

crops .......................................................... 4-40 
 

Table 4-13. CDD/CDF concentrations (pg TEQ/kg wood) in emissions from residential
 

wood stoves in Canada ............................................ 4-42 
 
 
 

xv 
LIST OF TABLES (continued) 

Table 4-14. CDD/CDF mean emission factors (ng/kg wood) for industrial wood 

combustors ..................................................... 4-45 
 

Table 4-15. NCASI CDD/CDF TEQ concentrations and emissions for wood 
 

residue-fired boilers .............................................. 4-50 
 

Table 4-16. CDD/CDF concentrations (ng/kg) in residential chimney soot from 
 

wood stoves and fireplaces ........................................ 4-53 
 

Table 4-17. CDD/CDF concentrations (ng/kg) in bottom ash from residential wood stoves 
 

and fireplaces ................................................... 4-54 
 

Table 4-18. CDD/CDF concentrations in chimney soot (Bavaria, Germany) ............ 4-56 
 

Table 4-19. CDD/CDF concentrations (ng/kg) in fly ash from wood-working industry . . . 4-57
 

Table 4-20. CDD/CDF concentrations (ng/kg) in electrostatic precipitator waste ash from
 

wood-fired industrial boiler ........................................ 4-59 
 

Table 4-21. Estimated CDD/CDF emission factors for oil-fired residential furnaces ..... 4-62 
 

Table 4-22. CDD/CDF emission factors (pg/L oil) for oil-fired utility/industrial boilers . . 4-65
 

Table 4-23. CDD/CDF concentrations (pg/Nm3) in stack emissions from U.S. coal-fired
 

power plants .................................................... 4-70 
 

Table 4-24. Characteristics of U.S. coal-fired power plants tested by the U.S. Department
 

of Energy ...................................................... 4-71 
 

Table 4-25. CDD/CDF emission factors (ng/kg coal) for coal-fired utility/industrial 
 

power plants .................................................... 4-72 
 

Table 4-26. CDD/CDF emission factors (ng/kg coal) for residential coal combustors .... 4-75 
 

Table 4-27. CDD/CDF concentrations and amounts disposed of in solid wastes from coal-
 

fired utilities .................................................... 4-79 
 

Table 5-1. CDD/CDF emission factors (ng/kg clinker produced) for cement kilns
 

burning hazardous waste for reference years 1987 and 1995 ............... 5-8 
 

Table 5-2. CDD/CDF emission factors (ng/hr) for cement kilns burning hazardous waste
 

for reference year 2000 ........................................... 5-10 
 
 
 

xvi 
LIST OF TABLES (continued) 

Table 5-3. CDD/CDF emission factors (ng/kg clinker produced) for cement kilns not 

burning hazardous waste for reference years 1987, 1995, and 2000 ......... 5-12 
 

Table 5-4. National emission estimates for cement kilns for reference years 
 

1987 and 1995 .................................................. 5-15 
 

Table 5-5. National emission estimates for cement kilns for reference year 2000 ....... 5-17 
 

Table 5-6. CDD/CDF concentrations (ng/kg) in ash samples from cement kiln
 

electrostatic precipitator and lightweight aggregate (LWA) kiln fabric filter . . 5-20
 

Table 5-7. CDD/CDF estimates in cement kiln dust (CKD) for reference years 1987,
 

1995, and 2000 .................................................. 5-21 
 

Table 5-8. Congener-specific emission factors (ng/kg asphalt produced) derived from the
 

testing of air emissions at two hot-mix asphalt plants in the United States .... 5-24 
 

Table 5-9. CDD/CDF emission factors (ng/barrel)a for petroleum catalytic reforming
 

units .......................................................... 5-29 
 

Table 5-10. CDD concentrations in Japanese cigarettes, smoke, and ash ............... 5-32 
 

Table 5-11. CDD/CDF concentrations (pg/cig) in cigarette smoke, normalized to a per-
 

cigarette basis ................................................... 5-34 
 

Table 5-12. CDD/CDF concentrations (pg/pack) in tobacco cigarette brands from various
 

countries ...................................................... 5-37 
 

Table 5-13. CDD/CDF mean emission factors (ng/kg feed) for black liquor 
 

recovery boilers ................................................. 5-44 
 

Table 5-14. CDD/CDF TEQ emission factors and emission estimates from Kraft recovery
 

furnaces and Kraft lime kilns ....................................... 5-47 
 
 
 

Table 5-15. CDD/CDF concentrations in candle materials and emissions .............. 5-47 
 
 

Table 6-1. CDD/CDF emission factors for a landfill flare .......................... 6-2 
 
 
 

Table 6-2 CDD/CDF mean emission factors (ng/kg) for forest fires ................. 6-16 
 
 
 

Table 6-3. Forest fire fuel loading factors (tons/acre) ............................. 6-19 
 
 

xvii
 
LIST OF TABLES (continued) 

Table 6-4. CDD/CDF average air emission factors (ng/kg waste burned) from barrel 

burning of household waste ........................................ 6-22 
 

Table 6-5. CDD/CDF analysis for composite ash samples from barrel 
 

burning (ng/kg ash) .............................................. 6-27 
 

Table 6-6. PCB analysis for composite ash samples from barrel burning (ng/kg ash) .... 6-28 
 

Table 6-7. CDDs/CDFs in dust fall and ashes from volcanoes ...................... 6-34 
 

Table 6-8. Residue of HpCDD/HpCDF and OCDD/OCDF (ng/kg) in paper cartridges
 

and charges of select pyrotechnic products ............................ 6-35 
 

Table 7-1. CDD/CDF emission concentrations (pg TEQ/m3 @ 11% oxygen) for primary
 

copper smelters .................................................. 7-2 
 

Table 7-2. CDD/CDF emissions data from primary and secondary copper and secondary
 

lead smelters ..................................................... 7-3 
 

Table 7-3. CDD/CDF emission factors (ng/kg scrap feed) for secondary aluminum
 

smelters for 2000 ................................................ 7-11 
 

Table 7-4. CDD/CDF emission factors (ng/kg scrap feed) for secondary aluminum
 

smelters for 1995 and 1987 ........................................ 7-12 
 

Table 7-5. CDD/CDF mean emission factors (ng/kg scrap feed) for secondary copper 
 

smelters ....................................................... 7-17 
 

Table 7-6. CDD/CDF emission factors (ng/kg lead produced) for secondary
 

lead smelters .................................................... 7-25 
 
 


Table 7-7. Estimated annual TEQ emissions (g TEQ) ............................ 7-27 
 
 


Table 7-8. CDD/CDF emission factors (ng/kg sinter) for sintering plants ............. 7-30 
 
 


Table 7-9. Operating parameters for U.S. iron ore sintering plants .................. 7-32 
 
 


Table 7-10. CDD/CDF emission estimates for Canadian coke oven facilities, blast furnace
 
 


facilities, and electric arc furnaces ................................... 7-34 
 

Table 7-11. CDD/CDF emission concentrations and rates for Canadian electric arc 
 

furnaces ....................................................... 7-37 
 


xviii 
LIST OF TABLES (continued) 

Table 7-12. CDD/CDF emission factors for a U.S. ferrous foundry ................... 7-39 
 

Table 7-13. Congener-specific profile for ferrous foundries ......................... 7-40 
 

Table 7-14. CDD/CDF emission factors for a scrap wire incinerator .................. 7-42 
 

Table 7-15. CDD/CDF concentrations in fly ash and ash/soil at metal recovery sites ..... 7-44 
 

Table 7-16. CDD/CDF emission factors for a drum and barrel reclamation facility ...... 7-46 
 

Table 8-1. CDD/CDF concentrations in pulp and paper mill bleached pulp, wastewater
 

sludge, and wastewater effluent (circa 1988) ............................ 8-2 
 

Table 8-2. CDD/CDF concentrations in pulp and paper mill bleached pulp, wastewater
 

sludge, and wastewater effluent (mid-1990s) ........................... 8-5 
 

Table 8-3. Summary of bleached chemical pulp and paper mill discharges (g/yr) of 
 

2,3,7,8-TCDD and 2,3,7,8-TCDF .................................... 8-7 
 

Table 8-4. CDD/CDF TEQ concentrations and emissions for the paper and pulp industry by
 

source ......................................................... 8-10 
 

Table 8-5. CDD/CDF concentrations (μg/kg) in graphite electrode sludge from chlorine
 

production ..................................................... 8-11 
 

Table 8-6. Releases of dioxin-like compounds in wastewater discharges from chlor-alkali and
 

mixed chemical manufacturing facilities to surface water in reference years 2000
 

and 1995 ....................................................... 8-13 
 

Table 8-7. Congener-specific and TEQ annual releases to air (g/yr) from chlor-alkali and
 

mixed chemical production facilities in 2000 .......................... 8-14 
 

Table 8-8. CDD/CDF concentrations (μg/kg) in metal chlorides .................... 8-15 
 

Table 8-9. CDD/CDF concentrations (mg/kg) in mono- through tetrachlorophenols .... 8-17 
 

Table 8-10. CDD/CDF concentrations (historical and current) (μg/kg) in technical-grade
 

pentachlorophenol (PCP) products .................................. 8-18 
 

Table 8-11. Historical CDD/CDF concentrations (μg/kg) in pentachlorophenol-Na 
 

(PCP-Na) ...................................................... 8-20 
 
 


xix 
LIST OF TABLES (continued) 

Table 8-12. 	 	 Summary of specific dioxin-containing wastes that must comply with land
 

disposal restrictions .............................................. 8-21 
 

Table 8-13. 	 	 CDD/CDF concentrations (μg/kg) in chlorobenzenes .................... 8-25 
 


Table 8-14. 	 	 Concentrations of CDD/CDF congener groups in unused commercial
 

polychlorinated biphenyl (PCB) mixtures (mg/kg) ...................... 8-28 
 

Table 8-15. 	 	 2,3,7,8-Substituted congener concentrations in unused polychlorinated biphenyl
 

(PCB) mixtures (μg/kg) ........................................... 8-29 
 

Table 8-16. 	 	 Reported CDD/CDF concentrations (μg/kg) in wastes from polyvinyl chloride
 

(PVC) manufacture .............................................. 8-32 
 

Table 8-17. 	 	 Releases of dioxin-like compounds (g/yr) in wastewater discharges from 

EDC/VCM/PVC and integrated chlorine chemical manufacturing facilities to 

surface water in reference years 2000 and 1995 ........................ 8-34 
 
 


Table 8-18. 	 	 Congener-specific releases to land from an EDC/VCM/PVC integrated chemical
 

manufacturing facilitya in reference years 2000 and 1995 ................. 8-37 
 

Table 8-19. 	 	 Congener-specific and TEQ releases to air (g/yr) from EDC/VCM/PVC integrated 
 

chemical manufacturing facilities in reference years 2000 and 1995 ........ 8-39 
 

Table 8-20. 	 	 Congener-specific and TEQ transfers to secure landfills (g/yr)a from
 

EDC/VCM/PVC integrated chemical production facilities in 2000 ......... 8-40 
 

Table 8-21. 	 	 CDD/CDF concentrations in products from U.S. EDC/VCM/PVC 
 

manufacturers ................................................... 8-42 
 

Table 8-22. 	 	 CDD/CDF concentrations (μg/kg) in samples of dioxazine dyes and pigments
 

(Canada) ....................................................... 8-46 
 

Table 8-23. 	 	 CDD/CDF concentrations (ng/kg) in printing inks (Germany) ............. 8-49 
 


Table 8-24. 	 	 Chemicals requiring Toxic Substances Control Act Section 4 testing under the
 

dioxin/furan rule ................................................ 8-51 
 

Table 8-25. 	 	 Congeners and limits of quantitation (LOQs) for which quantitation is required
 

under the dioxin/furan test rule and pesticide Data Call-In ................ 8-52 
 

Table 8-26. 	 	 Precursor chemicals subject to reporting requirements under Toxic Substances
 

Control Act Section 8(a) .......................................... 8-53 
 

xx 
LIST OF TABLES (continued) 

Table 8-27. 	 	 Results of analytical testing for dioxins and furans in the chemicals tested to date
 

under Section 4 of the dioxin/furan test rule ........................... 8-54 
 

Table 8-28. 	 	 CDD/CDF concentrations (μg/kg) in chloranil and carbazole violet samples
 

analyzed pursuant to the EPA dioxin/furan test rule ..................... 8-55 
 

Table 8-29. 	 	 Status of first pesticide Data Call-In: pesticides suspected of having the potential 

to become contaminated with dioxins if synthesized under conditions favoring 

dioxin formation ................................................ 8-62 
 


Table 8-30. 	 	 Status of second pesticide Data Call-In: pesticides suspected of being
 

contaminated with dioxins ......................................... 8-67 
 

Table 8-31. 	 	 Summary of analytical data submitted to EPA in response to pesticide Data CallIn(s) .......................................................... 8-71 
 

Table 8-32. 	 	 Summary of results for CDDs/CDFs in technical 2,4-D and 2,4-D ester 
 

herbicides ...................................................... 8-72 
 

Table 8-33. 	 	 CDD/CDF concentrations (μg/kg) in samples of 2,4-D and pesticide formulations
 

containing 2,4-D ................................................ 8-74 
 

Table 8-34. 	 	 Mean CDD/CDF concentrations and range of detection limits (DLs) (pg/L) in
 

effluents from nine U.S. publicly owned treatment works (POTWs) ........ 8-79 
 

Table 8-35. 	 	 Effluent concentrations (pg/L) of CDDs/CDFs from publicly owned treatment
 

works in Mississippi ............................................. 8-80 
 

Table 8-36. 	 	 CDD/CDF concentrations (ng/kg) measured in the 1998/1999 National Sewage
 

Sludge Survey .................................................. 8-82 
 

Table 8-37. 	 	 CDD/CDF concentrations (ng/kg) measured in 99 sludges collected from U.S.
 

publicly owned treatment works (POTWs) during 1994 .................. 8-84 
 

Table 8-38. 	 	 Concentrations of CDDs/CDFs (ng/kg dry matter) in sewage sludge from publicly
 

owned treatment works in Mississippi ................................ 8-85 
 

Table 8-39. 	 	 CDD/CDF median concentrations (ng/kg) measured in 1999 from a publicly
 

owned treatment works facility in Ohio ............................... 8-86 
 

Table 8-40. 	 	 CDD/CDF mean concentrations (ng/kg) measured in the 2001 National Sewage
 

Sludge Survey .................................................. 8-87 
 

xxi 
LIST OF TABLES (continued) 

Table 8-41. 	 	 Quantity of sewage sludge disposed of and potential dioxin TEQ releases from
 

primary, secondary, and advanced treatment publicly owned treatment works for
 

reference year 1987 .............................................. 8-90 
 

Table 8-42. 	 	 Quantity of sewage sludge disposed of and potential dioxin TEQ releases from
 

primary, secondary, and advanced treatment publicly owned treatment works for
 

reference year 1995 .............................................. 8-91 
 

Table 8-43. 	 	 Quantity of sewage sludge disposed of and potential dioxin TEQ releases from
 

primary, secondary, and advanced treatment publicly owned treatment works for
 

reference year 2000 .............................................. 8-92 
 

Table 8-44. 	 	 Biosolids disposal practices for reference year 2000 ..................... 8-92 
 
 


Table 8-45. 	 	 CDD/CDF concentrations in Swedish liquid soap, tall oil, and tall resin ..... 8-95 
 

Table 9-1. 	 	 Estimated quantity of animal manure produced in the United States in 2000 . . . 9-7
 

Table 9-2. 	 	 CDD and CDF concentrations (ng/kg dry weight) in samples of animal manure in
 

the United Kingdom ............................................... 9-8 
 

Table 9-3. 	 	 Concentrations of CDDs (pg/g, dry weight) in eight ball clay samples in the United
 

States ......................................................... 9-16 
 

Table 9-4. 	 	 Comparison of the mean CDD/CDF congener group concentrations in ball clay
 

with those in urban and rural soils in North America (pg/g, dry weight) ..... 9-17 
 

Table 10-1. 	 	 Confidence rating classes for 2000 for releases from all known and suspected
 

source categories of dioxin-like PCBs ................................ 10-2 
 

Table 10-2. 	 	 Inventory of contemporary releases of dioxin-like PCBs in the United States for
 

1987, 1995, and 2000 and preliminary release estimates of dioxin-like PCBs for
 

2000 (g TEQp-WHO98/yr) ......................................... 10-3 
 

Table 10-3. 	 	 Weight percent concentrations of dioxin-like PCBs in Aroclors, Clophens, and
 

Kanechlors ..................................................... 10-7 
 

Table 10-4. 	 	 Disposal requirements for PCBs and PCB items ....................... 10-12 
 
 


Table 10-5. 	 	 Off-site transfers of PCBs reported in the Toxics Release Inventory (TRI)
 

(1988–2000) ................................................... 10-16 
 

xxii 
LIST OF TABLES (continued) 

Table 10-6. 	 	 Releases of PCBs reported in the Toxics Release Inventory (TRI) 
 

(1988–2000) ................................................. 10-17 
 

Table 10-7. 	 	 Aroclor concentrations (ng/kg) measured in EPA’s National Sewage
 

Sludge Survey ................................................. 10-20 
 

Table 10-8. 	 	 Dioxin-like PCB concentrations measured in sludges collected from 74 U.S.
 

publicly owned treatment works (POTWs) during 1994 ................. 10-21 
 

Table 10-9. 	 	 Dioxin-like PCB concentrations in sewage sludge collected from U.S. publicly
 

owned treatment works during 1999 ................................ 10-22 
 

Table 10-10. 	 Quantity of sewage sludge disposed of annually in 1989 by primary, secondary, or 

advanced treatment publicly owned treatment works (POTWs) and potential 

dioxin-like PCB TEQ releases ..................................... 10-23 
 

Table 10-11. 	 Quantity of sewage sludge disposed of annually in 1995 by primary, secondary, or 

advanced treatment publicly owned treatment works (POTWs) and potential 

dioxin-like PCB TEQ releases ..................................... 10-24 
 

Table 10-12. 	 Quantity of sewage sludge disposed of annually in 2000 by primary, secondary, or 

advanced treatment publicly owned treatment works (POTWs) and potential 

dioxin-like PCB TEQ releases ..................................... 10-25 
 

Table 10-13. 	 	PCB congener group emission factors for industrial wood combustorsa ..... 10-28 
 

Table 10-14. 	 	PCB congener group emission factors for medical waste incinerators 
 

(MWIs) ...................................................... 10-29 
 

Table 10-15. 	 	PCB congener group emission factors for a tire combustora .............. 10-30 
 
 


Table 10-16. 	 Dioxin-like PCB concentrations in cigarette tobacco in brands from various
 

countries (pg/pack) .............................................. 10-32 
 

Table 10-17. 	 Dioxin-like PCB concentrations in stack gas collected from a U.S. sewage sludge
 

incinerator .................................................... 10-33 
 

Table 10-18. 	 	Dioxin-like PCB emission factors from backyard barrel burning .......... 10-35 
 
 


Table 10-19. 	 PCB congener group emission factors for a petroleum catalytic reforming
 

unit .......................................................... 10-35 
 

xxiii 
LIST OF TABLES (continued) 

Table 10-20. Estimated tropospheric half-lives of dioxin-like PCBs with respect to gas-phase 

reaction with the OH radical ...................................... 10-40 
 

Table 10-21. Estimated PCB loads in the global environment as of 1985 .............. 10-42 
 

Table 10-22. Estimated domestic sales of aroclors and releases of PCBs, 1957–1974 (metric
 

tons) ......................................................... 10-43 
 

Table 10-23. Estimated U.S. usage of PCBs by use category, 1930–1975 .............. 10-44 
 

Table 10-24. Estimated direct releases of Aroclors to the U.S. environment, 1930–1974a
 

(metric tons) ................................................... 10-45 
 

Table 10-25. Estimated releases of dioxin-like PCB TEQs to the U.S. environment, 
 

1930–1977 .................................................... 10-46 
 

Table 11-1. Historical production, sales, and usage of 2,4-dichlorophenoxy-acetic acid 
 

(2,4-D) (metric tons) ............................................. 11-4 
 

Table 11-2. Historical production, sales, and usage of 2,4,5-trichlorophen-oxyacetic acid
 

(2,4,5-T) (metric tons) ............................................ 11-7 
 

Table 11-3. CDD/CDF concentrations (:g/kg) in recent sample of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) ...................................... 11-10 
 

Table 11-4. PCB 138 fluxes predicted by Harner et al. (1995) ...................... 11-16 
 

Table 11-5. Summary of flux calculations for total PCBs in Green Bay .............. 11-19 
 

Table 11-6. Comparison of model-estimated PCB concentrations with observed 
 

values ........................................................ 11-22 
 


xxiv 
LIST OF FIGURES
 
 


Figure 1-1. Chemical structure of 2,3,7,8-TCDD and related compounds ............... 1-7 
 

Figure 1-2. Estimated CDD/CDF I-TEQ emissions to air from combustion sources in the
 

United States for reference year 1987 ................................ 1-24 
 

Figure 1-3. Estimated CDD/CDF I-TEQ emissions to air from combustion sources in the
 

United States for reference year 1995 ................................ 1-25 
 

Figure 1-4. Estimated CDD/CDF I-TEQ emissions to air from combustion sources in the
 

United States for reference year 2000. ................................ 1-26 
 

Figure 1-5. Comparison of estimates of annual I-TEQ emissions to air (g I-TEQ/yr) for
 

reference years 1987, 1995, and 2000. ................................ 1-27 
 

Figure 1-6. Estimated CDD/CDF WHO-TEQ emissions to air from combustion sources in the
 

United States for reference year 1987 ................................ 1-28 
 

Figure 1-7. Estimated CDD/CDF WHO-TEQ emissions to air from combustion sources
 

in the United States for reference year 1995. ........................... 1-29 
 

Figure 1-8. Estimated CDD/CDF WHO-TEQ emissions to air from combustion sources
 

in the United States for reference year 2000. ........................... 1-30 
 

Figure 1-9. Comparison of estimates of annual WHO-TEQ emissions to air (g WHOTEQ/yr) for reference years 1987, 1995, and 2000. ...................... 1-31 
 

Figure 1-10. Congener profiles (as percent distributions to the sum of CDDs and CDFs) of
 

anthropogenic sources of chlorinated dibenzo-p-dioxins and chlorinated
 

dibenzofurans in the United States .................................. 1-45 
 

Figure 2-1. Typical mean distribution of CDD and CDF congeners in contemporary municipal
 

solid waste ...................................................... 2-3 
 

Figure 2-2. The de novo synthesis of CDDs/CDFs from heating carbon particulate at 300°C at
 

varying retention times ............................................ 2-15 
 

Figure 2-3. Temperature effects on CDD/CDF formation .......................... 2-16 
 

Figure 3-1. Typical mass burn waterwall municipal solid waste combustor ............. 3-3 
 

Figure 3-2. Typical mass burn rotary kiln combustor ............................... 3-4 
 

Figure 3-3. Typical modular starved-air combustor with transfer rams ................. 3-5 
 

xxv
 
LIST OF FIGURES (continued) 

Figure 3-4. Typical modular excess-air combustor ................................ 3-6 
 


Figure 3-5. Typical dedicated refuse-derived fuel-fired spreader stoker boiler ........... 3-7 
 


Figure 3-6. Fluidized-bed refuse-derived fuel furnace .............................. 3-8 
 


Figure 3-7. Municipal waste combustor design classes for 1987 .................... 3-11 
 


Figure 3-8. Municipal waste combustor design classes for 1995 .................... 3-12 
 


Figure 3-9. Municipal waste combustor design classes for 2000 .................... 3-13 
 


Figure 3-10. Congener and congener group profiles for air emissions from a mass burn
 
 


waterwall municipal waste combustor equipped with a dry scrubber and fabric
 

filter .......................................................... 3-32 
 

Figure 3-11. 2,3,7,8-TCDD frequency distribution (negative natural log concentration) . . . 3-33
 

Figure 3-12. 1,2,3,7,8-PeCDD frequency distribution (negative natural log concentration) . 3-34
 

Figure 3-13. Congener profile for air emissions from 17 hazardous waste incinerators tested
 

from 1993 through 1996. .......................................... 3-50 
 

Figure 3-14. Congener and congener group profiles for air emissions from
 

boilers and industrial furnaces burning hazardous waste .................. 3-52 
 

Figure 3-15. Congener and congener group profiles for air emissions from medical waste
 

incinerators without air pollution control devices (nondetects set equal 
 

to zero) ........................................................ 3-66 
 

Figure 3-16. Congener and congener group profiles for air emissions from medical waste
 

incinerators equipped with a wet scrubber and fabric filter ................ 3-67 
 

Figure 3-17. Congener and congener group profiles for air emissions from the Camellia
 

Memorial Lawn crematorium and Woodlawn Cemetery crematorium ....... 3-75 
 

Figure 3-18. Congener profile for air emissions from the University of Georgia animal
 

crematorium .................................................... 3-78 
 

Figure 3-19. Congener and congener group profiles for air emissions from
 

sewage sludge incinerators ......................................... 3-81 
 


xxvi 
LIST OF FIGURES (continued) 

Figure 3-20. 	 Congener and congener group profiles for air emissions from a tire 
 

combustor ...................................................... 3-85 
 

Figure 4-1. 	 	 Congener and congener group profiles for air emissions from diesel-fueled 
 

vehicles ....................................................... 4-15 
 

Figure 4-2. 	 	 Congener and congener group profiles for air emissions from leaded 
 

gas-fueled vehicles ............................................... 4-16 
 

Figure 4-3. 	 	 Congener and congener group profiles for air emissions from unleaded 
 

gas-fueled vehicles ............................................... 4-17 
 

Figure 4-4. 	 	 Tunnel air concentrations .......................................... 4-29 
 


Figure 4-5a. 	 Congener and congener group profiles for air emissions from industrial wood
 

combustors ..................................................... 4-46 
 

Figure 4-5b. 	 Congener and congener group profiles for air emissions from bleached Kraft mill
 

bark combustors ................................................. 4-47 
 

Figure 4-6. 	 	 Congener group profile for air emissions from residential oil-fueled 
 

furnaces ....................................................... 4-63 
 

Figure 4-7. 	 	 Congener and congener group profiles for air emissions from industrial 
 

oil-fueled boilers ................................................ 4-66 
 

Figure 4-8. 	 	 Congener and congener group profiles for air emissions from 
 

industrial/utility coal-fueled combustors (nondetects set equal to zero) ...... 4-73 
 

Figure 4-9. 	 	 Congener group profile for air emissions from residential coal-fueled 
 

combustors ..................................................... 4-77 
 

Figure 5-1. 	 	 Congener profile for air emissions from cement kilns burning hazardous waste
 

for reference years 1987 and 1995 (nondetect set equal to zero) ............. 5-9 
 

Figure 5-2. 	 	 Congener profile for air emissions from cement kilns burning hazardous
 

waste for reference year 2000. ...................................... 5-11 
 

Figure 5-3. 	 	 Congener profile for air emissions from cement kilns burning nonhazardous
 

waste for reference years 1987, 1995, and 2000. ........................ 5-13 
 

Figure 5-4. 	 	 Congener and congener group profiles for air emissions from petroleum
 

catalytic reforming units .......................................... 5-30 
 

xxvii 
LIST OF FIGURES (continued) 

Figure 5-5. CDD profiles for Japanese cigarettes, smoke, and ash ................... 5-33 
 

Figure 5-6. Congener group profiles for mainstream and sidestream cigarette smoke .... 5-35 
 

Figure 5-8 Congener and congener group profiles for air emissions from
 

Kraft black liquor recovery boilers (nondetect set equal to zero) ........... 5-45 
 
 


Figure 5-7. Congener group profiles for cigarette tobacco from various countries ....... 5-38 
 
 


Figure 6-1. Congener profile for landfill flare air emissions ......................... 6-3 
 
 


Figure 6-2. Congener profile for forest fire simulation approach emissions ............ 6-17 
 
 


Figure 7-1. Congener and congener group profiles for air emissions from secondary
 
 


aluminum smelters ............................................... 7-13 
 

Figure 7-2a. Congener group profile for air emissions from a secondary copper smelter . . . 7-18
 

Figure 7-2b. Congener and congener group profiles for a closed secondary copper 
 

smelter ........................................................ 7-19 
 

Figure 7-3. Congener and congener group profiles for air emissions from
 

secondary lead smelters ........................................... 7-26 
 

Figure 7-4. Congener profiles for air emissions from U.S. iron ore sintering plants ...... 7-31 
 

Figure 7-5. Congener group profile for air emissions from a scrap wire incinerator ...... 7-43 
 

Figure 7-6. Congener group profile for air emissions from a drum incinerator .......... 7-47 
 

Figure 8-1. 104 Mill Study full congener analysis results for pulp (nondetects equal 
 

to zero) ......................................................... 8-3 
 

Figure 8-2. 104 Mill Study full congener analysis results for sludge ................... 8-3 
 

Figure 8-3. 104 Mill Study full congener analysis results for effluent .................. 8-4 
 

Figure 8-4. Congener profile for technical-grade PCP (developed from data in last column in
 

Table 8-10) ..................................................... 8-57 
 

Figure 8-5. Congener profile for 2,4-D (salts and esters) (based on mean concentrations
 

reported in Table 8-34) ........................................... 8-73 
 


xxviii 
LIST OF FIGURES (continued) 

Figure 8-6. Congener profiles for sewage sludge (created from data in Table 8-38) ...... 8-88 
 


Figure 11-1. Fluxes among environmental reservoirs .............................. 11-2 
 


xxix 
LIST OF ABBREVIATIONS AND ACRONYMS
 

AHA American Hospital Association 

AMSA Association of Metropolitan Sewerage Agencies 

APCD Air pollution control device 

BDDs polybrominated dibenzo-p-dioxins 

BDFs polybrominated dibenzofurans 

Btu British thermal unit 

CaCl2 calcium chloride 

CARB California Air Resources Board 

CBI Confidential Business Information 

CDD polychlorinated dibenzo-p-dioxin 

CDF polychlorinated dibenzofuran 

CFR Code of Federal Regulations 

CSF Confidential Statement of Formula 

CKD cement kiln dust 

CO carbon monoxide 

CO2 carbon dioxide 

CuCl copper (I) chloride 

CuCl2 copper (II) chloride 

DBF dibenzofuran 

DCBz dichlorobenzene 

DCI data call-in 

DCP dichlorophenol 

DL detection limit 

dscm dry standard cubic meter 

DSI dry sorbent injection 

EDC ethylene dichloride 

EIA Energy Information Administration 

EPA U.S. Environmental Protection Agency 

EPRI Electric Power Research Institute 

ESP electrostatic precipitator 

FF fabric filter 

FCEM field chemical emissions measurement 

FeCl3 ferric (iron) chloride 

xxx 
LIST OF ABBREVIATIONS AND ACRONYMS (continued) 

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act 

GAC granular activated carbon 

GC/ECD gas chromatography/electron capture detector 

GC/MS gas chromatography/mass spectrometry 

HCl hydrogen chloride 

HCBz hexachlorobenzene 

HDD halogenated dibenzo-p-dioxin 

HDF halogenated dibenzofuran 

HWI hazardous waste incinerator 

HxCB hexachlorobiphenyl 

IUPAC International Union of Pure and Applied Chemistry 

KCl potassium chloride 

LOQ limit of quantitation 

MB-WW mass burn waterwall 

MCBz monochlorobenzene 

MgCl2 magnesium chloride 

MgO magnesium oxide 

MSW municipal solid waste 

MWI medical waste incinerator 

NaCl sodium chloride 

NaOCl soldium hypochlorite 

NCASI National Council of the Paper Industry for Air and Stream Improvement 

NiCl2 nickel chloride 

NiO nickel oxide 

Nm3 standard cubic meter 

NMOC nonmethane organic compound 

OAQPS Office of Air Quality Planning and Standards 

O2 molecular oxygen 

OH hydroxide ion 

OPP Office of Pesticide Programs 

ORD Office of Research and Development 

OSW Office of Solid Waste 

Pb lead 

xxxi 
LIST OF ABBREVIATIONS AND ACRONYMS (continued) 

PCA Portland Cement Association 

PCB polychlorinated biphenyl 

PCP pentachlorophenol 

PCP-Na pentachlorophenate 

PeCB pentachlorobiphenyl 

PeCBz pentachlorobenzene 

PM particulate matter 

POTW publicly owned treatment works 

ppb parts per billion 

ppm parts per million 

ppmv parts per million (volume basis) 

ppt parts per trillion 

PVC polyvinyl chloride 

QA/QC quality assurance/quality control 

RCRA Resource Conservation and Recovery Act 

RDF refuse-derived fuel 

SIC Standard Industrial Classification 

SIP State Implementation Plan 

SNUR Significant New Use Rule 

SO2 sulfur dioxide 

TCBz trichlorobenzene 

TCDD 2,3,7,8-tetrachlorobidenzo-p-dioxin 

TCDF 2,3,.7,8-tetrachlorobidenzofuran 

TeCB tetrachlorobiphenyl 

TeCP tetrachlorophenol 

TEF toxicity equivalency factor 

TEQ toxicity equivalence 

TEQ/yr toxicity equivalents per year 

TiCl4 titanium tetrachloride 

TrCB trichlorobiphenyl 

TrCP trichlorophenol 

TRI Toxics Release Inventory 

TSCA Toxic Substances Control Act 

xxxii 
LIST OF ABBREVIATIONS AND ACRONYMS (continued) 

2,4-D 2,4-dichlorophenoxyacetic acid 

2,4-DB 4-(2,4-dichlorophenoxy) butyric acid 

2,4-DCP 2,4-dichlorophenol 

2,4-DP 2-(2,4-dichlorophenoxy) propionic acid 

2,4,5-T 2,4,5-trichlorophenoxy (phenoxy herbicides) 

U.K. United Kingdom 

USDA U.S. Department of Agriculture 

VCM vinyl chloride monomer 

WHO World Health Organization 

WS wet scrubber 

xxxiii 
FOREWORD
 

The purpose of this document is to present an inventory of sources and environmental 

releases of dioxin-like compounds in the United States. This inventory is associated with three 

distinct reference years: 1987, 1995, and 2000. The presentation of information in this manner 

permits the ranking of sources by magnitude of annual release and allows for the evaluation of 

environmental trends over time. 

The term “dioxin-like” includes congeners of polychlorinated dibenzo-p-dioxins (CDDs), 

polychlorinated dibenzofurans (CDFs) having chlorine atoms in the 2,3,7,8 positions on the 

molecule, and certain coplanar-substituted polychlorinated biphenyls (PCBs). Dioxin-like refers 

to the fact that these compounds have similar chemical structure and physical-chemical properties 

and invoke a common toxic response. Because of their hydrophobic nature and resistance 

towards metabolism, these chemicals persist and bioaccumulate in fatty tissues of animals and 

humans. Consequently, the principal route of chronic population exposure is through the dietary 

consumption of animal fats, fish, shellfish, and dairy products. Dioxin-like compounds are 

persistent in soils and sediments, with environmental half-lives ranging from years to several 

decades. Understanding the sources and environmental releases of dioxin-like compounds is 

fundamental to ultimately linking sources with population exposures. It is through such 

understanding that actions can be taken to reduce human exposures. 

The quantitative results of the inventory are expressed in terms of the toxicity equivalence 

(TEQ) of the mixture of polychlorinated dibenzo-p-dioxin (CDD) and polychlorinated 

dibenzofuran (CDF) compounds present in environmental releases using a procedure sanctioned 

by the World Health Organization (WHO) in 1998. This TEQ procedure translates the complex 

mixture of CDDs and CDFs characteristic of environmental releases into an equivalent toxicity 

concentration of 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD), the most toxic member of 

this class of compounds. With this procedure, the quantity of the mixture of CDDs and CDFs 

present as a release is given the notation grams (g) TEQDF-WHO98. 

This inventory of sources and environmental releases of dioxin-like compounds concludes 

that, between 1987 and 2000, there was approximately 90% reduction in the release of dioxin-like 

compounds to the circulating environment of the United States from all known sources combined. 

Annual emission estimates (TEQDF-WHO98) of releases of CDDs/CDFs to air, water, and land 

from reasonably quantifiable sources are approximately 1,422 g in reference year 2000; 3,444 g in 

reference year 1995; and 13,965 g in reference year 1987. In 1987 and 1995, the leading sources 

of dioxin emissions to the U.S. environment were municipal waste combustors. The inventory 

also identifies bleached chlorine pulp and paper mills as a significant source of dioxin to the 

aquatic environment in 1987 but a minor source in 1995 and 2000. The inventory concludes that 

xxxiv 
the major source of dioxin in 2000 was the uncontrolled burning of refuse in backyard burn 

barrels in rural areas of the United States. 

The reduction in environmental releases of dioxin-like compounds from 1987 to 2000 is 

attributable to source-specific regulations, improvements in source technology, advancements in 

the pollution control technologies specific to controlling dioxin discharges and releases, and the 

voluntary actions of U.S. industries to reduce or prevent dioxin releases. 

Peter W. Preuss, Ph.D. 

Director 

National Center for Environmental Assessment 

Office of Research and Development 

xxxv 
PREFACE
 

This document, An Inventory of Sources and Environmental Releases of Dioxin-Like 

Compounds in the United States for the Years 1987, 1995, and 2000, was prepared by the 

National Center for Environmental Assessment, which is the health risk assessment program in 

the Office of Research and Development. The document presents estimates of annual releases of 

dioxin-like compounds specific for each year. It is a detailed compilation and description of all 

known U.S. sources and their associated activities that cause these compounds to be released into 

the open and circulating environment, i.e., to air, water, and land. The overall purpose of this 

report is to document and describe sources in the United States that release dioxin-like 

compounds into the open environment, quantify annual releases to the environment from known 

sources in a scientific and transparent manner, and provide a reliable basis for observing trends in 

environmental releases. To the extent practical, the inventory is a comprehensive analysis of 

dioxin sources. 

This final document reflects a consideration of all comments received on an External 

Review Draft dated March 2005 (EPA600/P-03/002A) provided by an expert panel at a peerreview workshop held September 13–15, 2005, and comments received during a 60-day public 

review and comment period (May 6–July 5, 2005). 

Over 800 references were reviewed and cited in the preparation of this document. The 

citations generally reflect publications up to and including the year 2003. 

xxxvi 
AUTHORS, CONTRIBUTORS, AND REVIEWERS
 

The National Center for Environmental Assessment (NCEA) was responsible for 

preparing this document. Major portions of this report were prepared by Versar, Inc. under EPA 

Contract No. 68-W-99-041. David Cleverly of NCEA served as the EPA Work Assignment 

Manager, providing overall direction and coordination of the project, as well as author of several 

chapters. 

Authors 

David Cleverly 

National Center for Environmental Assessment 

Office of Research and Development 

U.S. Environmental Protection Agency 

Washington, DC 

Greg Schweer 

Office of Pollution Prevention and Toxics 

U.S. Environmental Protection Agency 

Washington, DC 

Ronald Lee 

Versar, Inc. 

Springfield, VA 

Kelly McAloon 

Versar, Inc. 

Springfield, VA 

Charles Peck 

Versar, Inc. 

Springfield, VA 

Marit Espevik Randall 

Versar, Inc. 

Springfield, VA 

Karie Traub Riley 

Versar, Inc. 

Springfield, VA 

xxxvii 
AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) 

Contributors 

Frank Behan, Office of Solid Waste, Office of Solid Waste and Emergency Response, U.S. 

Environmental Protection Agency, Washington, DC 

Laurel Driver, Office of Air Quality Planning and Standards, Office of Air and Radiation, U.S. 

Environmental Protection Agency, Research Triangle Park, NC 

Marietta Echeverria, National Center for Environmental Assessment, Office of Research and 

Development, U.S. Environmental Protection Agency, Washington, DC 

Joseph Ferrario, Environmental Chemistry Laboratory, Office of Pesticide Programs, Office of 

Prevention, Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Stennis 

Space Center, MS 

Roy Huntley, Office of Air Quality Planning and Standards, Office of Air and Radiation, U.S. 

Environmental Protection Agency, Research Triangle Park, NC 

John Schaum, National Center for Environmental Assessment, Office of Research and 

Development, U.S. Environmental Protection Agency, Washington, DC 

Walt Stevenson, Office of Air Quality Planning and Standards, Office of Air and Radiation, U.S. 

Environmental Protection Agency, Research Triangle Park, NC 

Dwain Winters, Office of Pollution Prevention and Toxics, Office of Prevention, Pesticides and 

Toxic Substances, U.S. Environmental Protection Agency, Washington, DC 

Reviewers 

Internal EPA Review: Internal EPA review occurred from May 2 through June 5, 2003. The 

key reviewers included the following individuals and offices: 

Elmer Akin, EPA Region 4, Atlanta, GA 

Don Anderson, Office of Science and Technology, Office of Water, Washington, DC 

John D Bachmann, Office of Air Quality Planning and Standards, Research Triangle Park, NC 

Angela L. Bandemehr, Office of International Affairs, Washington, DC 

Frank Behan, Office of Solid Waste, Washington, DC 

xxxviii 
AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) 

Tom Braverman, Office of Air Quality Planning and Standards, Research Triangle Park, NC 

Laurel Driver, Office of Air Quality Planning and Standards, Research Triangle Park, NC 

Dale Evarts, Office of Air Quality Planning and Standards, Research Triangle Park, NC 

Brian Gullett, National Risk Management Laboratory, Research Triangle Park, NC 

Bob Holloway, Office of Solid Waste, Washington, DC 

Roy Huntley, Office of Air Quality Planning and Standards, Research Triangle Park, NC 

Phil Lorang, Office of Air Quality Planning and Standards, Research Triangle Park, NC 

Ed Ohanian, Office of Water, Washington, DC 

Debbie Sisco, Office of Pesticide Programs, Washington, DC 

Alan Rubin, Office of Water, Washington, DC 

John Schaum, National Center for Environmental Assessment, Washington, DC 

Laurie Schuda, National Center for Environmental Assessment, Washington, DC 

Joe Somers, Office of Transportation and Air Quality, Ann Arbor, MI 

Walt Stevenson, Office of Air Quality Planning and Standards, Research Triangle Park, NC 

Dwain Winters, Office of Pollution Prevention and Toxics, Washington, DC 

Donn Viviani, Office of Policy, Economics, and Innovation, Washington, DC 

External Peer Review: This document has undergone two external peer reviews: the first 

External Review Draft (EPA/600/P-98/002Aa) published in April 1998 was reviewed by an expert 

panel at a public meeting held in Alexandria, Virginia, on June 3–4, 1998. The most recent 

External Review Draft (EPA600/P-03/002A) published in March 2005 was released for public 

review and comment on May 6, 2005, and was subsequently reviewed at a public meeting held in 

Arlington, Virginia, on September 13–15, 2005. 

xxxix 
AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) 

1. Peer review meeting September 13-15, 2005 

Panel Chair 

Heidelore Fiedler, Ph.D., Scientific Affairs Officer, United Nations Environment Programme 

Chemicals, Chatelaine, Switzerland 

Panel Members 

Judith A. Abbott, Ph.D., Research Scientist, Bureau of Toxic Substance Assessment, New York 

State Department of Health, Albany, NY 

Gary A. Amendola, Ph.D., President, Amendola Engineering, Westlake, OH 

Patrick Dyke, President, PD Consulting, Oxford, United Kingdom 

Christopher C. Lutes, Ph.D., Principal Scientist/Associate Vice President, ARCADIS G&M, 

Inc., Durham, NC 

Jozef M. Pacyna, Ph.D., Center for Ecological Economics, Norwegian Institute for Air Research, 

Kjeller, Norway 

2. Peer review meeting June 3–4, 1998 

Panel Chair 

Valerie Thomas, Ph.D., Research Scientist, Center for Energy and Environmental Studies, 

Princeton University 

Panel Members 

Patrick Dyke, President, PD Consulting, Oxford, United Kingdom 

Raouf Morcos, Manager, National Office of Pollution Prevention, Environment Canada 

Quebec, Canada 

William Randall Seeker, Ph.D., Senior Vice President, Energy and Environmental Research, 

Irvine, CA 

Chun Yi Wu, Senior Engineer, Air Quality Division, Minnesota Pollution Control Agency, 

St. Paul, MN 

xl 
ACKNOWLEDGMENT
 

The authors would like to acknowledge the contributions of the following individuals: 

Terri Konoza of NCEA who managed the document production activities and provided editing, 

proofreading, and word processing support; Patricia von Brook of KBM Group who provided 

editing support; and Christine Chang of KBM Group who provided word processing support. 

xli 
EXECUTIVE SUMMARY 

This report, An Inventory of Sources and Environmental Releases of Dioxin-Like 

Compounds in the United States for the Years 1987, 1995, and 2000, presents estimates of annual 

releases of dioxin-like compounds specific to each year. It is a detailed compilation and 

description of all known U.S. sources and their associated activities that cause these compounds 

to be released into the open and circulating environment, i.e., to air, water, and land. The overall 

purpose of this report is to document and describe sources in the United States that release dioxinlike compounds into the open environment, quantify annual releases to the environment from 

known sources in a scientific and transparent manner, and provide a reliable basis for observing 

trends in environmental releases. To the extent practical, the inventory is a comprehensive 

analysis of dioxin sources. 

The term “dioxin-like” refers to chemical compounds that mimic the chemical and 

physical properties of dioxin and have similar toxic effects. These include compounds of 

chlorinated dibenzo-p-dioxins (CDDs), chlorinated dibenzofurans (CDFs), and certain coplanar 

polychlorinated biphenyls (PCBs). The dioxin-like CDDs and CDFs have chlorine atoms in the 

2,3,7,8 positions on the molecule. Dioxin-like PCBs contain zero or one chlorine atom in the 

2,2',6 or 6' positions. All together there are 7 CDDs, 10 CDFs, and 12 PCBs that are considered 

to be dioxin like. It should be emphasized that releases of dioxin-like compounds presented in 

this inventory are, for the most part, for dioxin-like CDDs and CDFs. Sources of dioxin-like 

PCBs are generally poorly characterized. 

Approach 

Only sources judged to have a reasonable likelihood for releases of dioxin-like compounds 

to the air, water, and land of the United States are addressed in this report. The release estimates 

were derived in one of two ways: (1) dioxin was measured as an actual release from the source 

(i.e., points of release from the source were sampled and evaluated), or (2) dioxin releases were 

calculated on the basis of an emission factor and activity level. The emission factor is the amount 

of dioxin anticipated to be emitted per unit of activity and is derived from measurements made at 

sources having similar characteristics. The activity level is the amount of material processed, 

produced, or consumed by the source in the course of a year or, in the case of mobile sources, the 

number of kilometers driven. It can take several forms, such as kilograms of material processed 

per year by an industrial facility, vehicle kilometers traveled per year by trucks and automobiles, 

and liters of wastewater discharged into surface water from industrial sources. The activity level 

is multiplied by the emission factor to arrive at an estimate of annual dioxin releases from those 

sources lacking direct measurement of dioxin emissions. 

xlii 
Confidence in the accuracy of both the emission factor and the activity level are rated as 

low, medium, or high, based on the quality of the data. All the release estimates from sources in 

the inventory are assigned an overall confidence rating based on the lowest rating assigned to 

either the emission factor or the activity level. In some cases, the data were not adequate to 

support even a low confidence rating. These cases were treated in one of two ways. If the data 

were sufficient to make an approximate, but clearly nonrepresentative, estimate of releases, the 

estimates were labeled as preliminary and were not included in the national inventory. If limited 

data suggested that dioxin releases were possible from a source but were not adequate to support 

even rudimentary calculations of emissions, the source was labeled as unquantifiable. This 

approach resulted in the classification scheme shown below. 

Category A High Confidence Included in the national 

Category B Medium Confidence quantitative inventory 

Category C Low Confidence 

Category D Preliminary Not included in the national 

Category E Unquantifiable quantitative inventory 

Throughout this document, environmental release estimates are presented in terms of 

toxicity equivalence (TEQs). TEQs are derived from a toxicity weighting system that converts all 

mixture components to a single value normalized to the toxicity of 2,3,7,8-TCDD. This is done 

for convenience in presenting summary information and to facilitate comparisons across sources. 

For many situations, however, it is important to use the individual CDD/CDF and PCB congener 

values rather than TEQs. CDD and CDF congener-specific releases for most sources are given in 

tables in each chapter. The summary amounts of dioxin-like compounds released to the 

environment are reported in units of grams (g) TEQ, developed by the World Health Organization 

(WHO), and are given the abbreviated notation of TEQDF-WHO98 throughout the document. 

The major findings of the inventory of sources and environmental releases of dioxin-like 

compounds in the United States are: 

1. 	 In 1987, 1995, and 2000, approximately 13,965; 3,444; and 1,422 g TEQ, respectively, 

were released into the U.S. environment from all sources. Figure ES-1 graphically 

displays these releases. 

xliii 
13,965.3 

3,444.2 

1,422.0 

14,000 

12,000 

10,000 

) QE

8,000 

T ( s am

6,000 

gr 

4,000 

2,000 

0 

Year 1987 Year 1995 Year 2000 

Figure ES-1. Total environmental releases of dioxin-like compounds (g TEQ)
 
 


from all quantifiable sources during 1987, 1995, and 2000.
 
 


2. 	 Environmental releases of dioxin-like compounds decreased by approximately 90% 

between 1987 and 2000. As shown in Figure ES-1, most of the reductions (75%) 

occurred between 1987 and 1995. The overall reduction in releases of dioxin-like 

compounds is attributed to the control of air emissions of these compounds from 

municipal waste combustors (MWCs), medical waste incinerators, and cement kilns 

burning hazardous waste and of wastewater discharges of the compounds into surface 

waters from pulp and paper mills using chlorine. These reductions were achieved 

through a combination of regulatory activities, improved emission controls, voluntary 

actions on the part of industry, and the closing of a number of facilities. Table ES-1 

shows the reductions made by the largest sources of dioxin-like compound releases. 

Emission estimates for individual sources that could be quantified, i.e., Categories A, 

B, and C, are presented in the main text of this report. 

3. 	 The leading source of dioxin-like compounds in 2000 was the backyard burning of 

refuse in barrels (498.5 g TEQ, or 35% of total releases), as shown in Table ES-2, 

which presents the top 10 sources of releases for 2000, 1995, and 1987. Backyard 

barrel burning of refuse is an activity that occurs in rural areas of the United States. It 

is unregulated on a national level, but many states have banned or limited the practice 

(New Jersey, New York, California, Minnesota, Wisconsin, Massachusetts, and 

xliv 
Table ES-1. Reductions of releases of dioxin-like compounds to the 

environment in reference years 2000 and 1987 from major sources in the 

United States 

Source category Releases to:

 2000 

(g TEQ) 

1987 

(g TEQ) 

Percent 

reduction 

Municipal waste 

combustion Air 83.8 8,905.1 >99 

Medical waste 

incineration Air 378.0 2,570.0 85 

Cement kilns 

burning hazardous 

waste Air 18.8 117.8 84 

Bleached chemical 

wood pulp and 

paper mills 

Surface 

water 1.0 356.0 >99 

Florida to name a few). In 1995 and 1987 MWCs were the leading source of releases 

(1,393.5 g, or 40% of total releases in 1995; 8,905.1 g, or 64% of releases in 1987). 

However, due to strict regulatory requirements limiting dioxin emissions, MWCs were 

ranked fourth among the top 10 sources in 2000, with emissions of only 83.8 g, or 6% 

of total releases. Automobiles burning leaded gasoline were ranked as the eighth 

leading source of dioxin in 1987. The phase out of lead in gasoline eliminated this 

source by 2000. Cement kilns burning hazardous waste dropped out of the top 10 

sources in 2000, due primarily to voluntary actions of industry combined with national 

regulatory requirements to reduce dioxin emissions. 

4. 	 Environmental releases of dioxin-like compounds in the United States occur from a 

wide variety of sources but are dominated by releases to the air from combustion 

sources. Figure ES-2 presents the breakdown of releases to air, water, and land for 

each reference year. 

5. 	 There are potential sources of dioxin-like compounds that were not included in the 

inventory. Significant amounts of the dioxin-like compounds produced annually in the 

United States are not considered releases to the open and circulating environment and, 

therefore, are not included in the national inventory. Examples include dioxin-like 

compounds generated internal to a process but destroyed before release and waste 

streams that are disposed of in approved and secure landfills. There are also potential 

sources for which no information exists to permit any reliable estimates of 

xlv 
Table ES-2. Top 10 sources of dioxin-like compound releases and amounts released 

(g TEQ) for reference years 2000, 1995, and 1987 

2000 1995 1987 

Rank (1,422 g total) (3,444 g total) (13,965 g total) 

1 Backyard barrel burning 

of refuse (air) 

498.5 Municipal waste 

combustion 

(incineration of refuse) 

(air) 

1,393.5 Municipal waste 

combustion 

(incineration of 

refuse) (air) 

8,905.1 

2 Medical waste/ 

pathological incineration 

(air) 

378.0 Backyard barrel 

burning of refuse (air) 

628.0 Medical 

waste/pathological 

incineration (air) 

2,570.0 

3 Municipal wastewater 

treatment sludge (applied 

to land and incinerated) 

(land and air) 

89.7 Medical 

waste/pathological 

incineration (air) 

487.0 Secondary copper 

smelting (air) 

983.0 

4 Municipal waste 

combustion (incineration 

of refuse) (air) 

83.8 Secondary copper 

smelting (air) 

271.0 Backyard barrel 

burning of refuse 

(air) 

604.0 

5 Coal-fired utility boilers 

(electric generating 

plants) (air) 

69.5 Cement kilns 

(hazardous waste 

burning) (air) 

156.1 Bleached chemical 

wood pulp and 

paper mills (land, 

water) 

370.1 

6 Diesel heavy-duty trucks 

(air) 

65.4 Municipal wastewater 

treatment sludge 

(applied to land and 

incinerated) (land and 

air) 

133.3 Cement kilns 

(hazardous waste 

burning) (air) 

117.8 

7 Industrial wood 

combustion (air) 

41.5 Coal-fired utility boilers 

(electric generating 

plants) (air) 

60.1 Municipal 

wastewater 

treatment sludge 

(applied to land 

and incinerated) 

(land and air) 

85.0 

8 Diesel off-road 

equipment (includes 

ships, farm equipment, 

trains) (air) 

33.1 Ethylene 

dichloride/vinyl 

chloride production 

(land, air, water) 

35.7 Coal-fired utility 

boilers (electric 

generating plants) 

(air) 

50.9 

9 Ethylene dichloride/vinyl 

chloride production 

(land, air, water)

30.0 Diesel heavy-duty 

trucks (air) 

33.3 Automobiles using 

leaded gasoline 

(air) 

37.5 

10 Sintering plants (air) 27.6 Bleached chemical 

wood pulp and paper 

mills (land and water) 

30.0 2,4­

Dichlorophenoxy 

acetic acid (2,4-D) 

(land) 

33.4 

xlvi 
Year 2000	 Year 1995 

1.8% 5.7% 

92.4% 

7.4% 2.6% 

90.0% 

Year 1987 

2.5% 0.9% 

96.5% 

Releases to Air Releases to Water Releases to Land 

Figure ES-2. Releases of dioxin-like compounds to air, water, and land in 
 
 

2000, 1995, and 1987.
 
 


environmental releases; therefore, these potential sources could not be included in the 

inventory. EPA has classified these potential sources as Category D sources. 

Examples include forest and grassland fires and accidental fires at municipal solid 

waste landfills. Taken together, these sources have the potential to significantly 

increase the emission estimates in the present inventory. 

6. 	 The amount of dioxin-like PCBs released from man-made sources remains poorly 

characterized. Only a total of 19.5, 78.5, and 51.5 g of PCB TEQ could be quantified 

for 2000, 1995, and 1987, respectively. To date, only sewage sludge has been 

adequately characterized as a source of dioxin-like PCB releases. 

xlvii 
1. BACKGROUND, APPROACH, AND CONCLUSIONS
 

1.1. BACKGROUND
 

This report presents a comprehensive inventory of sources of releases of dioxin-like 

compounds in the United States for the years 1987, 1995, and 2000. It is a detailed review and 

description of all known sources and their associated activities that cause these compounds to be 

released into the “open and circulating environment,” i.e., air, water, and land. 

The aim of this report is to: 

•	 Document and describe sources that release dioxin-like compounds into the 

circulating environment of the United States. 

•	 Quantify annual releases to the environment of the United States from known sources 

in a scientific and transparent manner. 

•	 Provide a reliable basis for time-trends analyses such as observing changes in total 

releases to the circulating environment from 1987 to 2000. Time-trend analyses 

provide a quantitative indication of the achievements made (or lack thereof) in 

reducing environmental releases of dioxin-like compounds from specific sources in 

the United States. 

This is the second dioxin source inventory issued by the U.S. Environmental Protection 

Agency (EPA, or the Agency). The first one was issued in draft form and covered the years 1987 

and 1995 (U.S. EPA, 1998a). The current effort updates this earlier document and adds annual 

release estimates for 2000. 

The intended audience and users of the dioxin inventory include: 

•	 Members of the general public who are interested in learning more about sources of 

emissions of dioxin-like compounds to the U.S. environment and in obtaining peerreviewed estimates of releases. 

•	 State and local regulatory agencies that are interested in obtaining reliable and peerreviewed information on sources and environmental releases of dioxin-like 

compounds. 

•	 EPA Regional and Program Offices that are responsible for evaluating the need for 

regulating and/or preventing dioxin releases to the environment. 

•	 Risk assessors in the private and public sectors who need reliable information on 

sources and releases of dioxin-like compounds to improve quantitative risk 

assessments of dioxin sources. 

1-1
 
•	 Researchers who are interested in documented and time-specific dioxin source and 

emissions data to be used in sequential time-trends analyses. 

•	 Private and public stakeholder groups that are interested in obtaining reliable and 

peer-reviewed information on dioxin sources and releases and in observing time 

trends in environmental releases of dioxin-like compounds from specific source 

categories. 

A complete listing of the nomenclature used in this report is depicted in Table 1-1. 

Table 1-1. Nomenclature for dioxin-like compounds 

Term/symbol Definition

 CDD Chlorinated dibenzo-p-dioxin, halogens substituted in any position

 CDF Chlorinated dibenzofuran, halogens substituted in any position

 PCB Polychlorinated biphenyl

 M Symbol for mono (i.e., one halogen substitution)

 D Symbol for di (i.e., two halogen substitution)

 Tr Symbol for tri (i.e., three halogen substitution)

 T Symbol for tetra (i.e., four halogen substitution)

 Pe Symbol for penta (i.e., five halogen substitution)

 Hx Symbol for hexa (i.e., six halogen substitution)

 Hp Symbol for hepta (i.e., seven halogen substitution)

 O Symbol for octa (i.e., eight halogen substitution)

 2,3,7,8 Halogen substitutions in the 2,3,7,8 positions 

Congener Any one particular member of the same chemical family (e.g., there are 75 

congeners of CDDs). 

Congener group Group of structurally related chemicals that have the same degree of chlorination 

(e.g., there are eight congener groups of CDDs, monochlorinated [MCDD] 

through octachlorinated [OCDD]). 

Isomer Substances that belong to the same congener group (e.g., 22 isomers constitute 

the congener group of TCDDs). 

Specific isomer Denoted by unique chemical notation (e.g., 2,4,8,9-tetrachlorodibenzofuran is 

referred to as 2,4,8,9-TCDF). 

Source: Adapted from U.S. EPA (1989a). 

1-2
 

1.1.1. Reference Years 

A central part of EPA’s dioxin inventory is the organization of estimates of annual 

releases of dioxin-like compounds into reference years 1987, 1995, and 2000. The selection and 

use of three reference years provides a basis for comparing environmental releases over time. 

The year 1987 was selected as the initial reference year because it was the earliest time 

when it was feasible to assemble a reasonably comprehensive inventory. Prior to that time, very 

little data existed on dioxin emissions from stacks or other release points. The first study 

providing the type of data needed for a national inventory was EPA’s National Dioxin Study 

(U.S. EPA, 1987a). The year 1987 also corresponds roughly with the time when significant 

advances occurred in emissions measurement techniques and in the development of highresolution mass spectrometry and gas chromatography, which allowed analytical laboratories to 

detect low levels of polychlorinated dibenzo-p-dioxin (CDD) and polychlorinated dibenzofuran 

(CDF) congeners in environmental samples. Soon after this time, a number of facilities began 

upgrades specifically intended to reduce CDD/CDF emissions. Consequently, 1987 emissions 

are representative of levels of emissions that occurred before the widespread installation of 

pollution control systems and pollution prevention techniques specifically designed to reduce 

dioxin releases from man-made sources into the air, land, and water. 

EPA selected 1995 as the second reference year because it reflects the completion time of 

the first set of regulatory activities specifically tailored to reduce dioxin releases from major 

sources. By 1995, EPA had proposed or promulgated regulations limiting CDD/CDF emissions 

from municipal waste combustors (MWCs), medical waste incinerators (MWIs), hazardous 

waste incinerators, cement kilns burning hazardous waste, and pulp and paper mill facilities 

using bleached chlorine processes. 

The year 2000 was chosen as the most current date that could be addressed when this 

effort began in 2002. Also, it corresponds to a reasonable time interval since 1995 when one 

could expect to see further changes occurring in releases as a result of continuing regulatory 

activities, voluntary actions on the part of industry, and facility closures. 

1.1.2. Regulatory Summary 

Tables 1-2 through 1-7 present a synopsis of EPA emission standards for the control of 

dioxin releases. As discussed in Section 1.3.2, these regulations (along with other factors) 

contributed to the reductions in dioxin emissions observed over time. 

1-3
 

Table 1-2. Municipal waste combustorsa 

Categoryb 

Stack emission limitc 

(ng total 

CDD/CDF/dscm) Effective date 

New large 13 September 20, 1994d 

June 19, 1996e 

Existing large 

With electrostatic precipitators as the APCD 

With dry scrubber/fabric filters as the APCD 

60 

30 

When SIPs are 

approvedf 

New small 13 June 6, 2001g 

Existing small 

With electrostatic precipitators as the APCD 

With dry scrubber/fabric filters as the APCD 

60 

30 

When SIPs are 

approvedh 

a

Air emission standards promulgated December 19, 1995.
 
 b

Large = aggregate capacity $225 tons/day; small = aggregate capacity <225 tons/day.
 
 c

ng total CDD/CDF/dscm = nanogram total Cl4 - Cl8 CDDs plus CDFs per dry standard cubic meter of stack gas
 


volume, corrected to 7% O2.
 
 d

Began construction on this date. 
 e

Modified or upgraded on this date.
 
 f

When SIPs have been approved by EPA (approx. 3 yr from the final rule or 1998).
 
 g

For facilities constructed on or before this date.
 
 h

When SIPs have been approved by EPA (approx. 3 yr from the final rule or 2003).
 


APCD = Air pollution control device
 


SIP = State Implementation Plan
 


1.1.3. Definition of Dioxin-Like Compound 

This inventory of sources and environmental releases addresses specific compounds in 

the following chemical classes: CDDs, CDFs, and polychlorinated biphenyls (PCBs). These 

subsets of chemicals are defined as “dioxin like.” Dioxin-like refers to the fact that these 

compounds have similar chemical structures and physical-chemical properties, and they invoke a 

common battery of toxic responses. Because of their hydrophobic nature and resistance towards 

metabolism, these chemicals persist and bioaccumulate in the fatty tissues of animals and 

humans. The CDDs include 75 individual compounds; CDFs include 135 compounds. These 

individual compounds are technically referred to as congeners. Only 7 of the 75 congeners of 

CDDs, or of brominated dibenzo-p-dioxins (BDDs), are thought to have dioxin-like toxicity; they 

are the ones with chlorine substitutions in—at a minimum—the 2, 3, 7, and 8 positions. Only 10 

of the 135 possible congeners of CDFs are thought to have dioxin-like toxicity; they also are the 

ones with substitutions in the 2, 3, 7, and 8 positions. This suggests that 17 individual 

CDDs/CDFs exhibit dioxin-like toxicity. 

1-4
 
Table 1-3. Hazardous waste incinerators and cement kilns and lightweight 

aggregate kilns burning hazardous wastea 

Source Standards for new facilitiesb Standards for existing facilitiesb 

Hazardous waste 

incinerators 

0.11 ng I-TEQ/dscm for dry 

APCD and/or waste heat boiler 

sources 

0.20 ng I-TEQ/dscm for all other 

incinerators 

0.20 ng I-TEQ/dscm and 

temperature control <400°F at the 

APCD inlet 

0.40 ng I-TEQ/dscm when PM 

control device operated >400°F 

Cement kilns burning 

hazardous waste 

0.20 ng I-TEQ/dscm and 

temperature control <400°F at 

the APCD inlet 

0.40 ng I-TEQ/dscm when PM 

control device operated >400°F 

0.20 ng I-TEQ/dscm and 

temperature control <400°F at the 

APCD inlet 

0.40 ng I-TEQ/dscm when PM 

control device operated >400°F 

Lightweight aggregate kilns 

burning hazardous waste 

0.20 ng I-TEQ/dscm or rapid 

quench below 400°F at kiln exit 

0.20 ng I-TEQ/dscm or rapid 

quench below 400°F at kiln exit 

a

Air emission standards promulgated September 30, 1999, and December, 2005.
 b

ng I-TEQ/dscm = nanogram I-TEQ per dry standard cubic meter of stack gas volume, corrected to 7% O2.
 


APCD = Air pollution control device (dry = dry scrubber or fabric filter)
 


PM = Particulate matter
 


Table 1-4. Cement kilns not burning hazardous wastea 

Existing cement kilnsb New cement kilnsb 

0.20 ng I-TEQ/dscm and temperature control 0.20 ng I-TEQ/dscm and temperature control 

<400°F at the APCD inlet <400°F at the APCD inlet 

0.40 ng I-TEQ/dscm when PM control device 0.40 ng I-TEQ/dscm when PM control device 

operated >400°F operated >400°F a

Air emission standards promulgated June 14, 1999.
 
 b

ng I-TEQ/dscm = nanogram I-TEQ per dry standard cubic meter of stack gas volume, corrected to 7% O2.
 


APCD = Air pollution control device
 


PM = Particulate matter
 


There are 209 PCB congeners, of which only 12 are thought to have dioxin-like toxicity; 

those with four or more lateral chlorine atoms with one or no substitution in the ortho position. 

These compounds are sometimes referred to as coplanar, meaning that they can assume a flat 

configuration with rings aligned along the same plane. The physical/chemical properties of each 

congener vary according to the degree and position of chlorine substitution. 

1-5
 
Table 1-5. Secondary aluminum smeltersa 

Process Emission standard 

Sweat furnace 0.8 ng I-TEQ/dscm stack gas corrected to 7% O2 

Thermal chip dryer 2.50 μg I-TEQ per metric ton of scrap charged to the 

dryer 

Scrap dryer/delacquering kiln/decoating kiln 0.25 g I-TEQ per metric ton of scrap charged to the 

kiln 

Scrap dryer/delacquering kiln/decoating kiln 

equipped with an afterburner 

5.0 g I-TEQ per metric ton of scrap charged to the kiln 

a

Air emission standards promulgated March 23, 2000. 

Table 1-6. Medical waste incineratorsa 

Categoryb Standardc When built 

New 

Small 

Medium and large 

125 ng total CDD/CDF/dscm or 

2.3 ng I-TEQ/dscm 

25 ng total CDD/CDF/dscm or 

0.6 ng I-TEQ/dscm 

Constructed after June 20, 1996, 

or existing units that 

commenced modification after 

March 16, 1998. 

Existing (all sizes) 125 ng total CDD/CDF/dscm or 

2.3 ng I-TEQ/dscm 

Constructed on or before 

June 20, 1996; requires 

approval of SIPsd 

a

Air emission standards promulgated September 15, 1997. 
 b

Small = capacity <100 kg/hr; medium = capacity >100 to 227 kg/hr; large = capacity >227 kg/hr.
 
 c

ng/dscm = nanogram per dry standard cubic meter of stack gas volume, corrected to 7% O2.
 
 d

When SIPs have been approved by EPA (approx. 5 yr from the final rule or 2002).
 


Table 1-7. Pulp and paper millsa 

Pollutant Maximum 1-day wastewater discharge 

Tetrachlorodibenzo-p-dioxin <5 parts per quadrillion 

Tetrachlorodibenzofuran 31.9 picograms per liter a

Effluent standards promulgated November 14, 1997. 

1-6
 

Generally speaking, this document focuses on the 17 CDDs/CDFs and a few of the 

coplanar PCBs that are frequently encountered in source characterization or environmental 

samples. 

CDDs and CDFs are tricyclic aromatic compounds that have similar physical and 

chemical properties. Certain PCBs (the so-called coplanar or mono-ortho coplanar congeners) 

are also structurally and conformationally similar. The most widely studied of this general class 

of compounds is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). TCDD, often called simply 

“dioxin,” represents the reference compound for this class of compounds. The structures of 

TCDD and several related compounds are shown in Figure 1-1. Although sometimes confusing, 

the term “dioxin” is often also used to refer to the complex mixtures of TCDD and related 

compounds emitted from sources or found in the environment or in biological samples. It can 

also be used to refer to the total TCDD “equivalents” found in a sample. This concept of toxicity 

equivalence is discussed below. 

Cl Cl 

Cl Cl 

2,3,7,8-Tetrachlorodibenzo-p-dioxin 

O 

O 

Cl Cl 

Cl Cl 

2,3,7,8-Tetrachlorodibenzofuran 

O 

Cl Cl 

Cl Cl 

2,3,4,7,8-Pentachlorodibenzofuran 

O 

Cl 

Cl Cl 

Cl Cl 

O 

O 

1,2,3,7,8-Pentachlorodibenzo-p-dioxin 

Cl 

Cl Cl 

Cl Cl 

Cl Cl 

3,3',4,4',5,5'-Hexachlorobiphenyl 

Cl Cl 

Cl Cl 

Cl 

3,3',4,4',5-Pentachlorobiphenyl 

Figure 1-1. Chemical structure of 2,3,7,8-TCDD and related compounds. 

1-7
 

1.1.4. Toxicity Equivalence Factors 

CDDs, CDFs, and PCBs are commonly found as complex mixtures when detected in 

environmental media and biological tissues or when measured as environmental releases from 

specific sources. Humans are likely to be exposed to mixtures of CDDs, CDFs, and dioxin-like 

PCB congeners that vary by source and pathway, complicating the assessment of human health 

risk assessment. In order to address this problem, the concept of a “toxicity equivalence” (TEQ) 

has been considered and discussed by the scientific community, and toxicity equivalence factors 

(TEFs) have been developed and introduced to facilitate risk assessment of exposure to these 

chemical mixtures. 

On the most basic level, TEFs compare the potential toxicity of each dioxin-like 

compound in the mixture to the well-studied and well-understood toxicity of TCDD, the most 

toxic member of the group. The comparison procedure involves assigning individual TEFs to the 

2,3,7,8-substituted CDD/CDF congeners and dioxin-like PCBs. To accomplish this, scientists 

have reviewed the toxicological databases and, with considerations of chemical structure, 

persistence, and resistance to metabolism, have agreed to ascribe specific “order of magnitude” 

TEFs for each dioxin-like congener relative to TCDD, which is assigned a TEF of 1. The other 

congeners have TEF values ranging from 1 to 0.00001. 

Thus, these TEFs are the result of scientific judgment of a panel of experts using all of the 

available data and are selected to account for uncertainties in the available data and to avoid 

underestimating risk. In this sense, they can be described as “public health-conservative” values. 

To apply this TEF concept, the TEF of each congener present in a mixture is multiplied by the 

respective mass concentration, and the products are summed to represent the 2,3,7,8-TCDD TEQ 

of the mixture (eq 1-1). 

TEQ – 3i!n(Congeneri × TEFi) % (Congenerj × TEFj) % ......(Congenern × TEFn) (1-1) 

The TEF values for CDDs and CDFs were originally adopted by international convention 

(U.S. EPA, 1989a). These values were further reviewed and/or revised, and TEFs were also 

developed for PCBs (Ahlborg et al., 1994; Van den Berg et al., 1998). A problem arises in that 

past and present quantitative exposure and risk assessments may not have clearly identified 

which of three TEF schemes was used to estimate the TEQ. This document uses a new uniform 

TEQ nomenclature that clearly distinguishes between the different TEF schemes and identifies 

the congener groups included in specific TEQ calculations. The nomenclature uses the following 

abbreviations to designate which TEF scheme was used in the TEQ calculation: 

1-8
 

•	 I-TEQ refers to the international TEF scheme adopted by EPA in 1989 (U.S. EPA, 

1989a). See Table 1-8. 

•	 TEQ-WHO94 refers to the 1994 World Health Organization (WHO) extension of the 

I-TEF scheme to include 13 dioxin-like PCBs (Ahlborg et al., 1994). See Table 1-9. 

•	 TEQ-WHO98 refers to the 1998 WHO update to the previously established TEFs for 

dioxins, furans, and dioxin-like PCBs (Van den Berg et al., 1998). See Table 1-10. 

Table 1-8. The TEF scheme for I-TEQDF 

Dioxin congener TEF Furan congener TEF 

2,3,7,8-TCDD 1 2,3,7,8-TCDF 0.1 

1,2,3,7,8-PeCDD 0.5 1,2,3,7,8-PeCDF 0.05 

1,2,3,4,7,8-HxCDD 0.1 2,3,4,7,8-PeCDF 0.5 

1,2,3,6,7,8-HxCDD 0.1 1,2,3,4,7,8-HxCDF 0.1 

1,2,3,7,8,9-HxCDD 0.1 1,2,3,6,7,8-HxCDF 0.1 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.01 

0.001 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

0.1 

0.1 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.01 

0.01 

0.001 

Table 1-9. The TEF scheme for dioxin-like PCBs, as determined by the 

World Health Organization in 1994 

Chemical structure IUPAC number TEF 

3,3',4,4'-TCB 

2,3,3',4,4'-PeCB 

2,3,4,4',5-PeCB 

2,3',4,4',5-PeCB 

2',3,4,4',5-PeCB 

3,3',4,4',5-PeCB 

2,3,3',4,4',5-HxCB 

2,3',4,4',5,5'-HxCB 

3,3',4,4',5,5'-HxCB 

3,3',4,4',5,5'-HxCB 

2,2',3,3',4,4',5-HpCB 

2,2',3,4,4',5,5'-HpCB 

2,3,3',4,4',5,5'-HpCB 

PCB-77 

PCB-105 

PCB-114 

PCB-118 

PCB-123 

PCB-126 

PCB-156 

PCB-157 

PCB-167 

PCB-169 

PCB-170 

PCB-180 

PCB-189 

0.0005 

0.0001 

0.0005 

0.0001 

0.0001 

0.1 

0.0005 

0.0005 

0.00001 

0.01 

0.0001 

0.00001 

0.0001 

1-9
 

Table 1-10. The TEF scheme for TEQDFP-WHO98 

Dioxin congener TEF Furan congener TEF 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

1 

1 

0.1 

0.1 

0.1 

0.01 

0.0001 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.1 

0.05 

0.5 

0.1 

0.1 

0.1 

0.1 

0.01 

0.01 

0.0001 

Chemical structure IUPAC number TEF 

3,3',4,4'-TCB 

3,4,4',5-TCB 

2,3,3',4,4'-PeCB 

2,3,4,4',5-PeCB 

2,3',4,4',5-PeCB 

2',3,4,4',5-PeCB 

3,3',4,4',5-PeCB 

2,3,3',4,4',5-HxCB 

2,3,3',4,4',5'-HxCB 

2,3',4,4',5,5'-HxCB 

3,3',4,4',5,5'-HxCB 

2,3,3',4,4',5,5'-HpCB 

PCB-77 

PCB-81 

PCB-105 

PCB-114 

PCB-118 

PCB-123 

PCB-126 

PCB-156 

PCB-157 

PCB-167 

PCB-169 

PCB-189 

0.0001 

0.0001 

0.0001 

0.0005 

0.0001 

0.0001 

0.1 

0.0005 

0.0005 

0.00001 

0.01 

0.0001 

The nomenclature also uses subscripts to indicate which family of compounds is included 

in any specific TEQ calculation. Under this convention, a subscript D is used to designate 

dioxins, a subscript F to designate furans, and a subscript P to designate PCBs. As an example, 

TEQDF-WHO98 would be used to describe a mixture for which only dioxin and furan congeners 

were determined and where the TEQ was calculated using the WHO98 scheme. If PCBs had also 

been determined, the nomenclature would be TEQDFP-WHO98. Note that the designations TEQDFWHO98 and I-TEQDF are interchangeable, as the TEFs for dioxins and furans are the same in each 

scheme. Note also that in this document I-TEQ sometimes appears without the D or F subscripts. 

This indicates that the TEQ calculation includes both dioxins and furans. This document 

emphasizes the WHO98 TEF scheme as the preferred scheme to be used to assign TEQs to 

complex environmental mixtures. 

1-10
 
Throughout this document, environmental release estimates are presented in terms of 

TEQs. This is done for convenience in presenting summary information and to facilitate 

comparisons across sources. For purposes of environmental fate modeling, however, it is 

important to use the individual CDD/CDF and PCB congener values rather than TEQs. This is 

because the physical/chemical properties of individual CDD/CDF congeners vary and, 

consequently, the congeners will behave differently in the environment. For example, the 

relative mix of congeners released from a stack cannot be assumed to remain constant during 

transport through the atmosphere and deposition to various media. The full congener-specific 

release rates for most sources are given in an electronic database that will become available as a 

companion to this document. 

1.1.5. Information Sources 

In general, the literature used to prepare this report includes documents published in 2003 

or earlier. Some 2004 documents are cited, primarily in Chapter 2, which covers formation 

theory, but a thorough literature review was not extended past 2003. 

EPA’s Toxics Release Inventory (TRI) began collecting data on PCBs in 1988 and on 

CDDs/CDFs in 2000 (U.S. EPA, 2003c). These data were considered in this report for purposes 

of identifying possible sources, but they were not used for making quantitative release estimates 

because of the following considerations: 

•	 With respect to PCBs, the TRI data are reported as total PCBs rather than on a 

congener-specific basis. Thus, it is unknown what portion of these releases are 

dioxin-like PCBs, and TEQs cannot be calculated. In their present format, the PCB 

TRI data are not readily usable within the structure of this dioxin inventory. 

•	 With respect to CDDs/CDFs, the reporting format under TRI is the sum quantity of 

the 17 toxic CDDs/CDFs that are emitted in a given year (i.e., the sum of the 2,3,7,8­

chlorine-substituted compounds). Neither the releases of the individual CDD/CDF 

congeners nor the TEQs must be reported; therefore, the dioxin TRI data are not 

readily usable within the structure of this dioxin inventory. 

•	 The accuracy of the TRI data is unknown because they are self-reported and are not 

required to be based on measurements. 

•	 The TRI reports lack specific details and descriptions of the reporting industries. This 

information is needed for the dioxin inventory because the calculation of sourcespecific emission factors (representative of industrial source categories) strongly 

depends on closely matching facilities in terms of similarity of process, production, 

and pollution control. 

1-11
 
•	 The TRI reporting format does not include information on the strengths/weaknesses 

of the data, and therefore, it would be difficult to evaluate these data in terms of the 

confidence rating scheme developed for this inventory (presented in Section 1.2.3). 

1.2. APPROACH 

Only sources judged to have a reasonable likelihood for releases to the circulating 

environment were addressed in this report. For example: 

•	 CDDs/CDFs and dioxin-like PCBs in air emissions and wastewater discharges are 

included, whereas those in intermediate products or internal wastestreams are not. 

For example, the CDDs/CDFs in a wastestream going to an on-site incinerator are not 

addressed in this report, but any CDDs/CDFs in the stack emissions from the 

incinerator are included. 

•	 CDDs/CDFs and dioxin-like PCBs in wastestreams applied to land in the form of 

“land farming” are included, whereas those disposed of in permitted landfills were 

excluded. Properly designed and operated landfills are considered to achieve longterm isolation from the circulating environment. Land farming, however, involves the 

application of wastes directly to land, clearly allowing for releases to the circulating 

environment. 

1.2.1. Source Classes 

The major identified sources of environmental releases of dioxin-like compounds in the 

United States are grouped into five broad categories. 

Combustion. CDDs/CDFs are formed in most combustion systems (which can include 

those that incinerate wastes such as municipal solid waste, sewage sludge, medical waste, and 

hazardous wastes); in other high-temperature sources (such as cement kilns); in poorly or 

uncontrolled combustion sources (such as forest fires, brush fires, landfill fires, accidental fires, 

building fires, and open burning of wastes); and during the burning of various fuels (such as coal, 

wood, and petroleum products). 

Metals smelting, refining, and processing. CDDs/CDFs can be formed during various 

types of primary and secondary metals operations, including iron ore sintering, lead smelting, 

copper smelting, magnesium and titanium dioxide production, steel production, and scrap metal 

recovery. 

Chemical manufacturing. CDDs/CDFs can be formed as by-products of the 

manufacture of chlorine-bleached wood pulp, chlorinated phenols (e.g., pentachlorophenol 

[PCP]), PCBs, chlorobenzenes, phenoxy herbicides (e.g., 2,4-D and 2,4,5-T), and chlorinated 

aliphatic compounds (e.g., ethylene dichloride, vinyl chloride, polyvinyl chloride). 

1-12
 
Natural sources and processes. The evidence for the widespread existence of natural 

sources of dioxin is quite weak. Recent studies suggest that CDDs/CDFs can form under certain 

environmental conditions (e.g., composting) from the action of microorganisms on chlorinated 

phenolic compounds. Similarly, CDDs/CDFs have been reported to form during photolysis of 

highly chlorinated phenols. Certain clays used in ceramics (e.g., ball clay) are believed to have 

become contaminated with dioxin as a result of natural processes, but the source of 

contamination remains unknown. Although it has been suggested that volcanos may be a natural 

source, there is no reliable evidence that volcanos produce and emit significant amounts of 

dioxin during eruptions. 

Reservoirs. Reservoirs are environmental compartments and materials that have the 

capacity to store previously formed CDDs/CDFs or dioxin-like PCBs. These compounds are 

thus sequestered from the open and circulating environment. Potential reservoirs include soils, 

sediments, and biota as well as some anthropogenic materials, such as PCP treated telephone 

poles. Dioxin-like compounds in these reservoirs have the potential for redistribution and 

circulation in the environment through the physical processes of leaching, volatilization, erosion, 

sedimentation, and deposition. Whenever dioxins are released from their place of storage back 

into the circulating environment, the reservoir is considered a source of dioxin. 

Sources can also be categorized in terms of when releases occur: (1) contemporary 

formation sources (sources that have essentially simultaneous formation and release) and (2) 

reservoir sources (materials or places that contain previously formed CDDs/CDFs or dioxin-like 

PCBs that are re-released to the environment). The contemporary formation sources are 

discussed in Chapters 2 through 10 and the reservoir sources are discussed in Chapter 11. 

Table 1-11 provides a comprehensive list of all known or suspected sources of 

CDDs/CDFs in the United States. The checkmarks indicate how each source was classified in 

terms of the following six categories: 

1.	 Contemporary formation sources with reasonably well-quantified releases (see 

Section 1.4.2). These sources are listed in Table 1-11 and release estimates are shown 

in Table 1-12. 

2.	 Contemporary formation sources without quantified release estimates. These sources 

are listed in Table 1-12. 

3. 	 Reservoir sources with reasonably well-quantified releases. These sources would 

have been listed in Table 1-11, but none have yet been identified. 

1-13
 

Table 1-11. Known and suspected sources of CDDs/CDFs 

1-14
 


Source category 

Contemporary formation sources Reservoi

Quantifiable 

(Categories 

A, B and C) 

Preliminary 

estimate 

(Category D) 

Not 

quantifiable 

(Category E) 

Quantifiable 

(Categories 

A, B and C) 

Prelim

estim

(Categ

COMBUSTION SOURCES Waste incineration 

Municipal waste combustion Hazardous waste incineration Boilers/industrial furnaces Medical waste/pathological incineration Crematoria Sewage sludge incineration Tire combustion Pulp and paper mill sludge incinerators Biogas combustion 

TTTTTTTT

T

Power/energy generation 

Vehicle fuel combustion - leadeda

 - unleaded - diesel 

Wood combustion - residential 

- industrial 

Coal combustion - residential 

- industrial/utility 

Oil combustion - residential 

- industrial/utility 

TTTTTTT

TT

Other high-temperature sources 

Cement kilns burning hazardous waste Cement kilns burning nonhazardous waste 

Asphalt mixing plants Petroleum refining catalyst regeneration Cigarette combustion Carbon reactivation furnaces Kraft recovery boilers Manufacture of ball clay products

 Glass manufacturing Lime kilns Rubber manufacturing 

TTTTTT

T

TTTT
Table 1-11. Known and suspected CDD/CDF sources (continued) 

1-15
 


Source category 

Contemporary formation sources Reservoi

Quantifiable 

(Categories 

A, B and C) 

Preliminary 

estimate 

(Category D) 

Not 

quantifiable 

(Category E) 

Quantifiable 

(Categories 

A, B and C) 

Prelim

estim

(Categ

Minimally controlled or uncontrolled combustion 

Combustion of landfill gas in flares Landfill fires Accidental fires, structural Accidental fires, vehicles Forest, brush, and straw fires Backyard barrel burning Uncontrolled combustion of PCBs

 Burning of candles 

T

TTTTT

TT

METAL SMELTING/REFINING Ferrous metal smelting/refining 

Sintering plants Coke production Electric arc furnaces Ferrous foundries 

T

TTT

Nonferrous metal smelting/refining 

Primary aluminum Primary copper Primary magnesium Primary nickel Secondary aluminum Secondary copper

 Secondary lead 

TTTT

T

TT

Scrap electric wire recovery TDrum and barrel reclamation T
Table 1-11. Known and suspected CDD/CDF sources (continued) 

1-16
 


Source category 

Contemporary formation sources Reservoi

Quantifiable 

(Categories 

A, B and C) 

Preliminary 

estimate 

(Category D) 

Not 

quantifiable 

(Category E) 

Quantifiable 

(Categories 

A, B and C) 

Prelim

estim

(Categ

CHEMICAL MANUFACTURING (releases to the environment) Bleached chemical wood pulp and paper mills Mono- through tetrachlorophenols Pentachlorophenol Chlorobenzenes Chlorobiphenyls (leaks/spills) Ethylene dichloride/vinyl chloride Dioxazine dyes and pigments 2,4-Dichlorophenoxy acetic acid Municipal wastewater treatment Tall oil-based liquid soaps 

TT

T

TTTTTTT

BIOLOGICAL AND PHOTOCHEMICAL PROCESSES 

T 

RESERVOIR SOURCES Land Air Water Sediments 

T

Anthropogenic structures 

PCP-treated wood 

aLeaded fuel production and the manufacture of motor vehicle engines requiring leaded fuel for highway use are prohibited in the Unite

Section 4.1, for details.) 
Table 1-12. Inventory of contemporary releases (g/yr) of dioxin-like compounds from known sourc

United States for reference years 2000, 1995, and 1987 and preliminary release estimates for 2000 

Source Category 

2000 Inventory 

TEQDF -

WHO98 I-TEQ 

Category 

ratinga

1995 Inventory 

TEQDF -

WHO98 I-TEQ 

Category 

ratinga

1987 Inve

TEQDF -

WHO98 I-TEQ 

RELEASES TO AIR 

WASTE INCINERATION 

Municipal waste combustion 

Hazardous waste incineration 

83.8 

3.2 

76.3 

3.2 

A 

B 

1,393.5 

5.8 

1,101.3 

5.7 

B 

B 

8,905.1 

5.0 

7,858.8 

5.0 

1-17
 

Boilers/industrial furnaces 

Halogen acid furnaces 

Medical waste/pathological 

incineration 

1.8 

0.3 

378 

1.8 

0.3 

357 

C 

C 

C 

0.4 

NA 

487 

0.4 

NA 

459 

C 

C 

0.8 

NA 

2,570 

0.

NA 

2,440 

Crematoria 

 - human 0.3 0.3 C 0.2 0.2 C 0.2 0.

 - animal 

Sewage sludge incineration Tire combustion 

9.6 

0.5 

9.4 

0.5 

B 

C 

14.2 

0.1 

14 

0.1 

B 

C 

5.8 

0.1 

5.

0.

Pulp and paper mill sludge 

incineratorsb 

Biogas combustion 

POWER/ENERGY GENERATION 

Vehicle fuel combustion 

- leaded gasolinec 

- unleaded gasoline on-road 

- unleaded gasoline off-road 

- diesel on-road (Trucks) - diesel off-road 

7 

0.4 

65.4 

6.7 

0.4 

61.7 

C 

C 

C

1.6 

4.7 

NA 

33.3 

1.3 

4.4 

NA 

31.5 

C C 

C 

37.5 

3.6 

NA 

27.8 

31.9 

3.3 

NA 

26.3 

 - equipment  - railroad 

22 

6.8 

21 

6.4 

C C 

12 

7 

11 

6.6 

C 

C 

9.4 

5.8 

8.8 

5.5 

 - commercial marine

 vessel 

4.3 4 C 4.8 4.5 C 3.8 3.6 
Table 1-12. Inventory of contemporary releases (g/yr) of dioxin-like compounds from known sourc

United States for reference years 2000, 1995, and 1987 and preliminary release estimates for 2000 (

2000 Inventory 1995 Inventory 1987 Inve

Source Category 

TEQDF - Category TEQDF - Category TEQDF - 

WHO98 I-TEQ ratinga WHO98 I-TEQ ratinga WHO98 I-TEQ 

Wood combustion 

- residential 11.3d 11.3d C 15.7d 15.7d C 22d 22d 

- industrial 41.5 39.4 C 26.2 24.9 C 26.5 25.2 

Coal combustion 

- utility boilers 69.5 70.4 B 60.1 60.9 B 50.9 51.4 

- residentiale 

- commercial/industrial 

Oil combustion 

1-18
 - industrial/utility, residual oil 1.7 1.5 C 10.7 9.3 C 17.8 15.5 

- industrial/utility, distillate oil 7.3 6.3 C 7.3 6.4 C 8.3 7.2 - institutional/commercial 0.7 0.6 C 0.8 0.7 C 1.5 1.3 

heating, residual oil 

- institutional/commercial 2.9 2.5 C 3.1 2.7 C 3.7 3.2 

heating, distillate oil 

- residential, distillate oil 4.5 3.6 C 5.0 3.9 C 5.4 4.2 

OTHER HIGH-TEMPERATURE SOURCES 

Cement kilns burning hazardous 18.8 16.6 B 156.1 145.3 C 117.8 109.6 

waste 

Lightweight aggregate kilns 1.9d 1.8d C 2.4d 2.4d C 3.3d 3.3

burning hazardous waste 

Cement kilns burning 17.2 16.6 C 16.6 15.9 C 12.7 12.3 

nonhazardous waste 

Asphalt mixing plants 

Petroleum refining catalyst 2.2 2.1 C 2.2 2.1 C 2.2 2.

regeneration 

Cigarette combustion 0.4 0.4 C 0.8 0.8 C 1 1 Carbon reactivation furnaces 0.1d 0.1d C 0.1d 0.1d C 0.1d 0.

Kraft recovery boilers 0.8 0.8 B 2.3 2.3 B 2 2 
Table 1-12. Inventory of contemporary releases (g/yr) of dioxin-like compounds from known sourc

United States for reference years 2000, 1995, and 1987 and preliminary release estimates for 2000 (

2000 Inventory 1995 Inventory 1987 Inve

Source Category 

TEQDF - Category TEQDF - Category TEQDF - 

WHO98 I-TEQ ratinga WHO98 I-TEQ ratinga WHO98 I-TEQ 

MINIMALLY CONTROLLED OR UNCONTROLLED COMBUSTIONf 

Combustion of landfill gasLandfill fires 

 Accidental fires - structural 

 - vehiclesForest and brush firesh 

Backyard barrel burningi 1-19 498.5 472.6 C 628 595 C 604 573 

Residential yard waste burningj L 

and clearing debris burning 

METALLURGICAL PROCESSES 

Ferrous metal smelting/

 refining 

- sintering plants 27.6 24.4 A 28 25.1 B 32.7 29.

- coke production 

- electric arc furnaces 

 - foundries 

Nonferrous metal smelting/

 refining 

- primary copper 0.3d 0.3d B <0.5d <0.5d B <0.5d <0.5

- secondary aluminum 8.3 7.8 C 19.5 18.3 C 10.9 10.2 - secondary copper 0.9 0.9 C 271 266 C 983 966 - secondary lead 2.5 2.4 B 1.7 1.6 B 1.3 1.2 - primary magnesium 4.3d 4.3d A 4.1d 4.1d C NA NA 

Drum and barrel reclamation 0.6 0.6 C 0.1 0.1 C 0.1 0.
Table 1-12. Inventory of contemporary releases (g/yr) of dioxin-like compounds from known sourc

United States for reference years 2000, 1995, and 1987 and preliminary release estimates for 2000 (

Source Category 

2000 Inventory 

TEQDF -

WHO98 I-TEQ 

Category 

ratinga

1995 Inventory 

TEQDF -

WHO98 I-TEQ 

Category 

ratinga

1987 Inve

TEQDF -

WHO98 I-TEQ 

CHEMICAL MANUFACTURE/PROCESSING SOURCES Ethylene dichloride/vinyl chloride/PVC 

5.5d 5.5 A 11.2d 11.2 A NA NA 

 Chor-alkali facilities 1.8d 1.8d A 1.8 d 1.8 d C NA NA 

TOTAL RELEASES TO AIRk 1,314.5 1,243.6 A, B, C 3,239.9 2,857.1 A, B, C 13,482.6 12,230

RELEASES TO WATER 

1-20
 

CHEMICAL MANUFACTURE/PROCESSING SOURCES 

Bleached chemical wood pulp

 and paper mills 

POTW (municipal) wastewater Ethylene dichloride/vinyl

 chloride/PVC 

1.0 

23.1d 

1.0 

23.9d 

A 

A 

28 

23.1d 

28 

23.9d 

A 

C 

356 

NA 

356 

NA 

 Chor-alkali facilities 

TOTAL RELEASES TO WATERk 

1.8d 

25.9 

1.8d 

26.7 

A 

A, B, C 

1.8d 

52.9 

1.8 

d 

53.7 

C 

A, B, C 

NA 

356 

NA 

356 

RELEASES TO LAND 

CHEMICAL MANUFACTURING/PROCESSING SOURCES 

Bleached chemical wood pulp

 and paper mill sludge 

Ethylene dichloride/vinyl

 chloride/PVC 

Municipal wastewater treatment

 sludge 

Commercially marketed sewage 

sludge 

0.1 1.4 

78.2 

1.9 

0.1 

1.5 

78.2 

1.9 

A 

A 

A 

A 

2 

1.4 

116.1 

3 

2 

1.5 

156.5 

4 

A 

B 

A 

A 

14.1 

NA 

76.6 

2.6 

14.

NA 

103 

3.5 
Table 1-12. Inventory of contemporary releases (g/yr) of dioxin-like compounds from known sourc

United States for reference years 2000, 1995, and 1987 and preliminary release estimates for 2000 (

1-21
 

Source Category 

2000 Inventory 

TEQDF - Category 

WHO98 I-TEQ ratinga

1995 Inventory 

TEQDF - Category 

WHO98 I-TEQ ratinga

1987 Inve

TEQDF - 

WHO98 I-TEQ 

NA NA 28.9 18.4 A 33.4 21.3 

2,4-Dichlorophenoxy acetic acid 

TOTAL RELEASES TO LANDk 81.6 81.7 A, B, C 151.4 182.4 A, B, C 126.7 141.9 

OVERALL RELEASES TO OPEN AND CIRCULATING 

1,422.0 1,352.0 A, B, C 3,444.2 3,093.2 A, B, C 13,965.3 12728.6 

ENVIRONMENT 

aThe most reliable estimates of environmental releases are those sources in categories A, B, and C. bIncluded in estimate for wood combustion, industrial. cLeaded fuel production and the manufacture of motor vehicle engines requiring leaded fuel for highway use are prohibited in the Unite

 Section 4.1 for details.) 

dEstimate based on a TEQDF-WHO98 emissions estimate. eIncludes combustion of bituminous/subbituminous coal and anthracite coal. fRefers to conventional pollutant control, not dioxin emissions control. Very few sources listed in this inventory control specifically for gCongener-specific emissions data were not available; the Nordic TEQ estimate was used as a surrogate for the I-TEQDF emissions estimhIncludes forest wildfires and prescribed burning for forest management. iTerm refers to the burning of residential waste in barrels. jIncludes burning of brush and leaves. kTotal reflects only the total of the estimates made in this report. 

Category ratings: A = Characterization of the source category judged to be adequate for quantitative estimation with high confidence in the emission f

confidence in activity level. 

B = Characterization of the source category judged to be adequate for quantitative estimation with medium confidence in the emissi

medium confidence in activity level. 

C = Characterization of the source category judged to be adequate for quantitative estimation with low confidence in either the emis

activity level. 

D = Preliminary indication of the potential magnitude of emissions from “Unquantified” (Category D) sources; based on extremely li

clearly nonrepresentative. D estimates are not included in the inventory of source emissions, but serve the purpose of highlightin

adequate emissions information. 

NA = Not available (information is lacking). POTW = Publicly owned treatment works or sewage treatment plant. 
4. 	 Reservoir sources with preliminary release estimates. These sources are discussed in 

Chapter 11. 

5. 	 Reservoir sources without quantified releases. These sources are discussed in 

Chapter 11.

 Only contemporary formation sources (numbers 1 and 2 above) are considered for 

inclusion in the national inventory. Reservoir sources are not considered because they are not 

original releases, but rather the recirculation of past releases. To date, no reliable estimates of 

releases from the reservoir sources have been made because information is either lacking or is 

inadequate to allow for estimates to be made. 

This document includes discussions on products that contain dioxin-like compounds. 

Some of these products, such as 2,4-D, are considered to be sources because they are clearly used 

in ways that result in environmental releases (e.g., they are sprayed onto agricultural lands for 

weed control). If a release from the product occurs, it is added to the national dioxin inventory. 

Other products containing dioxin-like compounds, such as vinyl chloride products, do not appear 

to have environmental releases and are not considered sources. For all CDD-/CDF-containing 

products, this document summarizes the available information about contamination levels and, 

where possible, makes estimates of the total amount of CDDs/CDFs produced annually in these 

products. Estimates of the CDD/CDF TEQ amounts in products are summarized in Table 1-13. 

Table 1-13. Products containing CDDs/CDFs (g TEQDF-WHO98 /yr) 

Product 2000 1995 1987 

Bleached chemical wood pulp 0.58 40 505 

Ethylene dichloride/vinyl chloride 0.02 0.02 NA 

Chloranil 1.16 64 NA 

Pentachlorophenol 4,395 4,800 20,000 

2,4 -Dichlorophenoxy acetic acid (2,4-D)

a NA 28.9 33.4 

TOTAL 4,397 4,933 20,538 a

 Only 2,4-D is considered to be an environmental release. 

NA = Information not available 

1.2.2. Quantitative Method for Inventory of Sources 

Some source types havea high percentage of facilities with measured CDD/CDFreleases, 

such as municipal waste combustion, hazardous waste incineration, and cement kilns that burn 

hazardous waste (air emissions), and wastewater releases from chlorine-bleached pulp and paper 

mills. In addition, some source activities have been adequately sampled with respect to levels of 

1-22
 

lands and 2,4-D applied to agricultural lands. Other source categories have relatively few tested 

facilities and/or the activity has not been comprehensively evaluated for dioxin releases. In these 

cases, EPA relies on the use of emission factors to estimate CDD/CDF releases from the untested 

sources. This provides a method of extrapolation from tested sources to national estimates of 

environmental releases. Many of the national emission estimates, therefore, have been developed 

using this “top-down” approach. 

The first step in this approach is to derive from the available emissions monitoring data 

an emission factor (or series of emission factors) deemed to be representative of the source 

category (or segments of a source category that differ in, e.g., configuration, fuel type, air 

pollution control equipment). The emission factor relates mass of CDDs/CDFs or dioxin-like 

PCBs released into the environment with some measure of activity (e.g., kilograms of material 

processed per year, vehicle miles traveled per year, liters of wastewater discharged per year). It is 

developed by averaging the emission factors for the tested facilities or activities within the 

particular classification of sources. For example, mass burn MWCs equipped with dry scrubbers 

(DSs) combined with fabric filters (FFs) will have an average emission factor derived from the 

tested facilities within this source classification. This average emission factor is then multiplied 

by the measure of activity for the nontested facilities in the class (e.g., total kilograms of material 

processed by these facilities annually). Finally, emissions are summed for the tested facilities 

and nontested facilities. In general, this procedure can be represented by the followinge quations: 

Etotal = 3Etested, I % 3Euntested, I (1-2) 

Etotal = 3Etested, I % 3(EFi * Ai)untested (1-3) 

where: 

Etotal = annual emissions from all facilities (g TEQ/yr)
 

Etested, I = annual emissions from all tested facilities in class I (g TEQ/yr)
 

Euntested, I = annual emissions from all untested facilities in class I (g TEQ/yr)
 

EFi

 = mean emission factor for tested facilities in class I (g TEQ/kg)
 

Ai = activity measure for untested facilities in class I (kg/yr)
 

Figures 1-2 through 1-4 and 1-6 through 1-8 depict the various source categories and their 

emission factors, activity levels, and annual emissions for reference years 1987, 1995, and 2000, 

respectively, in I-TEQ and WHO-TEQ units. Figures 1-5 and 1-9 depict comparisons of the 

estimated I-TEQ and WHO-TEQ air emissions for these years. 

1-23
 

Emission Source Best Estimate of Total Annual "Activity" 

(tested / total units) I-TEQ Emission Factor (thousand metric tons/yr or Annual I-TEQ Emission 

(ng/kg or ng/L) million L/yr) (g I-TEQ/yr) 

Municipal Waste Combustion (19 / 113)
 


Medical Waste Incineration (8 / 5000)
 


Secondary Copper Smelting (2 / 4)
 


Backyard Barrel Burning (NA)
 


Cement Kilns Burning Haz Waste (10 / ?)
 


Utility / Industrial Coal Combustion (11 / ?)
 


On-road Leaded Gas Fuel Combustion (? / ?)
 


Iron Ore Sinter Plants (2 / ?)
 


On-road Diesel Fuel Combustion (NA)
 


Industrial Wood Burning (9 / ?)
 


Residential Wood Burning (7 / 25000000)
 


Off-road Diesel Fuel Combustion (NA)
 


Utility / Industrial Residual Oil Combustion (? / ?)
 


Secondary Aluminum Smelting (6 / 67)
 


Cement Kilns Not Burning Haz Waste (15 / ?)
 


Utility / Industrial Distillate Oil Combustion (>2 / ?)
 


Sewage Sludge Incineration (13 / 199)
 


Hazardous Waste Incineration (17 / 171-227)
 


On-road Unleaded Gas Fuel Combustion (? / ?)
 


Manufacture of EDC/VC (? / ?)
 


0.00.101.00

10.00

100.00

1000.00

10000.00

100000.00

1

10

100

1,000

10,000

100,000

1,000,000

1.00

10.00

100.00

1,000.00

10,000.00

100,000.00 

1,000,000.00 

The figures include sources with annual I-TEQ emission estimates greater than 5 g I-TEQ/yr in 

Low Confidence 

Medium Confidence 

High Confidence 

Legend 

one or both of the Reference Years 1987 or 1995. Derivation of the emission factors 

and annual "Activity" estimates (e.g., kg of waste incinerated) are presented in the following 

chapters of this report. The difference in bar shading indicates the degree of confidence in the 

estimate. The set of numbers following the source categories indicates the number of
 


facilities/sites for which emission test data are available versus the number of facilities/sites
 


in the category. A question mark (?) indicates that the precise number of facilities/sites
 


could not be estimated.
 


Figure 1-2. Estimated CDD/CDF I-TEQ emissions to air from combustion 

sources in the United States for reference year 1987 (municipal solid waste 

incineration is currently referred to as municipal waste combustion). 

Some source categories are made up of facilities that vary widely in terms of design and 

operating conditions. For these sources, as explained above, an attempt was made to create 

subcategories that grouped facilities with common features and then to develop separate emission 

factors for each subcategory. Implicit in this procedure is the assumption that facilities with 

similar design and operating conditions should have similar CDD/CDF release potential. For 

most source categories, however, the specific combination of features that contributes mostto 

1-24
 

Emission Source
 


(tested / total units)
 


Municipal Waste Combustion (39 / 130)
 


Medical Waste Incineration (20 / 2400)
 


Secondary Copper Smelting (2 / 3)
 


Backyard Barrel Burning (NA)
 


Cement Kilns Burning Haz Waste (10 / 34)
 


Utility / Industrial Coal Combustion (11 / ?)
 


On-road Leaded Gas Fuel Combustion (? / ?)
 


Iron Ore Sinter Plants (2 / 11)
 


On-road Diesel Fuel Combustion (NA)
 


Industrial Wood Burning (9 / ?)
 


Residential Wood Burning (7 / 25000000)
 


Off-road Diesel Fuel Combustion (NA)
 


Utility / Industrial Residual Oil Combustion (? / ?)
 


Secondary Aluminum Smelting (6 / 76)
 


Cement Kilns Not Burning Haz Waste (15 / 178)
 


Utility / Industrial Distillate Oil Combustion (>2 / ?)
 


Sewage Sludge Incineration (13 / 257)
 


Hazardous Waste Incineration (17 / 162)
 


On-road Unleaded Gas Fuel Combustion (? / ?)
 


Manufacture of EDC/VC (? / ?)
 


Best Estimate of Total Annual "Activity" 

I-TEQ Emission Factor (thousand metric tons/yr or Annual I-TEQ Emission 

(ng/kg or ng/L) million L/yr) (g I-TEQ/yr) 

0.01

0.1

1

10

100

1000

10000

100000

1

10

100

1,000

10,000

100,000

1,000,000

1

10

100

1000

10000

100000

1000000 

The figures include sources with annual I-TEQ emission estimates greater than 5 g I-TEQ/yr in Legend 

one or both of the Reference Years 1987 or 1995. Derivation of the emission factors Low Confidence 

and annual "Activity" estimates (e.g., kg of waste incinerated) are presented in the following Medium Confidence 

chapters of this report. The difference in bar shading indicates the degree of confidence in the High Confidence 

estimate. The set of numbers following the source categories indicates the number of
 


facilities/sites for which emission test data are available versus the number of facilities/sites
 


in the category. A question mark (?) indicates that the precise number of facilities/sites
 


could not be estimated.
 


Figure 1-3. Estimated CDD/CDF I-TEQ emissions to air from combustion 

sources in the United States for reference year 1995 (municipal solid waste 

incineration is currently referred to as municipal waste combustion). 

CDD/CDF or dioxin-like PCB releases is not well understood. Therefore, how to best 

subcategorize a source category was often problematic. For each subcategorized source category 

in this document, a discussion is presented about the variability in design and operating 

conditions, what was known about how these features contributed to CDD/CDF or dioxin-like 

PCB releases, and the rationale for creating subcategories. 

1-25
 
Emission Source
 


(tested / total units)
 


Municipal Waste Combustion (195 / 251)
 


Medical Waste Incineration (22 / ?)
 


Secondary Copper Smelting (1 / 2)
 


Backyard Barrel Burning (NA)
 


Cement Kilns Burning Haz Waste (10 / ?)
 


Utility / Industrial Coal Combustion (11 / ?)
 


On-road Leaded Gas Fuel Combustion (0 / 0)
 


Iron Ore Sinter Plants (2 / 11)
 


On-road Diesel Fuel Combustion (NA)
 


Industrial Wood Burning (9 / ?)
 


Residential Wood Burning (19 / 25000000)
 


Off-road Diesel Fuel Combustion (NA)
 


Utility / Industrial Residual Oil Combustion (? / ?)
 


Secondary Aluminum Smelting (6 / ?)
 


Cement Kilns Not Burning Haz Waste (15 / ?)
 


Utility / Industrial Distillate Oil Combustion (>2 / ?)
 


Sewage Sludge Incineration (14 / ?)
 


Hazardous Waste Incineration (22 / 132)
 


On-road Unleaded Gas Fuel Combustion (? / ?)
 


Manufacture of EDC/VC (8 / 12)
 


Best Estimate of Total Annual "Activity" 

I-TEQ Emission Factor (thousand metric tons/yr or Annual I-TEQ Emission 

(ng/kg or ng/L) million L/yr) (g I-TEQ/yr) 

0.010.101.00

10.00

100.00

1000.00

10000.00

100000.00

1

10

100

1,000

10,000

100,000

1,000,000

0.1

1

10

100

1000

10000

100000

1000000 

The figures include sources with annual I-TEQ emission estimates greater than 5 g I-TEQ/yr in Legend 

one or both of the Reference Years 1987 or 1995. Derivation of the emission factors Low Confidence 

and annual "Activity" estimates (e.g., kg of waste incinerated) are presented in the following Medium Confidence 

chapters of this report. The difference in bar shading indicates the degree of confidence in the High Confidence 

estimate. The set of numbers following the source categories indicates the number of
 

facilities/sites for which emission test data are available versus the number of facilities/sites
 

in the category. A question mark (?) indicates that the precise number of facilities/sites
 

could not be estimated.
 

Figure 1-4. Estimated CDD/CDF I-TEQ emissions to air from combustion 

sources in the United States for reference year 2000. 

The emission factors developed for the inventory are intended to be used for estimating 

total emissions for a source category rather than emissions from individual facilities. EPA has 

made uncertainty determinations for each of these emission factors, based, in part, on the 

assumption that by applying them to a group of facilities, the potential for overestimating or 

underestimating individual facilities will, to some extent, be self-compensating. This means that 

in using these emission factors one can place significantly greater confidence in an emission 

1-26
 

Municipal Waste Combustion 

Medical Waste Incineration 

Secondary Copper Smelting 

Backyard Barrel Burning 

Cement Kilns Burning Haz Waste 

Utility / Industrial Coal Combustion 

On-road Leaded Gas Fuel Combustion 

Iron Ore Sinter Plants 

On-road Diesel Fuel Combustion 

Industrial Wood Burning 

Residential Wood Burning 

Off-road Diesel Fuel Combustion 

Utility / Industrial Residual Oil Combustion 

Secondary Aluminum Smelting 

Cement Kilns Not Burning Haz Waste 

Utility / Industrial Distillate Oil Combustion 

Sewage Sludge Incineration 

Hazardous Waste Incineration 

On-road Unleaded Gas Fuel Combustion 

Manufacture of EDC/VC 

0.1 1 10 100 1000 10000 

1987 1995 2000 

Figure 1-5. Comparison of estimates of annual I-TEQ emissions to air 

(g I-TEQ/yr) for reference years 1987, 1995, and 2000. 

estimate for a class than in an estimate for any individual facility. Given the limited amount of 

data available for deriving emission factors and the limitations of our understanding about 

facility-specific conditions that determine formation and control of dioxin-like compounds, the 

current state of knowledge cannot support the development of emission factors that can be used 

to accurately estimate emissions on an individual facility-specific basis. The emission factors 

developed for each of the categories discussed in this national emissions inventory are listed in I

TEQ and TEQDF-WHO98 in Tables 1-14 and 1-15, respectively. 

1.2.3. Confidence Ratings 

Each source emission calculation required estimates of an emission factor and an activity 

level. For each emission source, the quantity and quality of the available information for both 

vary considerably. Consequently, it is important that emission estimates be accompanied by 

some indicator of the uncertainties associated with their development. For this reason, a 

­

1-27
 

Emission Source
 


(tested / total units)
 


Municipal Waste Combustion (19 / 105)
 


Medical Waste Incineration (8 / 5000)
 


Secondary Copper Smelting (2 / 4)
 


Backyard Barrel Burning (NA)
 


Cement Kilns Burning Haz Waste (10 / ?)
 


Utility / Industrial Coal Combustion (11 / ?)
 


On-road Leaded Gas Fuel Combustion (? / ?)
 


Iron Ore Sinter Plants (2 / ?)
 


On-road Diesel Fuel Combustion (NA)
 


Industrial Wood Burning (9 / ?)
 


Residential Wood Burning (7 / 25000000)
 


Off-road Diesel Fuel Combustion (NA)
 


Utility / Industrial Residual Oil Combustion (? / ?)
 


Secondary Aluminum Smelting (6 / 67)
 


Cement Kilns Not Burning Haz Waste (15 / ?)
 


Utility / Industrial Distillate Oil Combustion (>2 / ?)
 


Sewage Sludge Incineration (13 / 199)
 


Hazardous Waste Incineration (17 / 171-227)
 


On-road Unleaded Gas Fuel Combustion (? / ?)
 


Manufacture of EDC/VC (? / ?)
 


Best Estimate of Total Annual "Activity" 

WHO-TEQ Emission Factor (thousand metric tons/yr or Annual WHO-TEQ Emission 

(ng/kg or ng/L) million L/yr) (g WHO-TEQ/yr) 

0.01

0.1

1

10

100

1000

10000

100000

1

10

100

1,000

10,000

100,000

1,000,000

1

10

100

1000

10000

100000

1000000 

The figures include sources with annual WHO-TEQ emission estimates greater than 5 g WHO-TEQ/yr in Legend 

one or both of the Reference Years 1987 or 1995. Derivation of the emission factors Low Confidence 

and annual "Activity" estimates (e.g., kg of waste incinerated) are presented in the following Medium Confidence 

chapters of this report. The difference in bar shading indicates the degree of confidence in the High Confidence 

estimate. The set of numbers following the source categories indicates the number of
 


facilities/sites for which emission test data are available versus the number of facilities/sites
 


in the category. A question mark (?) indicates that the precise number of facilities/sites
 


could not be estimated.
 


Figure 1-6. Estimated CDD/CDF WHO-TEQ emissions to air from 

combustion sources in the United States for reference year 1987 (municipal 

solid waste incineration is currently referred to as municipal waste combustion). 

qualitative confidence rating scheme was developed as an integral part of the emissions estimate 

with the following considerations. 

Emission factor. The uncertainty in the emission factor estimate depends primarily on 

how well the tested facilities represent the untested facilities. In general, confidence in the 

emission factor increases with increases in the number of tested facilities relative to the total 

number of facilities. Variability in terms of physical design and operating conditions within a 

1-28
 

r

Emission Source 

(tested / total units) 

al Waste Combustion (39 / 130) 

Waste Incineration (20 / 2400) 

ary Copper Smelting (2 / 3) 

rd Barrel Burning (NA) 

Kilns Burning Haz Waste (10 / 34) 

ndustrial Coal Combustion (11 / ?) 

d Leaded Gas Fuel Combustion (? / ?) 

e Sinter Plants (2 / 11) 

d Diesel Fuel Combustion (NA) 

al Wood Burning (9 / ?) 

ntial Wood Burning (7 / 25000000) 

d Diesel Fuel Combustion (NA) 

ndustrial Residual Oil Combustion (? / ?) 

ary Aluminum Smelting (6 / 76) 

Kilns Not Burning Haz Waste (15 / 178) 

ndustrial Distillate Oil Combustion (>2 / ?) 

e Sludge Incineration (13 / 257) 

ous Waste Incineration (17 / 162) 

d Unleaded Gas Fuel Combustion (? / ?) 

cture of EDC/VC (? / ?) 

Best Estimate of Total Annual "Activity" 

WHO-TEQ Emission Factor (thousand metric tons/yr or Annual WHO-TEQ Emission 

(ng/kg or ng/L) million L/yr) (g WHO-TEQ/yr) 

0.01

0.1

1

10

100

1000

10000

100000 

1

10

100

1,000

10,000

100,000

1,000,000

1

10

100

1000

10000

100000

1000000 

gures include sources with annual WHO-TEQ emission estimates greater than 5 g WHO-TEQ/yr in Legend 

both of the Reference Years 1987 or 1995. Derivation of the emission factors Low Confidence 

nnual "Activity" estimates (e.g., kg of waste incinerated) are presented in the following Medium Confidence 

ers of this report. The difference in bar shading indicates the degree of confidence in the High Confidence 

ate. The set of numbers following the source categories indicates the number of
 


es/sites for which emission test data are available versus the number of facilities/sites
 


category. A question mark (?) indicates that the precise number of facilities/sites
 


not be estimated.
 


Figure 1-7. Estimated CDD/CDF WHO-TEQ emissions to air from 

combustion sources in the United States for reference year 1995. 

class or subclass must also be considered. The more variability among facilities, the less 

confidence that a test of any single facility is representative of that class or subclass. The quality 

of the supporting documentation also affects uncertainty. Whenever possible, original 

engineering test reports were used. Peer-reviewed reports from the open literature were also used 

for developing some emission factors. In some cases, however, draft reports that had undergone 

more limited review were also used. In a few cases, unpublished references (such as personal 

communication with experts) were used and are clearly noted in the text. 

1-29
 
Emission Source
 


(tested / total units)
 


Municipal Waste Combustion (195 / 251)
 


Medical Waste Incineration (22 / ?)
 


Secondary Copper Smelting (1 / 2)
 


Backyard Barrel Burning (NA)
 


Cement Kilns Burning Haz Waste (20 / 22)
 


Utility / Industrial Coal Combustion (11 / ?)
 


On-road Leaded Gas Fuel Combustion (0 / 0)
 


Iron Ore Sinter Plants (4 / 11)
 


On-road Diesel Fuel Combustion (NA)
 


Industrial Wood Burning (9 / ?)
 


Residential Wood Burning (19 / ?)
 


Off-road Diesel Fuel Combustion (NA)
 


Utility / Industrial Residual Oil Combustion (? / ?)
 


Secondary Aluminum Smelting (6 / ?)
 


Cement Kilns Not Burning Haz Waste (15 / ?)
 


Utility / Industrial Distillate Oil Combustion (>2 / ?)
 


Sewage Sludge Incineration (14 / ?)
 


Hazardous Waste Incineration (22 / 132)
 


On-road Unleaded Gas Fuel Combustion (? / ?)
 


Manufacture of EDC/VC/PVC (8 / 12)
 


Best Estimate of Total Annual "Activity" 

WHO-TEQ Emission Factor (thousand metric tons/yr or Annual WHO-TEQ Emission 

(ng/kg or ng/L) million L/yr) (g WHO-TEQ/yr) 

0.010.101.00

10.00

100.00

1000.00

10000.00

100000.00

1

10

100

1,000

10,000

100,000

1,000,000

0.1

1

10

100

1000

10000

100000

1000000 

The figures include sources with annual WHO-TEQ emission estimates greater than 5 g WHO-TEQ/yr in Legend 

one or both of the Reference Years 1987 or 1995. Derivation of the emission factors Low Confidence 

and annual "Activity" estimates (e.g., kg of waste incinerated) are presented in the following Medium Confidence 

chapters of this report. The difference in bar shading indicates the degree of confidence in the High Confidence 

estimate. The set of numbers following the source categories indicates the number of
 


facilities/sites for which emission test data are available versus the number of facilities/sites
 


in the category. A question mark (?) indicates that the precise number of facilities/sites
 


could not be estimated.
 


Figure 1-8. Estimated CDD/CDF WHO-TEQ emissions to air from
 


combustion sources in the United States for reference year 2000.
 


Activity level. The uncertainty in the activity level estimate was judged primarily on the 

basis of the extent of the underlying data. Estimates derived from comprehensive surveys 

(including most facilities in a source category) were assigned high confidence. As the number of 

facilities in the survey relative to the total decreased, confidence also decreased. The quality of 

the supporting documentation also affects uncertainty. Peer-reviewed reports from the open 

literature (including government and trade association survey data) were considered to be the 

1-30
 

Municipal Waste Combustion 

Medical Waste Incineration 

Secondary Copper Smelting 

Backyard Barrel Burning 

Cement Kilns Burning Haz Waste 

Utility / Industrial Coal Combustion 

On-road Leaded Gas Fuel Combustion 

Iron Ore Sinter Plants 

On-road Diesel Fuel Combustion 

Industrial Wood Burning 

Residential Wood Burning 

Off-road Diesel Fuel Combustion 

Utility / Industrial Residual Oil Combustion 

Secondary Aluminum Smelting 

Cement Kilns Not Burning Haz Waste 

Utility / Industrial Distillate Oil Combustion 

Sewage Sludge Incineration 

Hazardous Waste Incineration 

On-road Unleaded Gas Fuel Combustion 

Manufacture of EDC/VC 

0.1 1 10 100 1000 10000 

1987 1995 2000 

Figure 1-9. Comparison of estimates of annual WHO-TEQ emissions to air 

(g WHO-TEQ/yr) for reference years 1987, 1995, and 2000. 

most reliable. However, as with the emission factor estimates, draft reports that had undergone 

more limited review were used in some cases, and in a few cases unpublished references such as 

personal communication with experts were used. These are clearly noted in the text. 

1.2.3.1. Rating Scheme 

The confidence rating scheme shown in Table 1-16 represents the qualitative criteria used 

to assign a high, medium, or low confidence rating to emission factors and activity levels for 

those source categories for which emission estimates could be reliably quantified. The overall 

confidence rating assigned to an emissions estimate was determined by the confidence ratings 

assigned to the corresponding activity level and emission factor. If the lowest rating assigned to 

either the activity level or the emission factor is “high,” then the category rating assigned to the 

emission estimate is high (Category A). If the lowest rating assigned to either the activity level 

1-31
 

Table 1-14. I-TEQDF emission factors used to develop national emission inventory estimates of rele

1-32
 


Source category 

I-TEQDF emission factor 

2000 1995 1987 Emis

WASTE INCINERATION 

Municipal waste combustion 

Hazardous waste incineration Boilers/industrial furnaces Halogen acid furnaces Medical waste/pathological incinerationCrematoria - human

 - animal 

Sewage sludge incineration Tire combustion Pulp and paper mill sludge incineratorsb 

2.82 

2.12 

1.21 

0.803 

630 

a

 410 

0.11 

6.65 

0.282 

38.2a

3.83 

0.64 

598 

a

17,000

6.94 

0.282 

573 

a 

3.83 

0.64 

1,706 

a 

17,000 

6.94 

0.282 

ng TEQ/kg wa

ng TEQ/kg wa

ng TEQ/kg wa

ng TEQ/kg wa

ng TEQ/kg wa

ng TEQ/body

ng TEQ/kg ani

ng TEQ/kg dry 

ng TEQ/kg tire

POWER/ENERGY GENERATION 

Vehicle fuel combustion - leadedc 

- unleaded - diesel 

Wood combustion - residential 

- industrial 

Coal combustion - utility 

Oil combustion - industrial/utility 

NA 

1.5 

172 

0.5 

0.56–13.2d 

0.079 

0.2 

45 

1.5 

172 

2 

0.56–13.2d 

0.079 

0.2 

45 

1.5 

172 

2 

0.56–13.2d 

0.079 

0.2 

pg TEQ/km dri

pg TEQ/km dri

pg TEQ/km dri

ng TEQ/kg wo

ng TEQ/kg wo

ng TEQ/kg coa

ng TEQ/L oil c

OTHER HIGH-TEMPERATURE SOURCES 

Cement kilns burning hazardous waste Lightweight aggregate kilns 

Cement kilns not burning hazardous waste 

Petroleum refining catalyst regeneration Cigarette combustion Carbon reactivation furnaces Kraft recovery boilers 

1.444 

2.06 

0.27 

1.52 

0.00043–0.0029 

1.2 

0.029 

1.04–28.58e 

0.27 

1.52 

0.00043–0.0029 

1.2 

0.029 

1.04–28.58e 

0.27 

1.52 

0.00043–0.0029 

1.2 

0.029 

ng TEQ/kg clin

ng TEQ/ kg wa

ng TEQ/kg clin

ng TEQ/barrel 

ng TEQ/cigare

ng TEQ/kg of r

ng TEQ/kg sol
Table 1-14. I-TEQDF emission factors used to develop national emission inventory estimates of rele

(continued) 

Source category 

I-TEQDF emission factor 

2000 1995 1987 Emis

MINIMALLY CONTROLLED OR UNCONTROLLED COMBUSTION 

Backyard barrel burningf 72.8 72.8 72.8 ng TEQ/kg wa

METALLURGICAL PROCESSES 

Ferrous metal smelting/refining - sintering plants Nonferrous metal smelting/refining 

- primary copper - secondary aluminum smelting - secondary copper smeltingg 

- secondary lead smelters 

Drum and barrel reclamation 

0.55–4.14 

<0.31 

4.9 

0.05–8.31 

16.5 

0.55–4.14 

<0.31 

4.9 

0.05–8.31 

16.5 

0.55–4.14 

<0.31 

4.9 

0.05–8.31 

16.5 

ng TEQ/kg sin

ng TEQ/kg cop

ng TEQ/kg scr

ng TEQ/kg scr

ng TEQ/kg lea

ng TEQ/drum 

CHEMICAL MANUFACTURING/PROCESSING SOURCES 

Ethylene dichloride/vinyl chloride 0.95a ng TEQ/kg ED

aDifferent emission factors were derived for various subcategories within this industry; the value listed is a weighted average. bIncluded in total for wood combustion, industrial. cLeaded fuel production and the manufacture of motor vehicle engines requiring leaded fuel for highway use are prohibited in the Unite

 (See Chapter 4, Section 4.1, for details.) 

dEmission factor of 0.56 ng I-TEQDF/kg used for nonsalt-laden wood; emission factor of 13.2 ng I-TEQDF/kg used for salt-laden wood. eEmission factor of 1.04 ng I-TEQDF/kg used for kilns with air pollution control device (APCD) inlet temperatures less than 232°C; emi

 I-TEQDF/kg used for kilns with APCD inlet temperatures greater than 232°C. 

fIncludes the burning of brush and leaf residential yard waste. gFacility-specific emission factors were used ranging from 3.6 to 16,600 ng I-TEQDF/kg scrap consumed. 

1-33
 

Table 1-15. TEQDF-WHO98 emission factors used to develop national emission inventory estimates 

to air 

1-34
 


Emission source category 

TEQDF-WHO98 emission factor 

2000 1995 1987 Emis

WASTE INCINERATION 

Municipal waste combustion 

Hazardous waste incineration Boilers/industrial furnaces Halogen acid furnaces Medical waste/pathological incinerationCrematoria - humanb 

 - animal 

Sewage sludge incineration Tire combustion Pulp and paper mill sludge incineratorsc 

3.10

2.13 

1.21 

0.836 

630 

a

434 

0.12 

6.74 

0.281 

43.4 

a

3.88 

0.65 

633 

a

17,000 

7.04 

0.281 

644 

a 

3.88 

0.65 

1,811 

a 

17,000 

7.04 

0.281 

ng TEQ/kg w

ng TEQ/kg w

ng TEQ/kg w

ng TEQ/kg w

ng TEQ/kg w

ng TEQ/body

ng TEQ/kg an

ng TEQ/kg dr

ng TEQ/kg tir

POWER/ENERGY GENERATION 

Vehicle fuel combustion - leadedd 

- unleaded - diesel 

Wood combustion - residential

- industriale 

Coal combustion - utility 

Oil combustion - industrial/utility 

NA 

1.6 

182 

0.5 

b

0.6–13.2 

0.78 

0.23 

53 

1.6 

182 

2 

b

0.6–13.2 

0.078 

0.23 

53 

1.6 

182 

2 

b 

0.6–13.2 

0.078 

0.23 

pg TEQ/km d

pg TEQ/km d

pg TEQ/km d

ng TEQ/kg w

ng TEQ/kg w

ng TEQ/kg co

ng TEQ/L oil 

OTHER HIGH-TEMPERATURE SOURCES 

Cement kilns burning hazardous waste

Lightweight aggregate kilns 

Cement kilns not burning hazardous waste Petroleum refining catalyst regeneration Cigarette combustion Carbon reactivation furnaces 

Kraft recovery boilers 

1.635 1.99 

0.26 

1.59 

0.00044–0.003 

1.2b 

0.028 

1.11–30.7f 

0.26 

1.59 

0.00044–0.003 

1.2b 

0.028 

1.11–30.7f 

0.26 

1.59 

0.00044–0.003 

1.2b 

0.028 

ng TEQ/kg cl

ng TEQ/ kg w

ng TEQ/kg cl

ng TEQ/barre

ng TEQ/cigar

ng TEQ/kg of 

ng TEQ/kg so
Table 1-15. TEQDF-WHO98 emission factors used to develop national emission inventory estimates 

to air (continued) 

Emission source category 

TEQDF-WHO98 emission factor 

2000 1995 1987 Emis

MINIMALLY CONTROLLED OR UNCONTROLLED COMBUSTION 

Backyard barrel burningg 76.8b 76.8b 76.8b ng TEQ/kg w

METALLURGICAL PROCESSES 

Ferrous metal smelting/refining - sintering plants Nonferrous metal smelting/refining 

- primary copperb 

- secondary aluminum smelting - secondary copper smeltingh 

- secondary lead smelters 

Drum and barrel reclamation 

0.62–4.61 

<0.31 

5.2 

0.05–8.81 

17.5 

0.62–4.61 

<0.31 

5.2 

0.05–8.81 

17.5 

0.62–4.61 

<0.31 

5.2 

0.05–8.81 

17.5 

ng TEQ/kg sin

ng TEQ/kg co

ng TEQ/kg sc

ng TEQ/kg sc

ng TEQ/kg le

ng TEQ/drum 

CHEMICAL MANUFACTURING/PROCESSING SOURCES 

Ethylene dichloride/vinyl chloride 0.95a, b ng TEQ/kg ED

aDifferent emission factors were derived for various subcategories within this industry; the value listed is a weighted average. bCongener-specific data were not available; the TEQDF emission factor was used as a surrogate for the TEQDF-WHO98 emission factor. cIncluded in total for wood combustion, industrial. dLeaded fuel production and the manufacture of motor vehicle engines requiring leaded fuel for highway use are prohibited in the Unite

 (See Chapter 4, Section 4.1, for details). 

eEmission factor of 0.6 ng I-TEQDF/kg used for non-salt-laden wood; emission factor of 13.2 ng I-TEQDF/kg used for salt-laden wood. fEmission factor of 1.11 ng I-TEQDF/kg used for kilns with air pollution control device (APCD) inlet temperatures less than 232°C; emis

 I-TEQDF/kg used for kilns with APCD inlet temperatures greater than 232°C. 

gThis term refers to the burning of residential waste in barrels. hFacility-specific emission factors were used ranging from 3.6 to 16,600 ng TEQDF-WHO98/kg scrap consumed. 

1-35
 

Table 1-16. Confidence rating scheme for U.S. emission estimates 

Confidence rating Activity level estimate Emission factor estimate 

Categories/media for which releases can be reasonably quantified 

High Derived from comprehensive 

survey 

Derived from comprehensive survey 

Medium Based on estimates of average plant 

activity level and number of plants 

or limited survey 

Derived from testing at a limited but 

reasonable number of facilities 

believed to be representative of 

source category 

Low Based on data judged possibly 

nonrepresentative 

Derived from testing at only a few, 

possibly nonrepresentative facilities 

or from similar source categories 

Categories/media for which releases cannot be reasonably quantified 

Preliminary estimate Based on extremely limited data, 

judged to be clearly 

nonrepresentative 

Based on extremely limited data, 

judged to be clearly 

nonrepresentative 

Not quantified No data available (1) Argument based on theory but 

no data, or 

(2) Data available indicating 

formation but not in a form that 

allows developing an emission 

factor 

or the emission factor term is “medium,” then the category rating assigned to the emission 

estimate is medium (Category B). If the lowest rating assigned to either the activity level or the 

emission factor is “low,” then the category rating assigned to the emission estimate is low 

(Category C). It is emphasized that this confidence rating scheme should not be interpreted as a 

statistical measure, but rather as subjective judgment of the relative uncertainty among sources. 

For many source categories, either emission factor information or activity level 

information was inadequate to support development of reliable quantitative release estimates for 

one or more media. For some of these source categories, sufficient information was available to 

make preliminary estimates of emissions of CDDs/CDFs or dioxin-like PCBs; however, the 

confidence in the activity level estimates or emission factor estimates was so low that they could 

not be included in the sum of quantified emissions from sources with confidence ratings of A, B, 

and C. These preliminary estimates were given an overall confidence rating of D. The 

preliminary release estimates for sources with a confidence rating of D are given in the rightmost column of Table 1-12. Because these are order-of-magnitude estimates, they are made for 

2000 only. As preliminary estimates of source magnitude, they can be used to help prioritize 

1-36
 

future research and data collection. The actual magnitude of emissions from these sources could 

be significantly lower or higher than these preliminary estimates. Although EPA has chosen not 

to include them in the more thoroughly characterized emissions of the national inventory, some 

of these poorly characterized sources have the potential of being major contributors of releases to 

the environment. It is important to present these estimates because they may help determine 

priorities for future data collection efforts. As the uncertainty around these sources is reduced, 

they will be included in future inventory calculations. 

For other sources, some information exists that suggests that they may release dioxin-like 

compounds; however, the available data were judged to be insufficient for developing any 

quantitative emissions estimate. These source categories were assigned a confidence rating of E 

and also were not included in the national inventory (see the “Not quantifiable” column in Table 

1-11). 

1.3. CONCLUSIONS 

1.3.1. Total Environmental Releases 

Nationwide emission estimates of grams I-TEQDF and TEQDF-WHO98 released to the open 

and circulating environment of the United States are presented in Table 1-12. For the year 2000, 

EPA draws the following conclusions: 

•	 The total releases in the inventory (Categories A, B, and C) were 1,422 g TEQDFWHO98/yr. These were dominated by releases to the air (92%). Most of the air 

releases were from combustion sources. Table 1-17 presents a ranking of sources for 

2000, 1995, and 1987 based on the magnitude of environmental release. The top 

three sources were backyard barrel burning of refuse (498.5 g, 32% of total), MWIs 

(378 g, 27%), and the incineration and land application of municipal wastewater 

treatment sludge (89.7 g, 5%). 

•	 There is a significant potential for release of dioxin-like compounds from Category D 

sources. However, these sources a currently poorly characterized. The most important 

Category D sources are forest fires and accidental fires at MSW landfills. Research is 

recommended to confirm emissions from these sources and to provide a more 

accurate assessment of releases. 

•	 A total of 18 contemporary formation sources were classified as Category E. 

Information suggests these may be sources of dioxin-like compounds, but it is 

insufficient to make a national estimate of releases. Additional research on these 

sources is recommended in order to adequately identify them as actual sources and to 

provide data for estimating releases. 

1-37
 

Table 1-17. Ranking of sources of dioxin-like compounds based on environmental releases (from h

for reference years 2000, 1995, and 1987 

1-38
 


2000 1995 198

Releases Percent Releases Percent 

Source (released to) (grams) of total Source (released to) (grams) of total Source (released to) 

Backyard barrel 498.5 35.1 Municipal waste 1,393.5 40.5 Municipal waste 

burning of refuse (air) combustion (air) combustion (air) Medical 378.0 26.6 Backyard barrel 628.0 18.2 Medical waste/ 

waste/pathological burning of refuse (air) pathological incineration incineration (air) (air) Municipal wastewater 89.7 6.3 Medical 487.0 14.1 Secondary copper 

treatment sludge, land waste/pathological smelters (air) application and incineration (air) incineration (land, air) Municipal waste combustion (air) 

83.8 5.9 Secondary copper 

smelters (air) 

271.0 7.9 Backyard barrel burning 

of refuse (air) 

Coal fired-utility 69.5 4.9 Cement kilns burning 156.1 4.5 Bleached chemical boilers (air) hazardous waste (air) wood pulp and paper 

mills (land, water) 

Diesel heavy-duty trucks (air) 

65.4 4.6 Municipal wastewater 

treatment sludge, land application and incineration (land and air) 

133.3 3.9 Cement kilns burning 

hazardous waste (air) 

Industrial wood 41.5 2.9 Coal fired-utility 60.1 1.7 Municipal wastewater 

combustion (air) boilers (air) treatment sludge, land 

application and incineration (air, land) 

Diesel off-road 33.1 2.3 Ethylene 35.7 1.0 Coal fired-utility boilers 

equipment, ships, dichloride/vinyl (air) trains, tractors (air) chloride production 

(land, air, water) 
Table 1-17. Ranking of sources of dioxin-like compounds based on environmental releases (from h

for reference years 2000, 1995, and 1987 (continued) 

1-39
 


2000 1995 198

Source (released to) 

Releases 

(grams) 

Percent 

of total Source (released to) 

Releases 

(grams) 

Percent 

of total Source (released to) 

Ethylene dichloride/ vinyl chloride production (water, land, air) 

30.0 2.1 Diesel heavy-duty 

trucks (air) 

33.3 1.0 Automobiles using 

leaded gasoline (air) 

Sintering plants (air) 27.6 1.9 Bleached chemical 

wood pulp and paper mills (land, water) 

30.0 0.9 2,4-Dichlorophenoxy 

acetic acid (land) 

Cement kilns burning hazardous waste (air) 

18.8 1.3 2,4-Dichlorophenoxy 

acetic acid (land) 

28.9 0.8 Sintering plants (air) 

Cement kilns burning nonhazardous waste (air) 

17.2 1.2 Sintering plants (air) 28.0 0.8 Diesel heavy-duty trucks 

(air) 

Residential wood combustion (air) 

11.3 0.8 Industrial wood 

combustion (air) 

26.2 0.8 Industrial wood 

combustion (air) 

Secondary aluminum smelting (air) 

8.3 0.6 Diesel off-road 

equipment: ships, trains, tractors (air) 

23.8 0.7 Residential wood 

combustion (air) 

Industrial/utility oil combustion, distillate oil (air) 

7.3 0.5 Secondary aluminum 

smelters (air) 

19.5 0.6 Diesel off-road 

equipment: ships, trains, tractors (air) 

Automobiles using unleaded gasoline (air) 

7.0 0.5 Cement kilns burning 

nonhazardous waste (air) 

16.6 0.5 Industrial/utility oil 

combustion, residual oil (air) 

Residential heating, distillate oil (air) 

4.5 0.3 Residential wood 

combustion (air) 

15.7 0.5 Cement kilns burning 

nonhazardous waste (air) 

Primary magnesium production (air) 

4.3 0.3 Industrial/utility oil 

combustion, residual oil (air) 

10.7 0.3 Secondary aluminum 

smelting (air) 
Table 1-17. Ranking of sources of dioxin-like compounds based on environmental releases (from h

for reference years 2000, 1995, and 1987 (continued) 

1-40
 


2000 1995 198

Source (released to) 

Releases 

(grams) 

Percent 

of total Source (released to) 

Releases 

(grams) 

Percent 

of total Source (released to) 

Chlor alkali facilities (air, water) 

3.6 0.3 Industrial/utility oil 

combustion, distillate oil (air) 

7.3 0.2 Industrial/utility oil 

combustion, distillate oil (air) 

Hazardous waste incineration (air) 

3.2 0.2 Hazardous waste 

incineration (air) 

5.8 0.2 Residential heating, 

distillate oil (air) 

Institutional/commerci al heating, distillate oil (air) 

2.9 0.2 Residential heating, 

distillate oil (air) 

5.0 0.1 Hazardous waste 

incineration (air) 

Secondary lead smelting (air) 

2.5 0.2 Automobiles using 

unleaded gasoline (air) 

4.7 0.1 Institutional/commercial 

heating, distillate oil (air) 

Petroleum refining catalyst regeneration (air) 

2.2 0.2 Primary magnesium 

production (air) 

4.1 0.1 Automobiles using 

unleaded gasoline (air) 

Lightweight aggregate kilns burning hazardous waste (air) 

1.9 0.1 Chlor-alkali facilities 

(air, water) 

3.6 0.1 Lightweight aggregate 

kilns burning hazardous waste (air) 

Boilers/industrial furnaces (air) 

1.8 0.1 Institutional/commercia 

l heating, distillate oil (air) 

3.1 0.1 Petroleum refining 

catalyst regeneration (air) 

Industrial/utility oil combustion, residual oil (air) 

1.7 0.1 Lightweight aggregate 

kilns burning hazardous waste (air) 

2.4 0.1 Kraft recovery boilers 

(air) 

Bleached chemical wood pulp and paper mills (land, water) 

1.1 0.1 Kraft recovery boilers 

(air) 

2.3 0.1 Industrial/utility oil 

combustion, residual oil (air) 

Secondary copper smelting (air) 

0.9 0.1 Petroleum refining 

catalyst regeneration (air) 

2.2 0.1 Secondary lead smelting 

(air) 
Table 1-17. Ranking of sources of dioxin-like compounds based on environmental releases (from h

for reference years 2000, 1995, and 1987 (continued) 

1-41
 


2000 1995 198

Source (released to) 

Releases 

(grams) 

Percent 

of total Source (released to) 

Releases 

(grams) 

Percent 

of total Source (released to) 

Kraft recovery boilers (air) 

0.8 0.1 Secondary lead 

smelting (air) 

1.7 0.05 Cigarette combustion 

(air) 

Institutional/commerci al heating, residual oil (air) 

0.7 0.05 Automobiles using 

leaded gasoline (air) 

1.6 0.05 Boilers/industrial 

furnaces (air) 

Drum and barrel reclamation (air) 

0.6 0.04 Industrial/utility oil 

combustion, residual oil (air) 

0.8 0.02 Primary copper smelting 

(air) 

Tire incineration (air) 0.5 0.04 Cigarette combustion 

(air) 

0.8 0.02 Crematoria, human (air) 

Cigarette combustion (air) 

0.4 0.03 Primary copper 

smelting (air) 

0.5 0.01 Carbon reactivation 

furnaces (air) 

Unleaded gasoline off-road equipment (air) 

0.4 0.03 Boilers/industrial 

furnaces (air) 

0.4 0.01 Tire incineration (air) 

Halogen acid furnaces (air) 

0.3 0.02 Crematoria, human 

(air) 

0.2 0.01 Drum and barrel 

reclamation (air) 

Primary copper smelting (air) 

0.3 0.02 Tire incineration (air) 0.1 <0.01 

Crematoria, human (air) 

0.3 0.02 Carbon reactivation 

furnaces (air) 

0.1 <0.01 

Carbon reactivation furnaces (air) 

0.1 0.01 Drum and barrel 

reclamation (air) 

0.1 <0.01 

Automobiles using leaded gasoline (air) 

0 0.00 

TOTAL 1,422.0 100 3,444.2 100 
•	 Releases from reservoir sources could significantly add to the inventory. However, 

environmental reservoirs as sources of the redistribution of previously formed dioxins 

into the open and circulating environment are currently poorly understood and poorly 

characterized. This report suggests that urban runoff to surface water and rural soil 

erosion to surface water can be significant reservoir sources. Releases from reservoirs 

(air, sediment, water, and biota) could not be reliably quantified, given the lack of 

information in this area. 

•	 The amount of dioxin-like PCBs released from man-made sources remains poorly 

characterized. Only a total of 19.5, 78.5, and 51.5 g of PCB TEQ could be quantified 

for the years 2000, 1995, and 1987, respectively. To date, only sewage sludge has 

been adequately characterized in terms of the amount of dioxin-like PCBs that may be 

released from a source. 

1.3.2. Time Trends 

A significant reduction in total CDD/CDF environmental releases has occurred since 

1987. EPA’s best estimates of releases of CDDs/CDFs to air, water, and land from reasonably 

quantifiable sources (Categories A, B, and C) are approximately 1,422 g TEQDF-WHO98 in 

reference year 2000; 3,444.2 g in reference year 1995; and 13,965.3 g in reference year 1987. 

From 1987 to 2000 there was an approximately 90% reduction in releases to all media. Most of 

the reduction in dioxin releases (75%) occurred between 1987 and 1995. 

In 1987 and 1995, municipal waste combustion was the leading source of dioxin 

emissions to the U.S. environment; however, because of reductions in dioxin emissions from 

MWCs, it dropped to the fourth ranked source in 2000. Burning of domestic refuse in backyard 

burn barrels remained fairly constant over the years, but in 2000 it emerged as the largest source 

of dioxin emissions. 

Reductions in environmental releases of dioxin-like compounds are attributed primarily 

to reductions in air emissions from MWCs, MWIs, and cement kilns burning hazardous waste, 

and in wastewater discharged into surface waters from pulp and paper mills using chlorine. 

These reductions have occurred from a combination of regulatory activities (see Section 1.1.2), 

improved emission controls, improved industrial technologies, voluntary actions on the part of 

industry, and the closing of a number of antiquated facilities. 

1.3.3. Sources Not Included in the Inventory 

Significant amounts of the dioxin-like compounds produced annually in the United States 

are not considered releases to the open and circulating environment and are not included in the 

national inventory. Examples include dioxin-like compounds generated internal to a process but 

destroyed before release and waste streams that are disposed of in approved landfills. 

1-42
 
The only product judged to have the potential for environmental release—and therefore 

considered for the inventory—was the herbicide 2,4-D. Release estimates are provided for 1987 

and 1995. Since 1995, the chemical manufacturers of 2,4-D have been undertaking voluntary 

actions to significantly reduce the dioxin content of the product. No information is available on 

the extent of these reductions and, therefore, no release estimate could be made for 2000. 

Regarding other products, data are presented on the amounts of CDDs/CDFs contained in 

bleached pulp, ethylene dichloride/vinyl chloride, PCP-treated wood, and dioxazine dyes and 

pigments. None of these products, however, was considered to have release potential; they were 

not included in the inventory. 

A number of contemporary formation sources were classified as D or E and therefore 

were not included in the inventory. The largest contemporary formation Category D sources are 

forest fires and accidental fires at MSW landfills. Taken together, these sources have the 

potential to significantly increase the present inventory if preliminary release estimates are 

confirmed. 

The possibility remains that truly undiscovered sources exist. Many of the sources that 

are well-accepted today were discovered only in the past 20 years. For example, CDDs/CDFs in 

stack emissions from MWCs were not detected until the late 1970s; CDDs/CDFs in the 

wastewater effluent from bleached pulp and paper mills were found unexpectedly in the mid1980s; iron ore was not recognized as a source until the early 1990s. 

1.3.4. Formation Theory 

Current theory proposes that CDDs/CDFs are formed within the cool-down region of 

combustion processes, either de novo or from dioxin precursors. De novo synthesis involves 

solid-phase reactions with carbon, chlorine, and oxygen on combustion-generated particles 

promoted by copper chloride as a catalyst. A less efficient but plausible formation process is the 

gas-phase formation from precursors catalyzed by the presence of a transition metal such as 

copper chloride. The ideal temperatures for de novo dioxin formation are between 200 and 

400°C. Reducing temperatures to below 200°C, especially at the air pollution control device, 

will minimize dioxin formation and releases from combustion sources. Chlorine sources present 

in feeds are necessary for dioxin formation. Experiments suggest that a chlorine content of 1% in 

the feed/fuel is the threshold for a direct relationship to dioxin formation from combustion 

sources, i.e., a chlorine content $1% is strongly correlated to the amount of dioxin formed, but a 

chlorine content <1% is not. However, in well-designed, well-controlled, and well-operated fullscale combustion systems there does not appear to be a direct relationship between the amount of 

chlorine present in the waste and the amount of dioxin emissions from the stack. 

1-43
 
Controversy exists regarding the role of PVC in the formation of CDDs/CDFs during 

municipal waste combustion. Experimental evidence suggests that PVC combustion generates 

hydrogen chloride gas (HCl) and dioxin precursors such as chlorobenzenes and chlorophenols, 

both of which may contribute to dioxin formation. HCl is a progenitor of chlorine radicals that 

then participate in the dioxin formation chemistry. Precursors are foundation molecules to dioxin 

formation. If PVC is the only source of chlorine and dioxin precursors during the combustion of 

MSW, then the removal of PVC may reduce the amount of dioxin formed and emitted. 

However, the complex mixture of materials in MSW provides sufficient chlorine for de novo 

synthesis, and dioxin precursors are formed as products of the incomplete combustion of the 

waste constituents. Therefore, the elimination of PVC from the waste prior to combustion would 

not necessarily eliminate the formation and emissions of CDDs/CDFs from municipal waste 

combustion. 

Current information strongly suggests that releases of CDDs/CDFs to the U.S. 

environment occur principally from anthropogenic activities. However, scientific studies have 

identified the possibility of natural formation of some CDDs/CDFs (e.g., in ball clay). 

1.3.5. Congener Profiles of CDD/CDF Sources 

This document presents congener profiles for a number of sources, as shown in Figure 1­

10. These profiles show the relative amounts of CDD/CDF congeners in environmental releases. 

These profiles can be useful for (1) identifying source contributions to near-field air 

measurements of CDDs/CDFs, (2) comparing sources, and (3) providing insights into the 

formation of CDDs/CDFs in the releases. There are numerous procedures for deriving a 

congener profile, and there is no single agreed-upon convention (Cleverly et al., 1997; Lorber et 

al., 1996; Hagenmaier et al., 1994). 

For this report, congener profiles were developed primarily by calculating the ratio of 

specific 2,3,7,8-substituted CDDs/CDFs in the emissions or product to the total (Cl4 - Cl8) 

CDDs/CDFs. With respect to combustion sources, the profiles were derived by dividing the 

congener-specific emission factors by the total (Cl4 - Cl8) CDD/CDF emission factor for each 

tested facility and then averaging the congener profiles developed for all tested facilities within 

the combustor type. For chemical processes and commercial chemicals, CDD/CDF profiles were 

typically generated by dividing average congener concentrations (ppt) in the chemical by the total 

CDDs/CDFs present. Profiles for select source categories are presented in Figure 1-10. 

1-44
 

4D 2378 

5D 12378

6D 1234786D123678

6D123789

7D 1234678

8D 12346789 12346789

4F 2378

5F123785F 23478

6F 1234786F 123678 6F 1237896F234678

7F12346787F1234789

8F 12346789

4D 2378

5D 12378 

6D 123478

6D 123678 

6D 123789 

7D 1234678

8D 12346789 

4F 2378 

5F 12378 5F 23478

6F 123478

6F 123678

6F 123789

6F 234678

7F 12346787F 1234789 

8F 12346789

4D2378

5D 12378

6D 123478 6D 123678

6D 123789 

7D 1234678

8D 12346789

4F 2378

5F 12378 5F 23478

6F 123478

6F 1236786F 123789

6F 234678 

7F 1234678 

7F1234789

8F 12346789

4D 2378

5D 12378 

6D 123478

6D 123678

6D 123789

7D1234678 

8D 12346789

4F 2378 

5F 12378

5F 23478

6F 123478

6F 123678

6F 123789 

6F 234678

7F 1234678

7F 1234789 

8F 12346789 

-

0% 

2% 

4% 

6% 

8% 

10 % 

12 % 

0% 

2% 

4% 

6% 

8% 

10% 

12% 

14% 

16% 

-

0%

2%

4%

6%

8%

10%

12%

0%

2%

4%

6%

8%

10%

12%

14%

16%

Municipal Waste Combustor with Hot ESP Municipal Waste Combustor with DS/FF 

0% 

1% 

2% 

3% 

4% 

5% 

6% 

7% 

8% 

0% 

2% 

4% 

6% 

8% 

10% 

12% 

0%

1%

2%

3%

4%

5%

6%

7%

8%

Medical/hospital Waste Incinerators Hazardous Waste Incinerators

0%

2%

4%

6%

8%

10%

12%

Figure 1-10. Congener profiles (as percent distributions to the sum of 

CDDs and CDFs) of anthropogenic sources of chlorinated dibenzo-pdioxins and chlorinated dibenzofurans in the United States. 

On the basis of inspection and comparisons of the average CDD/CDF congener profiles 

across combustion and noncombustion sources, the following observations were made (Cleverly 

et al., 1997) (these generalizations are derived from this data set, and their application beyond 

these data is uncertain): 

•	 It appears that combustion sources emit all 2,3,7,8-substituted CDDs/CDFs, although 

in varying percentages of total CDDs/CDFs. 

•	 In combustion source emissions, 2,3,7,8-TCDD is usually 0.1 to 1% of total 

CDDs/CDFs. The exception is stack emissions from industrial oil-fired boilers, 

where the available but limited data indicate that 2,3,7,8-TCDD constitutes an 

average of 7% of total CDD/CDF emissions. 

1-45
 
4D 2378

5D 12378

6D123478

6D123678 

6D 123789 

7D 1234678

8D 12346789

4F 2378

5F 12378

5F 23478

6F 1234786F 1236786F 1237896F 234678

7F 12346787F 1234789

8F 12346789 

4D 2378

5D12378

6D 123478 6D 1236786D 123789 

7D 1234678 

8D 12346789 

4F 2378

5F 12378 5F23478

6F 123478 6F 1236786F 1237896F 234678

7F 12346787F 1234789

8F 12346789 

4D 2378

5D 12378 

6D 123478 6D 123678 6D 123789

7D1234678 

8D 12346789

4F 2378

5F12378

5F 23478

6F 123478

6F 123678

6F 1237896F 234678

7F 1234678 

7F1234789

8F 12346789

4D 2378

5D 12378 

6D 1234786D 1236786D123789

7D1234678 

8D 12346789 12346789

4F2378

5F 12378 5F23478 

6F 1234786F 123678

6F 1237896F234678

7F12346787F 1234789 

8F 12346789

0% 

1% 

1% 

2% 

2% 

3% 

3% 

0% 

5% 

10 % 

15 % 

20 % 

25 % 

0%

1%

1%

2%

2%

3%

3%

Backyard Refuse Barrel Burning Industrial Wood Waste Combustion 

0%

5%

10%

15%

20%

25%

Portland Cement Kilns	 Portland Cement Kilns Burning Hazardous Waste 

5%

4%

30%

25%

4%

20% 3%

3%

15%

2%

10% 2%

1%

5% 1%

0% 0%

Figure 1-10. Congener profiles (as percent distributions to the sum of CDDs and 

CDFs) of anthropogenic sources of chlorinated dibenzo-p-dioxins and chlorinated 

dibenzofurans in the United States (continued). 

•	 It cannot be concluded that OCDD is the dominant congener for all combustiongenerated emissions of CDDs/CDFs. OCDD dominates total emissions from massburn MWCs that have DSs and FFs for dioxin control, industrial oil-fired boilers, 

industrial wood-fired boilers, unleaded gasoline combustion, diesel fuel combustion 

in trucks, and sewage sludge incinerators. The dominant congeners for other 

combustion sources are 1,2,3,4,6,7,8-HpCDF in emissions from mass-burn MWCs 

equipped with hot-sided electrostatic precipitators (ESPs), hazardous waste 

incineration, and secondary aluminum smelters and 2,4-D salts and esters; OCDF in 

emissions from medical waste incineration and industrial/utility coal-fired boilers; 

2,3,4,7,8-PeCDF in cement kilns burning hazardous waste; and 2,3,7,8-TCDF in 

cement kilns not burning hazardous waste. 

1-46
 

4D 2378

5D 12378 

6D 1234786D 123678

6D 123789

7D 1234678 

8D 12346789

4F 2378

5F 12378 5F23478

6F 1234786F 1236786F 123789 

6F 234678

7F 1234678

7F 1234789

8F 12346789

4D 2378 

5D 12378 

6D 123478 

6D 123678 

6D 123789

7D 1234678

8D 12346789

4F 2378

5F 12378 

5F 23478

6F1234786F 123678 

6F 123789

6F 234678

7F 1234678 7F 1234789

8F 12346789 

4D 2378

5D 12378

6D 123478 

6D 1236786D 123789 

7D 1234678

4F 2378

5F 12378

5F 23478

6F 123478 

6F 1236786F 123789

6F 234678 

7F 12346787F 1234789

8D 12346789 

8F 12346789

4D 2378

5D 12378

6D 123478 

6D 123678

6D 123789

7D 1234678

8D 12346789

4F 2378 

5F 12378 

5F 23478

6F1234786F 123678 6F 1237896F 234678

7F 1234678

7F 1234789

8F 12346789 

0% 

5% 

10% 

15% 

20% 

25% 

-

0% 

10 % 

20 % 

30 % 

40 % 

50 % 

60 % 

70 % 

0%

5%

10%

15%

20%

25%

Coal-fired Electrical Generating Facilities Oil-fired Electrical Generating Facilities 

0%

10%

20%

30%

40%

50%

60%

70%

0% 

5% 

10% 

15% 

20% 

25% 

30% 

35% 

40% 

45% 

0% 

10 % 

20 % 

30 % 

40 % 

50 % 

60 % 

70 % 

80 % 

0%

5%

10%

15%

20%

25%

30%

35%

40%

45%

Petroleum Refineries Bleached Pulp and Paper Mills (effluent) 

0%

10%

20%

30%

40%

50%

60%

70%

80%

Figure 1-10. Congener profiles (as percent distributions to the sum of CDDs and 

CDFs) of anthropogenic sources of chlorinated dibenzo-p-dioxins and chlorinated 

dibenzofurans in the United States (continued). 

Evidence for a shift in the congener patterns potentially caused by the application of 

different air pollution control systems within a combustion source type can be seen in 

the case of mass-burn MWCs. For mass-burn MWCs equipped with hot-sided ESPs, 

the most prevalent CDD/CDF congeners are 1,2,3,4,6,7,8-HxCDF; OCDD; 

1,2,3,4,6,7,8-HpCDD/1,2,3,4,7,8-HxCDF; 2,3,4,6,7,8-HxCDF/OCDF; 1,2,3,6,7,8­

HxCDF. The most prevalent congeners emitted from MWCs equipped with DS/FF 

are OCDD; 1,2,3,4,6,7,8-HpCDD; 1,2,3,4,6,7,8-HpCDF; OCDF; and 2,3,7,8­

TCDF/1,2,3,4,7,8-HxCDD; 2,3,4,6,7,8-HxCDF. 

•	 There is evidence of marked differences in the distribution of CDD/CDF congeners 

between cement kilns that burn hazardous waste and those that do not. When not 

burning hazardous waste as supplemental fuel, the dominant congeners appear to be 

2,3,7,8-TCDF; OCDD; 1,2,3,4,6,7,8-HpCDD, and OCDF. When burning hazardous 

waste, the dominant congeners are 2,3,7,8-PeCDF; 2,3,7,8-TCDF; 1,2,3,4,7,8­

1-47
 

4D2378

5D 12378

6D 123478

6D 123678

6D 123789

7D1234678

8D 12346789

4F 2378

5F 12378

5F23478

6F 123478

6F123678

6F 1237896F234678

7F 1234678

7F 1234789

8F12346789

4D 2378 

5D 12378

6D 123478

6D 123678

6D 123789

7D 1234678 

8D12346789

4F 2378 

5F 123785F 23478

6F123478 

6F123678

6F 123789

6F 234678

7F 12346787F 1234789

8F 12346789

4D 2378 

5D 12378

6D1234786D 123678 6D 123789 

7D 1234678

8D 12346789

4F 2378

5F 12378

5F 23478

6F 123478

6F 1236786F 1237896F 234678

7F 1234678

7F 1234789 

8F 12346789

4D 2378

5D 12378 

6D 123478 6D 123678

6D 123789

7D 1234678 

8D 12346789

4F 2378 

5F 123785F 23478

6F 123478

6F 123678 

6F 123789 

6F 234678 

7F 1234678 

7F 1234789

8F 12346789 

Secondary Aluminum Smelters Secondary Copper Smelters 

14%

12%

10%

8%

6%

4%

2%

0%

5%

5%

4%

4%

3%

3%

2%

2%

1%

1%

0%

Secondary Lead Smelters	 Iron Foundries 

3.0000% 0.07%

0.06% 2.5000%

0.05%

2.0000%

0.04%

1.5000%

0.03%

1.0000% 0.02%

0.5000% 0.01%

0.00% 0.0000%

Figure 1-10. Congener profiles (as percent distributions to the sum of CDDs and 

CDFs) of anthropogenic sources of chlorinated dibenzo-p-dioxins and chlorinated 

dibenzofurans in the United States (continued). 

xCDF; and 1,2,3,4,6,7,8-HpCDD. When burning hazardous waste, OCDD and 

OCDF are minor constituents of stack emissions. 

•	 The congener profile of 2,4-D salts and esters seems to mimic a combustion source 

profile in the number of congeners represented and in the minimal amount of 2,3,7,8­

TCDD relative to all 2,3,7,8-substituted congeners. A major difference is the 

prevalence of 1,2,3,7,8-PeCDD in 2,4-D (14%), which is not seen in any other 

combustion or noncombustion source presented here. 

1-48
 

4D 2378 

5D 12378 

6D 123478 

6D 123678 

6D 123789 

7D 1234678 

8D 12346789 

4F 2378 

5F 12378 

5F 23478 

6F 123478 6F 123678 6F 123789 

6F 234678 

7F 1234678 

7F 1234789 

8F 12346789 

4D 2378

5D 12378

6D 123478

6D 123678 

6D 123789

7D 1234678

8D 12346789

4F 2378

5F 123785F 23478 

6F 123478 

6F 123678

6F 123789 6F234678

7F1234678

7F 1234789 

8F 12346789

4D 2378

5D 12378

6D 123478

6D 123678 6D 123789

7D 1234678

8D 12346789

4F 2378

5F 123785F 23478 

6F 123478 

6F 123678

6F 123789 6F234678

7F 1234678

7F 1234789

8F 12346789

4D 2378

5D 12378

6D 123478

6D 123678 

6D 123789

7D 1234678

8D 12346789 

4F 2378

5F 123785F 23478 

6F 123478 

6F 123678

6F 123789 6F234678

7F1234678

7F 1234789

8F 12346789

•	 There are similarities in the congener profiles of PCP, diesel truck emissions, 

unleaded gasoline vehicle emissions, and emissions from industrial wood combustors. 

In these sources, OCDD dominates total emissions, but the relative ratio of 

1,2,3,4,6,7,8-HpCDD to OCDD is also quite similar. 

•	 The congener profiles for diesel truck exhaust and those for air measurements from a 

tunnel study of diesel traffic are quite similar. 

Automobiles Burning Leaded Gasoline	 Diesel Truck Exhaust 

14% 45%

40% 12%

35%

10%

30%

8% 25%

6% 20%

15% 4%

10%

2%

5%

0% 0%

0% 

5% 

10% 

15% 

20% 

25% 

30% 

35% 

0% 

2% 

4% 

6% 

8% 

10% 

12% 

0%

5%

10%

15%

20%

25%

30%

35%

Automobiles Burning UnLeaded Gasoline Crematoria 

0%

2%

4%

6%

8%

10%

12%

Figure 1-10. Congener profiles (as percent distributions to the sum of CDDs 

and CDFs) of anthropogenic sources of chlorinated dibenzo-p-dioxins and 

chlorinated dibenzofurans in the United States (continued). 

1-49
 

4D2378

5D 12378

6D 1234786D 1236786D 123789

7D 1234678

8D 12346789

4F2378

5F12378

5F 23478

6F 1234786F 123678 6F123789

6F 234678

7F12346787F 1234789

8F 12346789

0% 

10% 

20% 

30% 

40% 

50% 

60% 

70% 

80% 

0%

10%

20%

30%

40%

50%

60%

70%

80%

Forest Fires 

Figure 1-10. Congener profiles (as percent distributions to the sum of CDDs 

and CDFs) of anthropogenic sources of chlorinated dibenzo-p-dioxins and 

chlorinated dibenzofurans in the United States (continued). 

1-50
 

2. MECHANISMS OF FORMATION OF DIOXIN-LIKE COMPOUNDS 

DURING COMBUSTION OF ORGANIC MATERIALS 

More than a decade of combustion research has contributed to a general understanding of 

the central molecular mechanisms that form CDDs/CDFs emitted from combustion sources. 

Current understanding of the conditions necessary to form CDDs/CDFs were derived primarily 

from studies of full-scale municipal waste combustors (MWCs), augmented with observations 

involving the experimental combustion of synthetic fuels and feeds in the laboratory. However, 

the formation mechanisms elucidated by these studies are generally relevant to most combustion 

systems in which organic material is burned with chlorine. 

Intensive studies have examined MWCs from the perspective of identifying the specific 

formation mechanism(s) that occurs within the system. This knowledge may lead to methods 

that prevent the formation of CDDs/CDFs and their release into the environment. Although 

much has been learned from such studies, a method that completely prevents CDDs/CDFs from 

forming during the combustion of certain organic materials in the presence of a source of 

chlorine and oxygen is still unknown. The wide variability of organic materials incinerated and 

thermally processed by a wide range of combustion technologies that have varying temperatures, 

residence times, and oxygen requirements adds to this complex problem. However, central 

chemical events involved in the formation of CDDs/CDFs can be identified by evaluating 

emission test results from MWCs in combination with results from laboratory experiments. 

CDD/CDF emissions from combustion sources can potentially be explained by three 

principal mechanisms that should not be regarded as being mutually exclusive. In the first 

mechanism (referred to as “pass through”), CDDs/CDFs are present as contaminants in the 

combusted organic material; they pass through the furnace and are emitted unaltered. This 

mechanism is discussed in Section 2.1. In the second mechanism (referred to as “precursor”), 

CDDs/CDFs ultimately form from the thermal breakdown and molecular rearrangement of 

precursor ring compounds, which are defined as chlorinated aromatic hydrocarbons that have a 

structural resemblance to the CDD/CDF molecules. Ringed precursors that emanate from the 

combustion zone are a result of the incomplete oxidation of the constituents of the feed (i.e., 

products of incomplete combustion). The precursor mechanism is discussed in Section 2.2. The 

third mechanism (referred to as “de novo synthesis”) is similar to the precursor mechanism and is 

described in Section 2.3. De novo synthesis describes a pathway of CDD/CDF formation from 

heterogeneous reactions on fly ash (particulate matter [PM]) involving carbon, oxygen, hydrogen, 

chlorine, and a transition metal catalyst. With these reactions, intermediate compounds that have 

an aromatic ring structure are formed. 

2-1
 
Studies in this area suggest that aliphatic compounds, which arise as products of 

incomplete combustion, may play a critical role in initially forming simple ring molecules, which 

later evolve into complex aromatic precursors. CDDs/CDFs are then formed from the 

intermediate compounds. In both the second and the third mechanism, formation occurs outside 

the furnace, in the so-called post-combustion zone. Particulate-bound carbon is suggested as the 

primary reagent in the de novo synthesis pathway. 

Section 2.4 presents an overview of studies that have investigated the role that chlorine 

plays in forming CDDs/CDFs. Although chlorine is an essential component for the formation of 

CDDs/CDFs in combustion systems, the empirical evidence indicates that for commercial-scale 

incinerators, chlorine levels in feed are not the dominant controlling factor for rates of CDD/CDF 

stack emissions. There are complexities related to the combustion process itself, and some types 

of air pollution control equipment tend to mask any direct association. Therefore, the chlorine 

content of fuel and feeds to a combustion source is not a good indicator of levels of CDDs/CDFs 

emitted from the stack of that source. 

Section 2.6 discusses the generation and formation of coplanar polychlorinated biphenyls 

(PCBs). The presence of coplanar PCBs in stack emissions from combustors is an area in need 

of further research. Evidence to date suggests that PCB emissions are mostly attributable to PCB 

contamination in waste feeds and that emissions are related to the first mechanism described 

above. However, newly published research has also indicated that it is possible that PCBs form 

in much the same way as described in the second and third mechanisms identified in the 

formation of CDDs/CDFs within the post-combustion zone. 

Section 2.7 provides a closing summary of the three principal formation mechanisms and 

the role of chlorine. From the discussions in this chapter, it should be evident that no clear 

distinction exists between the precursor and the de novo synthesis mechanisms of CDD/CDF 

formation. Both formation pathways depend on the evolution of precursors within combustion 

gases, the interaction of reactive fly ashes, a generally oxidative environment, the presence of a 

transition metal catalyst, the presence of gaseous chlorine, and a favorable range of temperatures. 

The temperature of the combustion gases (i.e., flue gases) is perhaps the single most important 

factor in forming dioxin-like compounds. Temperatures between 200 and 450°C are most 

conducive to the formation of CDDs/CDFs, with maximum formation occurring at around 

350°C. If the temperature falls outside this range, the amount of CDDs/CDFs formed is 

minimized. 

2-2
 
2.1. MECHANISM 1 (PASS THROUGH): CDD/CDF CONTAMINATION IN FUEL AS 

A SOURCE OF COMBUSTION STACK EMISSIONS 

The first mechanism involved in stack emissions of CDDs/CDFs is the incomplete 

destruction of CDD/CDF contaminants present in the fuel or feeds delivered to the combustion 

chamber. Not all of these molecules are destroyed by the combustion system, thus allowing 

trace amounts to be emitted from the stack. Most work in this area has involved the study of 

incineration of municipal solid waste (MSW), where CDDs/CDFs were analytically measured in 

the raw refuse fed into the incinerator. CDDs/CDFs are ubiquitous in the environment (air, 

water, and soil) and in foods and paper; therefore, they clearly are present in municipal waste 

(Tosine et al., 1983; Ozvacic, 1985; Clement et al., 1988; Federal Register, 1991a; Abad et al., 

2002). 

Abad et al. (2002) provided contemporary measurements of CDDs/CDFs in raw MSW. 

Twenty-two samples were collected and analyzed for CDDs/CDFs over a 1-year period, from 

September 1998 through September 1999. The congeners that dominated the total mass of 

CDDs/CDFs were OCDD and 1,2,3,4,7,8,9-HpCDD. Figure 2-1 displays the mean CDD and 

CDF congener distribution from this study. Abad et al. found that the I-TEQ concentration in the 

MSW was highly variable and ranged from 1.55 to 45.16 ng I-TEQ/kg MSW. 

350 

300 m

a gr

250 r pe

200 

s

m

a

150 

ogr

ci

100 

P

50 

0 

2,3,7,8-TCDD

1,2,3,7,8-P

eCDD

1,2,3,4,7,8-HxCD1,2,3,6,7,8-HxCD1,2,3,7,8,9-HxCD

1,2,3,4,6,7,8-Hp

OCDD

2,3,7,8-TCDF

1,2,3,7,8-P

eCDF

2,3,4,7,8-P

eCDF

1,2,3,4,7,8-HxCD1,2,3,6,7,8-HxCD1,2,3,7,8,9-HxCD2,3,4,6,7,8-HxCD

1,2,3,4,6,7,8-Hp1,2,3,4,7,8,9-Hp

OCDF 

Congener 

Figure 2-1. Typical mean distribution of CDD and CDF congeners in 

contemporary municipal solid waste. 

Source: Adapted from Abad et al. (2002). 

2-3 
A number of studies have provided evidence that most of the CDDs/CDFs present in 

MSW are destroyed during combustion (Abad et al., 2002; Clement et al., 1988; Commoner et 

al., 1984, 1985, 1987; Hay et al., 1986; Environment Canada, 1985). These studies involved a 

mass balance of the input versus output of CDDs/CDFs at two operational MWCs. The mass of 

CDDs/CDFs outside the incinerator furnace was found to be much greater than the mass of 

CDDs/CDFs in the raw MSW fed into the incinerator, and the profiles of the distributions of 

CDD/CDF congeners were strikingly different. Primarily, the more highly chlorinated congeners 

were detected as contaminants in the waste, whereas the total array of tetra- through octaCDDs/CDFs could be detected in the stack gases. Moreover, the ratio of the total CDD 

concentration to the total CDF concentration in the MSW was greater than 1, whereas in typical 

incinerator stack emissions this ratio is less than 1 (meaning more dibenzofurans than dioxins are 

emitted). From such evidence it can be concluded that CDDs/CDFs are being synthesized after 

the contaminated feed has been combusted (Abad et al., 2002). It is also expected that the 

conditions of thermal stress imposed by high temperatures reached in typical combustion would 

destroy and reduce the CDDs/CDFs present as contaminants in the waste feed to levels that are 

0.0001 to 10% 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%, but lower 

temperatures result in a higher survival rate. 

Theoretical modeling has shown that unimolecular destruction of CDDs/CDFs at 99.99% 

can occur at the following temperatures and retention times within the combustion zone: 977°C 

with a retention time of 1 sec, 1,000°C at a retention time of 0.5 sec, 1,227°C at a retention time 

of 4 msec, and 1,727°C at a retention time of 5 :sec (Schaub and Tsang, 1983). Thus, 

CDDs/CDFs would have to be in concentrations of parts per million in the feed in the combustor 

to be found in the parts-per-billion or parts-per-trillion level in the stack gas emissions (Shaub 

and Tsang, 1983). However, it cannot be ruled out that CDDs/CDFs in the waste or fuel may 

contribute (up to some percentage) to the overall concentration leaving the stack. The only other 

possible explanation for CDD/CDF emissions from high-temperature combustion of organic 

material is formation outside and downstream of the furnace. 

The above studies point to formation mechanisms other than simple pass through of 

noncombusted feed contamination. These formation mechanisms are discussed and reviewed in 

the following sections. 

2.2. MECHANISM 2 (PRECURSOR): FORMATION OF CDDs/CDFs FROM 

PRECURSOR COMPOUNDS 

The second mechanism involves the formation of CDDs/CDFs from aromatic precursor 

compounds in the presence of a chlorine donor. This mechanism has been elucidated by 

2-4
 
laboratory experiments involving the combustion of known precursors in quartz ampules under 

starved-air conditions and in experiments that investigated the role of combustion fly ash in 

promoting the formation of CDDs/CDFs from precursor compounds. The general reaction in this 

formation pathway is an interaction between an aromatic precursor compound and chlorine 

promoted by a transition metal catalyst on a reactive fly ash surface (Stanmore, 2004; Dickson 

and Karasek, 1987; Liberti and Brocco, 1982). Examples of well-studied precursor compounds 

include chlorobenzenes, chlorophenols, phenol, and benzene (Esposito et al., 1980). Gaseous 

hydrogen chloride (HCl), free chlorine (Cl2), and chlorine radicals (ClA) are the chlorinating 

agents within the combustion gases. CDD/CDF formation results from heterogeneous gas-phase 

reactions involving chlorinated precursor compounds and a source of chlorine. Chlorophenol 

and chlorobenzene compounds have been measured in flue gases from MWCs (Dickson and 

Karasek, 1987). 

Precursors are carried from the furnace to the flue duct as products of incomplete 

combustion. These compounds can adsorb on the surface of combustion fly ash or entrain in the 

gas phase within the flue gases. Thus, there are two formation pathways from precursor 

compounds: heterogeneous solid-phase reactions and homogeneous gas-phase reactions. In the 

post-combustion region outside the furnace, heterogeneous reactions on the surface of reactive 

fly ash can ensue to form CDDs/CDFs from the precursor compounds. This occurs at the cooldown temperatures of 200 to 400°C. The heterogeneous gas-phase reactions occur from the 

breakdown and molecular rearrangement of precursor compounds followed by condensation and 

chlorination at the higher temperatures of 500 to 800°C. Both reaction pathways are catalyzed by 

copper chloride (CuCl2) or another transition metal. 

Laboratory experiments involving the controlled combustion of precursor compounds 

have caused the breakdown of the precursor reagent and the subsequent appearance of 

CDDs/CDFs as products of the reaction. For example, Jansson et al. (1977) produced CDDs 

through the pyrolysis of wood chips treated with tri-, tetra-, and pentachlorophenol (PCP) in a 

bench-scale furnace operated at 500 to 600°C. Stehl and Lamparski (1977) combusted grass and 

paper treated with the herbicide 2,4,5-trichlorophenoxyacetic acid in a bench-scale furnace at 600 

to 800°C and generated ppmv levels of TCDD. Ahling and Lindskog (1982) reported CDD 

formation during the combustion of tri- and tetrachlorophenol formulations at temperatures of 

500 to 600°C. Decreases in oxygen during combustion generally increased the CDD yield. 

Ahling and Lindskog (1982) noted that adding copper salts to the tetrachlorophenol 

formulation significantly enhanced the yield of CDDs. This may have been an early indication 

of copper’s role in catalyzing the condensation of chlorophenol to dioxin. Combustion of PCP 

resulted in low yields of CDDs. However, when PCP was burned with an insufficient supply of 

oxygen in the presence of copper, the investigators noted the formation of tetra- through 

2-5
 
octachlorinated congeners. Buser (1979) generated CDDs/CDFs on the order of 0.001 to 0.08% 

(by weight) by heating tri-, tetra-, and pentachlorobenzenes at 620°C in quartz ampules in the 

presence of oxygen. It was noted that chlorophenols formed as combustion by-products; Buser 

speculated that these chlorophenols were acting as reaction intermediates in the formation of 

CDDs/CDFs. 

The second condition postulated to regulate the synthesis of CDDs/CDFs from the 

aromatic precursor compound is the adsorption and interaction with the reactive surface of 

combustion-generated fly ash (PM) entrained in the combustion plasma and the presence of a 

transition metal catalyst (Stanmore, 2004; Dickson et al., 1992; Bruce et al., 1991; Cleverly et al., 

1991; Gullet et al., 1990a; Commoner et al., 1987; Dickson and Karasek, 1987; Vogg et al., 

1987). These are heterogeneous solid-phase reactions that occur at temperatures below 450°C. 

The molecular precursor leaves the gas phase and condenses onto the fly ash particle. This 

condition, which places greater emphasis on heterogeneous surface reactions and less emphasis 

on homogeneous gas-phase reactions, was first postulated by Shaub and Tsang (1983) using 

thermal-kinetic models based on the temperature of the heat of formation, adsorption, and 

desorption. Shaub and Tsang modeled CDD production from chlorophenols and concluded that 

solid-phase formation of CDDs/CDFs was of greater importance than gas-phase formation within 

an incineration system. 

The temperature of the combustion gases is a critical factor in the formation of 

CDDs/CDFs from aromatic precursor compounds (Weber and Hagenmaier, 1999; Fangmark et 

al., 1994; Vogg et al., 1987, 1992; Oberg et al.,1989). Vogg et al. (1987) found that formation 

probably occurs outside of and downstream from the combustion zone of a furnace, in regions 

where the temperature of the combustion offgases has cooled within a range of 200 to 450°C. 

After carefully removing organic contaminants from MWC fly ash, Vogg et al. (1987) 

added known concentrations of isotopically labeled CDDs/CDFs to the matrix. The MWC fly 

ash was then heated for 2 hr in a laboratory furnace at varying temperatures. The treated fly ash 

was exposed to temperatures increasing in 50°C increments within a temperature range of 150 to 

500°C. Table 2-1 summarizes these data. Because the relative concentration of CDDs/CDFs 

increased while exposed to varying temperatures, it was concluded that the temperature of the 

combustion gas is crucial to promoting the formation of CDDs/CDFs on the surface of fly ash. 

Within a temperature range of 200 to 450°C, the concentration of CDDs/CDFs increases to some 

maxima; outside this range, the concentration diminishes. 

2-6
 
Table 2-1. Concentration of CDDs/CDFs on municipal incinerator fly ash at 

varying temperatures 

Congener 

CDD/CDF concentration on fly ash (ng/g) by temperature 

200°C 250°C 300°C 350°C 400°C 

CDD 

Tetra 

Penta 

Hexa 

Hepta 

Octa 

CDF 

Tetra 

Penta 

Hexa 

Hepta 

Octa 

15 

40 

65 

100 

90 

122 

129 

61 

48 

12 

26 

110 

217 

208 

147 

560 

367 

236 

195 

74 

188 

517 

1,029 

1,103 

483 

1,379 

1,256 

944 

689 

171 

220 

590 

550 

430 

200 

1,185 

1,010 

680 

428 

72 

50 

135 

110 

60 

15 

530 

687 

260 

112 

12 

Source: Adapted from Vogg et al. (1987). 

The region of cooler gas temperature is often referred to as the “post-combustion region.” 

This region extends from near the exit of the furnace to the point of release of the combustion 

gases at stack tip. The heat loss may be inherent in the conduction and transfer through the 

combustion gas metal ducting system or related to the adsorption/exchange of heat to water in 

boiler tubes. 

Fangmark et al. (1994) found that CDDs/CDFs exhibit a similar dependence at a 

temperature range of 260 to 430°C, with maximum formation occurring around 340°C. Using a 

pilot-scale combustor, Behrooz and Altwicker (1996) found that the formation of CDDs/CDFs 

from the precursor 1,2-dichlorobenzene rapidly occurred within the post-combustion region in a 

temperature range of 390 to 400°C, with residence times of only 4 to 5 sec. On the other hand, 

CDD/CDF formation from 1,2-dichlorophenol seemed to require higher temperatures. 

Oberg et al. (1989) examined the role of temperature in the formation kinetics using a 

full-scale hazardous waste incinerator (HWI) operating in Sweden. The investigators observed 

that maximum CDD/CDF formation transpired in the boiler used to extract heat for cogeneration 

of energy. In this study, significant increases in total concentration of I-TEQDF occurred between 

280 and 400°C, and concentrations declined at temperatures above 400°C. Weber and 

Hagenmaier (1999) showed that in gas-phase reactions, chlorophenols react in the presence of 

2-7
 
oxygen at temperatures above 340°C to form CDDs/CDFs. Phenoxyradicals were formed, which 

in turn caused the formation of CDDs. Polychlorinated dihydroxybiphenyls were identified as 

reaction intermediates in the gas-phase dimerization of chlorophenols, and these intermediates 

could form CDFs. 

Konduri and Altwicker (1994) proposed that rate-limiting factors were the nature and the 

concentrations of the precursors, the reactivity and availability of the fly ash surface, and the 

residence time in the post-combustion zone. Dickson and Karasek (1987) investigated fly ash 

reactivity with 13C6-chlorophenol compounds. Several samples of fly ash from MWCs and 

copper smelters and a variety of combustion fuels were heated at 300°C in quartz tubes under 

conditions known to catalyze the conversion of chlorophenols to CDDs/CDFs. The MSW fly ash 

included a sample from a poorly operated mass burn refractory incinerator and a sample from a 

well-operated fluidized-bed combustor. The MWC fly ash proved to be the most active catalytic 

medium, despite similarities among the samples with respect to specific surface area and average 

pore diameter. The fly ash from the refractory MWC generated about seven times more mass of 

dioxin-like compounds than did the fluidized-bed MWC. In the MSW fly ashes, all CDD/CDF 

congener groups were formed from labeled chlorophenols; however, only trace amounts of 

heptachloro- and octachlorodioxin were formed with the copper smelter/refiner. X-ray 

photoelectron spectroscopy revealed the presence of chlorine adsorbed to the surface of the 

MWC fly ash but an absence of chlorine sorbed to the copper smelter fly ash. 

CDD congener groups have been postulated to form from the labeled PCP precursors by 

(1) first forming octachlorodioxin by the condensation of two PCP molecules, and (2) forming 

other less-chlorinated dioxins through dechlorination of the more highly chlorinated isomers. 

These steps seemed to proceed by an increased reactivity of the chemisorbed precursor molecule 

caused by the removal of one or more hydrogen or chlorine atoms along the ring structure 

(Dickson and Karasek, 1987), an observation consistent with the kinetic model of Shaub and 

Tsang (1983). 

In related experiments, Dickson and Karasek (1987) more specifically reported on 

forming CDDs/CDFs from condensation reactions of chlorophenols on the surface of MWC fly 

ash 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 whether MWC fly ash could promote these 

reactions. MWC fly ash was obtained from facilities in Canada and Japan. The fly ash was 

rinsed with solvent to remove any organic constituents prior to initiating the experiment. Twenty 

grams of fly ash were introduced into a bench-scale furnace (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 PCP and two trichlorophenol isotopes (13C12-2,3,5­

2-8
 
trichlorophenol and 3,4,5-trichlorophenol) were added to the surface of the clean fly ash matrix 

and placed in the oven for 1 hr at 300°C. Pure inert nitrogen gas (flow rate of 10 mL/min) was 

passed through the flow tube and a constant temperature was maintained. 

Tetra- through octa-CDDs were formed from the labeled PCP experiment; more than 100 

:g/g of total CDDs were produced. The congener pattern was similar to that found in MWC 

emissions. The 2,3,5-trichlorophenol experiment primarily produced HxCDDs and very small 

amounts of tetra- through octa-CDDs. The 3,4,5-trichlorophenol experiment mainly produced 

OCDD and 1,2,3,4,6,7,8-HpCDD. 

Dickson and Karasek (1987) proposed that CDDs on fly ash surfaces may result from 

chlorophenol undergoing molecular rearrangement or isomerization as a result of dechlorination, 

dehydrogenation, and transchlorination before condensation occurs. These reactions were 

proposed as controlling the types and amounts of CDDs that are ultimately formed. Born et al. 

(1993) conducted experiments on the oxidation of chlorophenols with fly ash in a quartz tube 

reactor heated to about 300°C. The MWC fly ash mediated the oxidation of chlorophenols to 

produce carbon dioxide (CO2) and carbon monoxide (CO) as major products and polychlorinated 

benzenes, monobenzofurans, and nonhalogenated dibenzo-p-dioxins as trace species. Formation 

of these trace aromatic species occurred after residence times of only 7 to 8 sec, which was 

consistent with the later experimental result of Behrooz and Altwicker (1995), which showed the 

potential for rapid formation from a precursor. 

Milligan and Altwicker (1996) fitted experimental flow-tube reactor data to classical 

catalytic reaction models to empirically explain the interaction of 2,3,4,6-tetrachlorophenol (as a 

model precursor) with reactive MWC fly ash during MSW incineration. The precursor was 

found to be highly adsorptive on the surface of fly ash, with a first-order dependence on gasphase precursor concentration to CDD formation. The investigators concluded that 

chlorophenol’s dependence on gas-phase concentration to form CDDs on fly ash reflects the 

highly heterogeneous nature of the fly ash surface. Moreover, the estimated 6 × 1018 adsorption 

sites per gramof fly ash suggested the presence of highly energetic sites, which may be important 

in the surface-catalyzed reactions forming CDDs. An interesting observation by Milligan and 

Altwicker was that precursor molecules appeared to compete with oxygen molecules for the 

reactive sites; therefore, chlorophenols are expected to adsorb less readily to the fly ash surface 

in the presence of oxygen. 

Experimental evidence suggests that condensation to CDD of chlorophenol compounds 

via isomerization and the Smiles rearrangement on reactive MWC fly ash surfaces is a proven 

pathway for the formation of dioxins from a precursor compound (Addink and Olie, 1995). 

However, no detailed mechanisms have been presented for CDD/CDF formation from other 

precursors such as chlorobenzenes under conditions simulating incineration. 

2-9
 
A condition in the synthesis of CDDs/CDFs from aromatic precursor compounds is that 

the presence of a transition metal catalyst promotes the chemical reaction on the surface of fly 

ash. CuCl2 is a strong catalyst for promoting surface reactions on PM to convert aromatic 

precursor compounds to CDDs/CDFs (Vogg et al., 1987). CuCl2 promotes ring condensation 

reactions (of the chlorophenols) on fly ash to form CDDs/CDFs (Addink and Olie, 1995) via the 

Ullman reaction (Born et al., 1993). In the Ullman reaction, copper catalyzes the formation of 

diphenyl ethers by the reaction of halogenated benzenes with alkali metal phenolates (Born et al., 

1993), with copper participating in a nucleophilic aromatic substitution reaction. Thus, Born et 

al. proposed a similar mechanism in catalyzing the formation of dioxin-like compounds. Using 

the Ullman reaction as a model, the authors proposed that the copper-catalyzed condensation of 

two ortho-substituted chlorophenol molecules form chlorine-free dibenzo-p-dioxins. 

Vogg et al. (1987) proposed an oxidation reaction pathway, giving rise to the formation of 

CDDs/CDFs in the post-furnace regions of the incinerator in the following order: (1) 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 CuCl2 as a 

catalyst, yields free gaseous chlorine via the Deacon reaction; (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 Cl2; and (4) a chlorinated precursor to dioxin 

(e.g., chlorophenol) is further oxidized (with CuCl2 as a catalyst) to yield CDDs/CDFs and 

chlorine. 

Gullett et al. (1990a, b, 1991a, b, 1992) studied the formation mechanisms through 

extensive combustion research at EPA and verified the observations of Vogg et al. (1987). It was 

proven that CDDs/CDFs could ultimately be produced from low-temperature (i.e., 350°C) 

reactions between chlorine (Cl) and a phenolic precursor combining to form a chlorinated 

precursor, followed by oxidation of the chlorinated precursors (catalyzed by a copper catalyst 

such as CuCl2), as shown below. 

1. The initial step in dioxin formation is the formation of chlorine from HCl in the 

presence of oxygen (the Deacon reaction), as follows (Bruce et al., 1991; Vogg et al., 1987):

 Heat 
 

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 of the breakdown and molecular 

2-10
 
rearrangement of the precursor. The reaction is promoted by CuCl2 acting as a catalyst (Vogg et 

al., 1987; Dickson and Karasek, 1987; Gullett et al., 1992): 

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

 CuCl2 

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

Eklund et al. (1986) observed the high-temperature formation of a large variety of 

chlorinated toxic compounds, including CDDs/CDFs, from precursors during a simple 

experiment in which phenol was oxidized with HCl at 550°C. One milligram of phenol was 

placed in a quartz tube reactor with an aqueous solution (10 :L) of HCl and heated at a 

temperature of 550°C for 5 min. Trichlorobenzene, dichlorophenol, dichlorobenzofuran, 

tetrachlorobenzene, trichlorophenol, and tetrachlorophenol were identified as major products 

formed. Monochlorobenzene, chlorophenol, dichlorobenzene, tetrachloropropene, 

pentachloropropene, trichlorobenzofuran, TCDF, TrCDD, TCDD, HxCDD, HxCDF, 

pentachlorobenzene, pentachlorobiphenyl, and pentachlorodihydroxycylohexane were observed 

as minor products. Trace species formed included MCDF, PeCDF, PeCDD, OCDF, and OCDD. 

Eklund et al. (1986) hypothesized that chlorinated organic compounds can be produced 

from phenols, acids, and any chlorine source in the hot post-combustion region (just beyond the 

exit to the furnace). The reaction was seen as very sensitive to HCl concentration. No 

chlorinated compounds could be detected when HCl concentrations were <10-3 mol. 

Nestrick et al. (1987) reported that the thermolytic reaction between benzene (an 

unsubstituted precursor) and iron (III) chloride on a silicate surface yielded CDDs/CDFs at 

temperatures $150°C. The experimental protocol introduced 100 to 700 mg benzene and 13C6­

benzene into a macroreactor 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 carried the benzene vapor to the furnace area. The exit from 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 ferric trichloride (FeCl3) was added to form an 

FeCl3/silica reagent. The thermolytic reaction took place in a temperature range of 150 to 400°C 

at a residence time of 20 min. Although di- through octa-CDDs/CDFs were formed by this 

reaction at all temperatures studied, the percent yields were extremely small. Table 2-2 

summarizes these data. 

2-11
 
Table 2-2. CDDs/CDFs formed from the thermolytic reaction of 690 mg 

benzene + FeCl3 silica complex 

Congener Mass produced (ng) Number of mols produced Percent yielda 

DiCDD 

TrCDD 

TCDD 

PeCDD 

HxCDD 

HpCDD 

OCDD

 Total CDDs 

4.9 

54.0 

130.0 

220.0 

170.0 

98.0 

20.0 

696.9 

0.019 

0.019 

0.400 

0.620 

0.440 

0.230 

0.040 

1.940 

4.3 e!7 

4.3 e!6 

9.0 e!6 

1.4 e!5 

9.9 e!6 

5.2 e!6 

9.0 e!7 

4.4 e!5 

DiCDF 

TriCDF 

TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF

 Total CDFs 

990 

7,800 

12,000 

20,000 

33,000 

40,000 

74,000 

187,000 

4.2 

29.0 

39.0 

59.0 

88.0 

98.0 

167.0 

484.2 

9.5 e!5 

6.6 e!4 

8.8 e!4 

1.3 e!3 

2.0 e!3 

1.1 e!3 

3.8 e!3 

1.1 e!2 a

 Number of mols of CDD or CDF/mols benzene x 100. 

FeCl3 = ferric chloride 

Source: Nestrick et al. (1987). 

2.3. MECHANISM 3 (DE NOVO SYNTHESIS): SYNTHESIS OF CDDs/CDFs DURING 

COMBUSTION OF ORGANIC MATERIALS 

The third mechanism promotes CDD/CDF formation in combustion processes from the 

oxidation of carbon particulate catalyzed by a transition metal in the presence of chlorine. As in 

the precursor mechanism (mechanism 2), synthesis is believed to occur in regions outside of the 

furnace zone of the combustion process, where the combustion gases have cooled to a range of 

temperatures considered favorable to formation chemistry. A key component to de novo 

synthesis is the production of intermediate compounds (either halogenated or nonhalogenated) 

that are precursors to CDD/CDF formation. Research in this area has produced CDDs/CDFs 

directly by heating carbonaceous fly ash in the presence of a transition metal catalyst without the 

apparent generation of reactive intermediates. Thus, the specific steps involved in the de novo 

process have not been fully and succinctly delineated. However, laboratory experimentation has 

proven that MWC fly ash itself is a reactive substrate, and the matrix can actually catalyze the de 

2-12
 
novo formation chemistry. Typically, fly ash is composed of an alumina-silicate construct, with 

5 to 10% concentrations of silicon, chlorine (as inorganic chlorides), sulfur, and potassium 

(NATO, 1988). Twenty percent of the weight of fly ash particles is carbon, and the particles 

have specific surface areas in the range of 200 to 400 m2 /kg (NATO, 1988). 

The de novo synthesis essentially is the oxidative breakdown of macromolecular carbon 

structures, and CDDs/CDFs are formed partially from the aromatic carbon-oxygen functional 

groups embedded in the carbon skeleton (Huang et al., 1999). The distinguishing feature of the 

de novo synthesis over the precursor synthesis is the oxidation of carbon in particulate at the start 

of the process to yield precursor compounds. In mechanism 2, the precursor compound is the 

starting molecule of the condensation reactions forming CDDs/CDFs (Dickson et al., 1992). By 

this distinction, however, one could argue that mechanism 3 is really an augmentation of 

mechanism 2 because the production of CDDs/CDFs may still require the formation of a 

CDD/CDF precursor as an intermediate species. Nevertheless, a distinction is presented here to 

describe additional pathways suggested for the thermal formation of these compounds. 

To delineate the de novo synthesis of CDDs/CDFs, Stieglitz et al. (1989) conducted 

experiments that involved heating particulate carbon containing adsorbed mixtures of 

magnesium-aluminum (Mg-Al) silicate in the presence of CuCl2 (as a catalyst to the reaction). 

The authors described heating mixtures of Mg-Al silicate with activated charcoal (4% by 

weight), chloride as potassium chloride (7% by weight), and CuCl2 (1% in water) in a quartz 

flow tube reactor at 300°C. The retention time was varied at 15 min, 30 min, and 1, 2, and 4 hr 

to obtain differences in the amounts of CDDs/CDFs that could be formed. The results are 

summarized in Table 2-3. In addition to the CDDs/CDFs formed as primary products of the de 

novo synthesis, the investigators observed precursors formed at the varying retention times 

during the experiment. In particular, similar yields of tri- through hexachlorobenzenes, trithrough heptachloronaphthalenes, and tetra- through heptachlorobiphenyls were quantified; this 

was seen as highly suggestive of the role these compounds may play as intermediates in the 

continued formation of CDDs/CDFs. 

Stieglitz et al. (1989) made the following observations: 

C	 The de novo synthesis of CDDs/CDFs via the oxidation of carbonaceous PM occurred 

at a temperature of 300°C. Additionally, the experiment yielded parts-per-billion to 

parts-per-million concentrations of chlorinated benzenes, chlorinated biphenyls, and 

chlorinated naphthalenes 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 

formed as reaction products. 

2-13
 
Table 2-3. De novo formation of CDDs/CDFs after heating Mg-Al silicate, 

4% charcoal, 7% Cl, 1% CuCl2 in H2O at 300°C 

Congener 

Concentration of CDD/CDF (ng/g) by reaction time (hr) 

0.25 0.5 1 2 4 

TCDD 

PeCDD 

HxCDD 

HpCDD 

OCDD

 Total CDDs 

2 

110 

730 

1,700 

800 

3,342 

4 

120 

780 

1,840 

1,000 

3,744 

14 

250 

1,600 

3,500 

2,000 

7,364 

30 

490 

2,200 

4,100 

2,250 

9,070 

100 

820 

3,800 

6,300 

6,000 

17,020 

TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF

 Total CDFs 

240 

1,360 

2,500 

3,000 

1,260 

8,360 

280 

1,670 

3,350 

3,600 

1,450 

10,350 

670 

3,720 

6,240 

5,500 

1,840 

17,970 

1,170 

5,550 

8,900 

6,700 

1,840 

24,160 

1,960 

8,300 

14,000 

9,800 

4,330 

38,390 

Cl = chlorine 

CuCl2 = copper chloride 

Mg-Al = magnesium-aluminum 

Source: Stieglitz et al. (1989). 

C	 The transition metal compound CuCl2 catalyzed the de novo synthesis of CDDs/CDFs 

on the surface of particulate carbon in the presence of oxygen, yielding CO2 and 

chlorinated/brominated aromatic compounds. 

C	 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. 

Stieglitz et al. (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. 

The investigators found that the fly ash contained 900 ng/g of bound organic chlorine, of which 

only 1% was extractable. Heating the fly ash at 300 to 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 (5% extractable total organic chlorine at 300°C 

and 25 to 30% at 400°C). From this, the authors concluded that the oxidation and degradation of 

carbon in fly ash are the sources of the formation of CDDs/CDFs; therefore, they are essential in 

the de novo synthesis of these compounds. 

2-14
 
 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 g of carbon-free MWC fly ash 

were combined with 0.1 g of activated carbon and placed into a glass tube between two glass 

wool plugs. The glass tube was then placed into a furnace at specific temperatures ranging from 

200 to 400°C. This protocol was repeated for a series of retention times and temperatures. The 

investigators observed that CDD/CDF formation was optimized at 300°C and at the furnace 

retention times of 4 to 6 hr. Figure 2-2 displays the relationship between retention time and 

temperature in CDD/CDF production from the heating of carbon particulate. 

m

a gr r pe

s

m nogr

a Na

Rete 

ntion time (hr) 

Figure 2-2. The de novo synthesis of CDDs/CDFs from heating carbon 

particulate at 300°C at varying retention times. 

Source: Addink et al. (1991). 

Addink et al. (1991) also investigated the relationship between furnace temperature and 

CDD/CDF production from the heating of carbonaceous fly ash. Figure 2-3 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 

2-15
 
m

a gr r pe

s

m nogr

a Na

Temperature (°C) 

Figure 2-3. Temperature effects on CDD/CDF formation. 

Source: Addink et al. (1991). 

temperature and the CDD/CDF congener profile: at 300 to 350°C, the less-chlorinated tetra- and 

penta-CDD/CDF congeners increased in concentration, whereas hexa-, hepta-, and octaCDD/CDF congeners either remained the same or decreased in concentration. The congener 

profile of the original MWC 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 that some 

form of equilibrium is reached in the formation kinetics. 

Gullett and Lemieux (1994) used a pilot-scale combustor to study the effect of varying 

combustion gas composition, temperature, residence time, quench rate, and sorbent (Ca[OH]2) 

injection on CDD/CDF formation. The fly ash loading was simulated by injecting fly ash 

collected from a full-scale MWC. Sampling and analysis indicated that CDDs/CDFs formed on 

the injected fly ash at levels representative of those observed at full-scale MWCs. A statistical 

2-16
 
analysis of the results showed that, although the effect of combustor operating parameters on 

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 excess air (which also affects the ratio of CDDs to CDFs formed), and 

increase quench rate. 

Milligan and Altwicker (1995) found that increases in the carbon gasification rate caused 

increases in the amounts of CDDs/CDFs formed and gave further evidence linking the oxidation 

of carbon to the formation of CDDs/CDFs. Neither the gas-phase CO2 or CO (products of 

carbon oxidation) act as precursors to chlorobenzenes or CDDs/CDFs from reactions with carbon 

particulate (Milligan and Altwicker, 1995). Activated carbon, with its high surface area and 

excellent adsorptive characteristics, also has the highest gasification rate of all residual carbon 

(Addink and Olie, 1995). 

Experimental evidence suggests the following factors for the de novo synthesis of 

CDDs/CDFs from carbon: (a) carbon consisting of imperfect and degenerated layers of graphite, 

(b) the presence of oxygen, (c) the presence of chlorine, (d) catylization of the reactions by CuCl2 

or some other transition metal, and (e) temperatures in the range of 200 to 350°C (Huang and 

Buekens, 1995). The oxidation of carbon in fly ash is apparently inhibited at temperatures below 

200°C, thus indicating the lower temperature limit for the thermal inertization of de novo 

synthesis (Lasagni et al., 2000). 

Lasagni et al. (2000) determined that at a temperature of 250°C, the primary product of 

the gasification of carbon in fly ash is CO2, but in a temperature range of 250 to 325°C, organic 

compounds are formed as products of the oxidation of the carbon. Addink and Olie (1995) 

raised the possibility that the molecular backbone of CDDs/CDFs may be present in carbon. If 

this is the case, the generation of dioxins and furans from the oxidation of carbon would not 

require the formation of intermediate aromatic ring structures. More work is needed to confirm 

these possibilities. 

The de novo synthesis of CDDs/CDFs also involves the possibility that aromatic 

precursors are formed within the post-combustion zone in the following manner: (1) fuel 

molecules are broken down into smaller molecular species (e.g., C1 and C2 molecules) during 

primary combustion, and (2) these simple molecules recombine in the post-combustion zone to 

form larger-molecular aromatic species (i.e., chlorobenzenes and chlorophenols) (Altwicker et 

al., 1993). Thus, small molecular products that evolve in the hot zone of the furnace as a 

consequence of incomplete fuel or feed material combustion may be important foundation 

molecules to the subsequent formation of precursor compounds in the cooler, post-combustion 

region. 

2-17
 
Eklund et al. (1988) reported formation of a wide range of chlorinated organic 

compounds, including CDDs, CDFs, and PCBs, from the oxidation of methane with HCl at 

temperatures of 400 to 950°C in a quartz flow tube reactor. No active catalysts or reactive fly 

ashes were added to the combustion system. From these experimental results, the authors 

hypothesized that chlorocarbons, including CDDs/CDFs, are formed at high temperatures via a 

series of reversible reactions starting with chloromethyl radicals. The chloromethyl radicals can 

be formed from the reaction of methyl radicals and HCl in a sooting flame. Methane is 

chlorinated by HCl in the presence of oxygen at high temperatures, forming chlorinated 

methanes, which react with methyl radicals at higher temperatures (e.g., 800°C) to form aromatic 

compounds. In an oxidative atmosphere, chlorinated phenols are formed, but alkanes and 

alkenes are the primary products. The chlorinated phenols then act as precursors for the 

subsequent formation of CDDs/CDFs. 

Aliphatic compounds are common products of incomplete combustion, and they may be 

critical to the formation of simple ring structures in the post-combustion zone (Weber et al., 

1999; Sidhu, 1999; Froese and Hutzinger, 1996a, b; Jarmohamed and Mulder, 1994). The 

aromatic precursor compounds may be formed in a potentially rich reaction environment of 

aliphatic compounds, reactive fly ash particles, HCl, and oxygen. Sidhu (1999) noted that 

combustion of acetylene on carbon (a common combustion effluent) in the presence of gaseous 

HCl and CuCl2 (as a catalyst) at 300°C led to the formation of intermediate precursors and, 

subsequently, CDDs/CDFs. 

Propene oxidized at 350 to 550°C when in contact with reactive MWC fly ash in a flow 

tube reactor formed a wide range of chlorinated aromatic compounds when the resulting 

combustion gases were mixed with HCl (Jarmohamed and Mulder, 1994). Although the 

conversion was low (1 to 3%), the oxidation of propene on fly ash in the presence of HCl can 

yield chlorinated benzenes and monobenzofurans. Incorporating an oxygen atom into the 

monobenzofuran structure then leads to the formation of monodibenzofuran. The HCl 

contributes chlorine to the aromatic ring through the Deacon reaction, and cyclization on the fly 

ash surface can yield cyclohexadienyl-substituted benzenes, which in turn can be further oxidized 

into CDFs. 

Froese and Hutzinger (1996a) investigated the heterogeneous combustion reactions of the 

nonchlorinated C2 aliphatics. Acetylene, as a model aliphatic compound, was allowed to react 

with precleaned MWC fly ash in a tube flow reactor at approximately 600°C. Metal oxides 

(silicon dioxide [SiO2], iron oxide [Fe2O3], and copper oxide [CuO])—rather than the metal 

chlorides used in other precursor experiments—were added separately as catalysts. The reactants 

were put into contact with HCl vapor, which was introduced at a constant flow rate. The 

2-18
 
acetylene flow was set at 1.1 mL/min and constantly fell to near 0.9 mL/min over 30 min. 

Regulated air flow maintained homeostatic oxidation conditions. 

Chlorobenzenes and chlorophenols were formed, with isomer patterns generally 

resembling isomer patterns of chlorobenzene and chlorophenol emissions from MWCs. CuO 

was seen as catalyzing condensation and chlorination reactions under heterogeneous conditions 

to form the chlorinated CDD/CDF precursor compounds. Other more volatile compounds 

formed were short-chain aliphatic products, such as chloromethane, dichloromethane, and 

chloro- and dichloroacetylene. Chlorobenzene congeners were not the major products formed; 

perchlorinated aliphatic compounds dominated as gas-phase reaction products. 

Froese and Hutzinger (1997) noted that perchlorinated aliphatic compounds (e.g., 

hexachloropropene, hexachloro-1,3-butadiene, and hexachlorocyclopentadiene) are important 

intermediates in aromatic ring formation; they concluded that the catalytic reaction of C2 

aliphatic compounds at 600°C dramatically contributes to the formation of chlorinated and 

nonchlorinated aromatic compounds during combustion. Thus, aliphatic compounds can form 

CDD/CDF precursor compounds. Variable temperature effects were observed in the formation 

of CDDs/CDFs in the same reactions. Maximal OCDD formation occurred at 400°C, and the 

tetra through hepta homologue groups were maximally formed at 600°C. For CDFs, production 

of more highly chlorinated homologues occurred at 400°C, and the formation of TCDFs occurred 

at 500°C. Froese and Hutzinger (1996a) noted a 100-fold increase in TCDF formation at 500°C 

when compared with formation at 400°C. An explanation for this increase is that the higher 

temperature maximized the formation of the CDD/CDF precursor (chlorophenol) from the 

aliphatic starting compound. 

Froese and Hutzinger (1996b) produced polychlorinated benzene and phenol compounds 

at a temperature range of 300 to 600°C, caused by the heterogeneous combustion reactions of 

ethylene and ethane over fly ash in the presence of HCl, oxygen, and a metal catalyst. No 

chlorobenzene congener precursors were formed from ethylene and ethane at 300°C; however, 

the formation rate increased with temperature until a maximum production was achieved at 

600°C. No definitive temperature dependence was observed for the formation of chlorophenols 

from the aliphatic starting compounds. However, at 500°C, 2,4,6-trichlorophenol dominated the 

reaction products; at 300°C, PCP was initially produced. 

Froese and Hutzinger (1996b) also investigated the effects of elemental catalysts on 

potentiating the heterogeneous combustion reactions by measuring the amount of chlorobenzene 

and chlorophenol product formed from the reactions of ethylene/HCl over each catalyst at 

600°C. The reaction with SiO2 did not have a catalytic effect. Aluminum oxide (Al2O3) catalytic 

action showed high intensity for the dichlorobenzene isomers and decreasing intensity for the 

higher-chlorinated isomers. Comparison of the amount of dichlorobenzene product formed 

2-19
 
indicated that an equal quantity was produced with either Al2O3 or fly ash; however, Al2O3 

formed four to five times more product than did the CuO catalyst. For tri- to hexachlorobenzene 

congeners, MWC fly ash reactions produced 5 to 10 times more product than did the metal 

catalysts. However, the presence of the CuO catalyst in these reactions produced a 

chlorobenzene congener pattern comparable to that of the fly ash reactions. With regard to 

chlorophenol production, Al2O3 also produced a unique dichlorophenol pattern, suggesting that 

Al2O3 has a unique catalytic effect in the high-temperature reactions of C2 aliphatic compounds. 

Reactions with CuO produced additional products, including chlorinated methyl 

compounds, chlorinated C2 aliphatics, and perchlorinated C3–C5 alkyl compounds. Froese and 

Hutzinger noted that these perchlorinated alkyl groups, formed by reacting ethylene and ethane 

over fly ash in the presence of the CuO catalyst, were key intermediate compounds to the 

formation of first aromatic rings in typical combustion systems. This emphasizes the importance 

of copper’s catalytic effects in a combustion fly ash system. Al2O3 catalyzed reactions produced 

nonchlorinated naphthalene and alkylbiphenyl compounds. Furthermore, the organic chlorine in 

aliphatic compounds may also act as a direct source of chlorine for the formation of CDDs/CDFs 

in a carbon fly ash system (Weber et al., 1999). 

In an earlier experiment using a similar flow tube apparatus, Froese and Hutzinger (1994) 

formed chlorinated benzenes and phenols in fly ash catalyzed reactions with trichloroethylene at 

temperatures of 400 to 500°C. In this case, metal oxides (CuO, FeO3, and Al2O3) were used as 

catalysts, but no HCl was added for oxychlorination of product compounds. Under combustion 

conditions, temperature-dependent formation of chlorinated aromatics occurred from the 

trichloroethylene starting compound. Reaction with fly ash at 600°C formed hexachlorobenzene 

in concentrations that were about 1,000 times greater than those at 400 and 500°C, with similar 

results for chlorophenols. The authors hypothesized that key aromatic precursors for 

CDDs/CDFs are formed in the higher-temperature region of a post-combustion zone (about 

600°C) and are then carried to the cooler post-combustion region (about 300°C), where the 

precursors form CDDs/CDFs. 

2.4. THE ROLE OF CHLORINE IN THE FORMATION OF CDDs/CDFs IN 

COMBUSTION SYSTEMS 

The formation of CDDs/CDFs in the post-combustion region of combustion systems via 

either the precursor or de novo synthesis mechanisms requires the availability of a source of 

chlorine (Luijk et al., 1994; Addink et al., 1995; Stanmore, 2004; Wikstrom et al., 2003 ). 

Chlorine concentration in this region is somehow related to the chlorine content of combustion 

fuels and feed materials in incineration/combustion systems because there can be no othersource. 

The main question regarding the role of chlorine in forming CDDs/CDFs is whether a positive 

2-20
 
and direct correlation exists between the amount of chlorine in feeds and the amount of 

CDDs/CDFs formed and emitted from the stack of a combustion system. If a direct relationship 

appears to exist, then reductions in the chlorine content of fuels/feeds prior to combustion should 

result in a corresponding reduction in the concentrations of CDDs/CDFs formed after 

combustion. If the oxychlorination reactions require a number of steps, then the relationship 

between chlorine in uncombusted fuels and CDDs/CDFs formed after combustion may not be 

linear, although it may still be dependent in some nonlinear association. The main question can 

best be addressed by examining both formation mechanisms revealed in laboratory-scale 

combustion experiments and correlations between chlorine inputs with CDD/CDF outputs in 

commercial-scale combustors. 

2.4.1. Review of Laboratory-Scale Studies 

A wide body of experimental evidence has elucidated the direct and indirect associations 

between chlorine in feeds and fuels and the potential formation of CDDs/CDFs during 

combustion. The de novo synthesis of CDDs/CDFs requires two basic reactions: (1) the transfer 

of chlorine to residual carbon particulate, with subsequent formation of carbon-chlorine bonds, 

and (2) the oxidation of this macromolecular complex to yield CO2 and volatile and semivolatile 

organic compounds as side products (Weber et al., 1999). Transition metal compounds such as 

CuCl2 catalyze these reactions. Gaseous HCl, Cl2, and Cl@ are the most abundant sources of 

chlorine available for participation in the formation of CDDs/CDFs, and they are initially formed 

as a combustion by-product from the inorganic and organic chlorine contained in the fuel 

(Wikstrom et al., 2003; Rigo, 1998; Addink et al., 1995; Rigo et al., 1995; Halonen et al., 1994; 

Luijk et al., 1994; Altwicker et al., 1993; Wagner and Green, 1993; Dickson et al., 1992; Bruce 

et al., 1991; Gullet et al., 1990b; Commoner et al., 1987; Vogg et al., 1987). 

MSW contains approximately 0.45 to 0.90% (w/w) chlorine (Domalski et al., 1986). The 

most predominant chlorine species formed from MSW combustion is gaseous HCl, which 

averages between 400 and 600 ppm in the combustion gas (Wikstrom et al., 2003; U.S. EPA, 

1987a). Chlorine is initially released from the chlorine in the MSW and is rapidly transformed to 

HCl by the abstraction of hydrogen from reaction with hydrocarbons present in the fuel 

(Wikstrom et al., 2003). HCl may oxidize to yield Cl2 gas by the Deacon reaction, and the Cl2 

directly chlorinates a CDD/CDF precursor along the aromatic ring structure. Further oxidation of 

the chlorinated precursor in the presence of a transition metal catalyst (of which CuCl2 was found 

to be the most active) yields CDDs/CDFs (Altwicker et al., 1993). Increasing the yield of 

chlorine in vapor phase from HCl oxidation generally increases the rate of CDD/CDF formation. 

Formation kinetics are most favored at temperatures ranging from 200 to 450°C. However HCl 

2-21
 
is considered a weak chlorinating agent because of the tenacity of the hydrogen-to-carbon bond 

of aromatic compounds (Wikstrom et al., 2003). 

Chlorine production from gaseous HCl can be reduced either by limiting initial HCl 

concentration or by shortening the residence time (Bruce et al., 1991; Gullett et al., 1990b; 

Commoner et al., 1987). Bruce et al. (1991) observed a general increase in CDD/CDF formation 

with increases in the vapor-phase concentration of chlorine and verified a dependence of the 

formation of CDDs/CDFs in the post-combustion zone on the concentration and availability of 

gaseous chlorine. This latter finding is in agreement with the results of a simple experiment by 

Eklund et al. (1986) in which unsubstituted phenol was mixed with HCl at 550°C in a quartz 

tube reactor. A wide range of toxic chlorinated hydrocarbons were formed, including 

CDDs/CDFs. Eklund et al. (1988) also found a dependence of the amounts of chlorinated phenol 

product formed from the nonchlorinated starting material on the increased amount of HCl 

introduced into the reaction. Under the conditions of this experiment, no chlorinated compounds 

were formed at an HCl concentration of less than 10-3 mol, and maximum chlorophenol 

concentration occurred at around 108 mol. 

Born et al. (1993) also observed that increasing levels of HCl gave rise to increasing rates 

of oxychlorination of precursors, with increasing chances for the post-combustion formation of 

CDDs/CDFs. However, Addink et al. (1995) observed that an HCl atmosphere and/or chlorine 

produced approximately equal quantities of CDDs/CDFs during the de novo synthesis from 

oxidation of particulate carbon. Such results suggest that chlorine production via the Deacon 

reaction in the de novo synthesis may not be the only chlorination pathway, and they may 

indicate that the HCl molecule can be a direct chlorinating agent. In addition, some chlorine is 

expected to be formed from the oxidation of metal chlorides (e.g., CuCl2), but Cl2 formation from 

the Deacon reaction is greater because of the continuous supply of HCl delivered from the 

combustion chamber (Bruce et al., 1991). In this case, a first-order dependence of HCl to Cl2 is 

observed. 

However, Wikstrom et al. (2003) reported on the importance of chlorine species on the de 

novo formation of CDDs/CDFs. HCl can react with oxidizing radicals (e.g., hydroxyl radical, or 

OH) to produce ClA. ClA are highly reactive and can replace hydrogen atoms with chlorine atoms 

in the H-C bond of the aromatic structure. Thus, HCl is most likely an indirect chlorinating agent 

via the formation of ClA. 

Experimentally, about 18% of the total chlorine content in fuels can be thermally 

converted to ClA in the post-combustion region (Procaccini et al., 2003). Although HCl is the 

primary chlorine-containing product formed from the combustion of chlorine-rich fuels, it may 

not be the major chlorinating agent in the formation of chloro-organics in the cooled-down 

region of the combustor. The experiments by Procaccini et al. (2003) indicate that the major role 

2-22
 
of HCl in the formation of chloro-organic compounds at cooler temperatures may be that of a 

chemical progenitor of ClA. HCl reacts with the oxidizing radicals OH and O that are abundantly 

present in combustion off-gases to reform ClA. ClA readily abstract hydrogen atoms from the H-C 

bond of aromatic compounds formed as combustion by-products of organic fuels. By this means, 

unsubstituted aromatic compounds, e.g., benzene, undergo oxy-chlorination reactions with the 

ClA to form chlorobenzenes and chlorophenols. These products are well-defined precursor 

compounds for the synthesis of CDDs/CDFs. 

Wagner and Green (1993) investigated the correlation of chlorine content in feed to stack 

emissions of chlorinated organic compounds in a pilot-scale incinerator using HCl flue gas 

measurements as a surrogate for fuel-bound organic chlorine. In addition to MSW as a fuel, 

variable amounts of polyvinyl chloride (PVC) resin were added during 6 of 18 stack test runs. 

The resulting data were regressed to determine the coefficient of correlation between HCl 

measurements and total chlorobenzene compound emission measurements. In nearly all of the 

regression analyses performed, the relationship between HCl emissions and emissions of 

chlorinated organic compounds was positive and well defined. In addition, the investigators 

found a direct dependence of HCl emission levels on the level of PVC in the waste, with 

generally increasing amounts of HCl formed as increasing amounts of PVC were added. From 

these experiments, they concluded that decreased levels of organically bound chlorine in the 

waste incinerated led to decreased levels of chlorinated organic compounds in stack emissions. 

Kanters and Louw (1994) investigated a possible relationship between chlorine content in 

waste feed and chlorophenol emissions in a bench-scale thermal reactor. MSW incineration with 

a higher content of chlorine in the feed caused higher emissions of chlorophenols via the de novo 

synthesis pathway. The investigators lowered the chlorine content of the prototype MWC by 

replacing chlorine-containing fractions with cellulose. They observed appreciable decreases in 

the amounts of chlorophenol formed from combustion, and concluded that reductions in the 

chlorine content of waste feeds or elimination of PVC prior to municipal waste combustion 

should result in a corresponding reduction in chlorophenol and CDD/CDF emissions. 

In a similar experiment, Wikstrom et al. (1996) investigated the influence of chlorine in 

feed materials on the formation of CDDs/CDFs and benzenes in a laboratory-scale fluidized-bed 

reactor. Seven artificial fuels (composed of 34% paper, 30% wheat flour, 14% saw dust, 7% 

polyethylene (PE), and 2% metals), to which varying amounts of organic chlorine and inorganic 

chlorine (CaCl2 A 6H2O) were added, were combusted. The chlorine content of these fuels varied 

from 0.12 to 2%. All combustion was performed with a high degree of combustion efficiency 

(99.999%) to avoid the formation of polyvinylidene chloride and naphthalenes as products of 

incomplete combustion of pure PVC. With the combustion conditions held constant, only the 

2-23
 
chlorine content of the fuel was varied. Flue gases were sampled for CDDs/CDFs and 

chlorobenzenes. 

In these experiments, concentrations of PCB isomers were approximately 1,000-fold 

higher than CDDs/CDFs (expressed as concentration of I-TEQDF). Moreover, a correlation was 

found between I-TEQDF and PCB levels in the flue gases and the chlorine content of the fuel. A 

fivefold increase in both I-TEQDF and PCB concentrations was observed in the flue gases from 

combustion of fuels containing 0.5 and 1.7% total chlorine. Furthermore, no differences were 

observed in the amount of chlorinated product produced or when the source of chlorine in the 

fuel was organic or inorganic. No correlation was observed between total CDD/CDF and PCB 

formation and total chlorine in the feed when chlorine levels in feed were 0.5% or lower. The 

highest amounts of CDDs/CDFs and PCBs were formed from the fuel with the highest total 

chlorine content (1.7%). 

Under the conditions of this experiment, Wikstrom et al. (1996) observed that a chlorine 

fuel content of 1% was a threshold for formation of excess CDDs/CDFs and PCBs during 

combustion. The authors noted that MSW in Sweden contained about 0.7% chlorine, of which 

approximately 40% was organic chlorine. They concluded that MSW was below the observed 

threshold value of 1% chlorine content associated with a general increase in CDD/CDF and PCB 

formation in the post-combustion region. They also stated that their study did not support the 

hypothesis that elimination of only PVC from waste prior to combustion will cause a significant 

reduction in CDD/CDF emissions if the combustion process is well controlled (high combustion 

efficiency). Wang et al. (2003) verified the existence of a theoretical chlorine-in-fuel threshold 

when they demonstrated de novo synthesis when combusting fuels with 0.8 to 1.1% chlorine. 

A primary by-product of PVC combustion is HCl. Paciorek et al. (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% of the theoretical amount, based on the percent 

weight of chlorine on the molecule. Ahling et al. (1978) 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. 

They observed data from 11 separate experiments conducted with temperatures ranging from 570 

to 1,130°C. These data 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 ClA and HA generated in the incineration process may 

attack the chlorinated benzenes and abstract hydrogen atoms to produce orthochlorine-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. The chlorine in aliphatic 

2-24
 
compounds has been observed as both yielding high amounts of HCl during combustion and 

acting as a direct chlorine source for the de novo synthesis of CDDs/CDFs (Weber et al., 1999). 

Kim et al. (2004) determined that the combustion of pure PVC yielded appreciable 

amounts of polycyclic aromatic hydrocarbons (PAHs), PCBs, chlorobenzenes, and 

chlorophenols. They suggested that the gas-phase production of PCBs and chlorobenzenes 

contributed to the gas-phase formation of CDDs/CDFs through the precursor mechanism. 

Chlorophenols, however, contributed to the de novo formation. Kim et al. (2004) reported that 

the de novo synthesis of CDDs/CDFs from chlorophenols was approximately 100 times greater 

than their formation from PCB and chlorobenzene precursors. 

Katami et al. (2002) found a clear correlation between dioxin formation and the chlorine 

content of mixed plastics combusted in a laboratory-scale incinerator. PVC, PE, polystyrene 

(PS), polyethylene terephthalate (PET), and their various mixtures were burned at temperatures 

greater than 600°C. Average CO concentrations in the exhaust gases were varied from 2 to 880 

ppm as a general indication of the quality of the fire in the combustion chamber. When 

incinerated, each type of plastic formed CDDs/CDFs in the exhaust gases. Of the total CDDs 

formed, HxCDD and TCDD formed in the greatest amounts when PE was combusted. MonoCDF was the most abundant CDF formed from PE combustion. Mono-ortho coplanar PCBs 

were preferentially formed over nonortho-PCBs. The combustion of PS caused TCDD to be 

formed in the greatest abundance of all possible CDDs, whereas TCDF was the most abundant 

dibenzofuran. Mono-ortho PCBs formed more than nonortho coplanar PCBs when PS was 

combusted. The combustion of PET mostly formed MCDD and MCDF among the CDDs/CDFs 

formed. 

When PVC was combusted with the conditions of high temperature and low CO (good 

combustion), a total of 53.5 ng/g of total CDD was formed, with the HxCDD predominating. In 

addition, good combustion conditions formed a total of 771 ng/g of CDFs, with Cl2 and Cl3 CDF 

congeners dominating. When PVC was combusted with the conditions of low temperature and 

high CO (poor combustion), the total CDDs and CDFs formed increased significantly to 429 ng/g 

and 8,492 ng/g, respectively. TrCDD and DiCDF dominated the congener distributions, 

suggesting that poor combustion of PVC tends to form high levels of lower-chlorinated 

CDDs/CDFs. The investigators observed that maintaining good combustion tended to minimize 

the formation of CDDs/CDFs from the combustion of chlorinated plastics. 

Shibata et al. (2003) reported on the formation of CDDs/CDFs from the combustion of 

PVC in quartz ampules. Synthesis of CDDs/CDFs proceeded de novo in a temperature range of 

200 to 400°C, with the reaction catalyzed by CuO. Maximum formation occurred at 300°C. 

HpCDDs and OCDD were the dominant CDDs observed in the flue gases, whereas TCDFs, 

2-25
 
PeCDFs, and HxCDFs dominated the CDFs. The ratio of CDFs to CDDs from PVC combustion 

was greater than 1, which is typical of MSW combustion (Shabata et al., 2003). 

Addink and Altwicker (1999) reported on the role of the inorganic chloride ion in the 

formation of CDDs/CDFs using the labeled compound Na37Cl. The inorganic chloride ion forms 

carbon-chlorine bonds on soot particles during combustion. The chlorine in the soot can be 

directly inserted into a CDD/CDF molecule during formation, or it can exchange with the 

chloride ions in the transitional metal catalyst, which promotes CDD/CDF formation. Thus, the 

inorganic chlorine ion participates as a chlorine donor to CDD/CDF formation. 

De Fre and Rymen (1989) reported on the formation of CDDs/CDFs from hydrocarbon 

combustion in a domestic gas/oil burner in the presence of 15 and 300 ppm concentrations of 

HCl. More than 100 chlorinated organic compounds were detected in the flue gases whenever 

HCl was injected into the system. The investigators observed formation of CDDs and CDFs in 

all experiments where HCl was injected in a hydrocarbon flame. In this case, CDFs were always 

more abundant than CDDs. It was concluded that the relationship between the HCl concentration 

and the emitted concentration of CDDs/CDFs under fixed combustion conditions appeared to be 

exponential for a wide range of temperatures (240 to 900°C). 

2.4.2. Review of Full-Scale Combustion Systems 

The review of experimental data clearly indicates an association between chlorine content 

of feed/fuels and the potential synthesis of CDDs/CDFs. Paradoxically, the review of full-scale 

operating incineration processes does not yield such unequivocal results, indicating that complex 

kinetic events make strong associations difficult in full-scale systems. The following is a review 

of studies of the association between chlorine in feeds and stack releases of CDDs/CDFs in fullscale incineration systems. 

 In the stack testing of a variety of industrial stationary combustion sources during the 

National Dioxin Study in 1987, EPA made a series of qualitative observations about the 

relationship between total chlorine present in the fuel/waste and the magnitude of emissions of 

CDDs/CDFs from the stack of the tested full-scale combustion facilities (U.S. EPA, 1987a). In 

general, combustion units with the highest CDD emission concentrations had greater quantities 

of chlorine in the fuel; 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 was reported by Lustenhouwer et al. (1980) as coal, 1,300 :g/g; MSW, 2,500 :g/g; leaded 

gasoline, 300 to 1,600 :g/g; and unleaded gasoline, 1 to 6 :g/g. 

Thomas and Spiro (1995) also analyzed the relationship between CDD/CDF emissions 

from combustion and the chlorine content of feed materials. Thomas and Spiro (1996) plotted 

average CDD/CDF emission factors for a variety of combustion systems and processes (black 

2-26
 
liquor boilers, unleaded gasoline combustion, leaded gasoline combustion, wire incineration, 

cigarette combustion, sewage sludge incineration, MWC, PCP-treated wood combustion, 

hazardous waste incineration, and hospital waste incineration) against the average chlorine 

concentration of the combusted material. The plot showed that average CDD/CDF emissions of 

combustion source categories tended to increase with the average chlorine content of the 

combusted fuel. This analysis indicated that combustion sources with relatively high combustion 

efficiency and adequate air pollution controls tended to have emissions two orders of magnitude 

lower than those of poorly operated sources. This suggests that the magnitude of CDD/CDF 

emissions is strongly dependent on chlorine concentration in fuels in the context of the more 

poorly controlled and operated combustion sources, and the association becomes less apparent in 

the well-controlled facilities operating with good combustion practices. The slope of the log-log 

plot was between 1 and 2 for the poorly controlled and operated facilities, indicating that the 

relationship between chlorine content and CDD/CDF emissions was more than proportional. 

Costner (1998) reported finding a positive correlation between chlorine content of feed 

material and CDD/CDF emissions at a full-scale hospital waste incinerator. Costner concluded 

that emissions at this facility were dependent on chlorine input at a concentration as low as 

0.031% and that there was no evidence of a threshold in the relationship between chlorine in feed 

and CDD/CDF emissions. 

Rigo et al. (1995) summarized the results of a study commissioned by the American 

Society of Mechanical Engineers (ASME, 1995). The study was a statistical evaluation of the 

relationship between HCl concentration in flue gases and various combustion systems (MWCs, 

hospital waste incinerators, HWIs, biomass combustors, laboratory combustors, and bench-scale 

combustors) and stack emissions of total CDDs/CDFs. In this study, HCl was used as a 

surrogate for total chlorine content in the fuel. The data analysis was sufficient for 92 facilities in 

the database that showed both HCl and CDD/CDF emissions. Of the 92 facilities, 72 did not 

show a statistically significant relationship between chlorine input and CDD/CDF output in 

emissions streams, 2 showed increasing CDD/CDF concentrations with increasing chlorine, and 

8 showed decreasing CDD/CDF concentrations with increasing chlorine. ASME (1995) reports 

the following conclusion: 

The failure to find simultaneous increases in most cases and finding inverse 

relationships in a few indicates that any effect chlorine has on CDD/CDF 

emissions is smaller than the variability of other causative factors. Whatever 

effect chlorine has on CDD/CDF emissions in commercial-scale systems is 

masked by the effect of APCS (air pollution control systems) temperature, ash 

chemistry, combustion conditions, measurement imprecision, and localized flow 

stratification. 

2-27
 
Liberson and Belanger (1995) reported the results of an analysis of the formation and 

emission of CDDs/CDFs as a function of total chlorine in combustion feed materials at a rotary 

kiln HWI. The data were generated from multiple test series conducted over a 13-month period 

at the HWI while operating a carbon injection system specifically designed to control and reduce 

CDD/CDF stack emissions. The chlorine feed rates ranged from 0 to 3,300 lb/hr, and the 

CDD/CDF emission rates ranged from 0.7 to 39 ng/dscm. The authors noted that multiple series 

of CDD/CDF control systems were used on this HWI (a high-temperature secondary combustion 

chamber, a spray dryer-evaporative quench that further cools the combustion gases, activated 

carbon injection to adsorb semivolatile organics, and a cool-side electrostatic precipitator 

followed by an acid gas scrubber to collect HCl and Cl2). From analyses of the data, the authors 

concluded that no correlation exists between CDD/CDF emissions and chlorine feed in a modern 

MWC using carbon injection for CDD/CDF control. 

More recently, Wang et al. (2003) investigated the association between chlorine content 

of waste feeds and CDD/CDF emissions from full-scale combustion systems. Previously, 

Wikstrom et al. (1996) had discerned a chlorine content in feeds of 1% as being a threshold 

concentration for the formation of CDDs/CDFs, i.e., an association with the magnitude of 

CDDs/CDFs formed occurred only when chlorine content in the feed was $1%. Wang et al. 

confirmed the apparent existence of a chlorine threshold for emissions of total CDDs/CDFs after 

statistically reviewing input of chlorine in feed versus output of CDDs/CDFs in emissions at two 

tested medical incinerators and two tested MWCs. Additionally, the authors examined secondhand data from 13 other dioxin sources obtained from the literature and found that the formation 

of CDFs was greater than the formation of CDDs when the chlorine content of the waste feed 

exceeded the threshold. However, when the chlorine content was below the approximate 1% 

threshold, the formation of CDDs was greater than the formation of CDFs. The authors proposed 

that chlorine content below the threshold formed chlorinated precursors to CDDs rather than 

forming the dibenzofuran molecule. Chlorine content above the threshold contributed to 

deterioration of combustion conditions, causing the formation of PAHs, which, in turn, 

contributed to the formation of CDFs. 

2.5. POTENTIAL PREVENTION OF CDD/CDF FORMATION IN COMBUSTION 

SYSTEMS 

Given what is currently understood about oxychlorination reactions in the synthesis of 

CDDs/CDFs, researchers have identified certain interventions that could be taken to reduce or 

impede formation in combustion systems. Raghunathan and Gullett (1996) demonstrated in a 

pilot-scale incinerator that sulfur compounds can combine with the metal catalyst necessary to 

stimulate the Deacon reaction of HCl and oxygen to yield Cl2, thereby neutralizing the catalyzing 

2-28
 
agent and reducing the formation of CDDs/CDFs. The Deacon reaction, which forms Cl2 in the 

combustion plasma, is seen as occurring only in the presence of a catalyst. Thus, the sulfur 

dioxide (SO2) molecule (formed when sulfur in the fuel combines with oxygen) can inhibit the 

catalytic activity of the fly ash by either combining with a metal-based Deacon catalyst in the fly 

ash or depleting the Cl2 formed. The authors observed that the principal action of sulfur in 

inhibiting the formation of CDDs/CDFs in combustion systems is through SO2 depletion of Cl2, 

as follows: 

Cl2 + SO2 + H2O : 2HCl + SO2 

The relevance of this finding is that the co-combustion of MSW with coal (that contains 

sulfur) should lead to dramatic reductions in the amount of CDDs/CDFs formed and emitted, and 

it may explain why, in the United States, coal combustion at power plants results in CDD/CDF 

emission rates more than a magnitude lower than those at MWCs. 

Naikwadi and Karasek (1989) investigated the addition of calcium oxide (CaO) and 

triethylamine (TEA) to the flue gases of a combustion system as an inhibitor of the catalytic 

activity of fly ash. They placed 500 :g 13C-labeled PCP (a dioxin precursor) in a combustion 

flow tube and allowed it to react with organic-extracted MWC fly ash at 300°C under an air 

stream. Under these conditions, CDDs/CDFs were formed at concentrations ranging from 1,660 

to 2,200 ng/100 :g 13C-PCP. The experimental method was then modified by mixing reactive 

MWC fly ash with either CaO or TEA. The results showed that the amount of CDDs/CDFs 

formed could be reduced by an order of magnitude from the reaction of PCP with fly ash and the 

addition of TEA as an inhibitor. When CaO was mixed with fly ash, the amount of CDDs/CDFs 

formed decreased more than 20-fold. 

2.6. THEORY ON THE EMISSION OF PCBs 

Air emissions of PCBs from MSW incineration is less well studied. Probably the 

formation mechanisms that apply to CDDs/CDFs would also apply to PCBs. Mechanism 1 (pass 

through) is implicit in the Toxic Substances Control Act rule, which requires 99.9999% 

destruction in HWIs. When this occurs, 0.0001% of the initial amount of PCBs fed into the HWI 

may be emitted from the stack. This may indicate that some small fraction of the PCBs present 

in the fuel fed into an incineration process may result in PCB emissions from the stack of the 

process. 

PCBs have been measured as contaminants in raw refuse prior to incineration in an MWC 

(Choudhry and Hutzinger, 1983; Federal Register, 1991a). Using this information, it is possible 

to test mechanism 1 for CDD/CDF emissions: that the PCB contamination present in the fuel is 

2-29
 
_______ 

mainly responsible for emissions from the stack. The mass balance of total PCBs, beginning 

with measurement in the raw refuse and ending with measurement at the stack of a refusederived fuel (RDF) MWC (Federal Register, 1991a), 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, 1991a), a computation of DRE can be made 

using the following equation (Brunner, 1984): 

WI - WO 

DRE = × 100%

 WI 

where: 

WI = mass rate of contaminant fed into the incinerator system 

WO = mass rate of contaminant exiting the incinerator system 

In test 11, 811 ng total PCB/g refuse were measured in the MSW fed into the incineration 

system and 9.52 ng/g 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% can 

be calculated. Therefore, it appears that PCB contamination in the raw MSW fed into this 

particular incinerator may have accounted for the PCB emissions from the stack of the MWC. 

PCBs can be thermolytically converted into CDFs (Choudhry and Hutzinger, 1983; 

U.S. EPA, 1984). This process occurs at temperatures somewhat lower than those typically 

measured inside the firebox of an MWC. Laboratory experiments conducted by EPA indicate 

that the optimum conditions for CDF formation from PCBs are near a temperature of 675°C in 

the presence of 8% oxygen and a residence time of 0.8 sec (U.S. EPA, 1984). This resulted in a 

3 to 4% efficiency of conversion of PCBs into CDFs. Because 1 to 2% of the PCBs present in 

the raw refuse may survive the thermal stress imposed in the combustion zone of the incinerator 

(Federal Register, 1991a), 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. 

Although it appears that contamination of waste feed with PCBs may be an important 

factor in detecting PCBs in stack emissions from combustion processes, recent research has 

indicated that these compounds may also be formed in the post-combustion region, either from 

de novo synthesis or from precursor compounds. Zheng et al. (1999) observed the formation of 

PCBs in the post-combustion region from the pyrolysis of chlorobenzenes using a laboratoryscale furnace. The investigators observed that PCBs were optimally formed from lesschlorinated chlorobenzenes (e.g., 1,3-dichlorobenzene) catalyzed by CuCl2. In this experiment, 

maximum PCB production occurred at a temperature of 350°C. Wikstrom et al. (1998) reported 

2-30
 
secondary formation of PCBs in the post-combustion region similar to the de novo synthesis of 

CDDs/CDFs, albeit PCBs were formed in only small amounts relative to CDDs/CDFs. 

Fangmark et al. (1994) postulated that formation of PCBs and CDDs/CDFs in the postcombustion region may occur through the same mechanisms. On the other hand, Blumenstock et 

al. (1998) produced results in a pilot-scale furnace that were inconsistent with the de novo 

formation of CDDs/CDFs in the post-combustion region (i.e., PCBs seemed to be optimally 

formed at high temperatures in oxygen-deficient atmospheres). Shin and Chang (1999) noted a 

positive correlation between PCB concentrations on MSW incineration fly ash and fly ash 

concentrations of CDDs/CDFs, suggesting that high PCB levels in fly ash may be a contributory 

cause of the post-combustion formation of CDDs/CDFs (i.e., PCBs are precursors to 

CDDs/CDFs). Nito et al. (1997) noted the formation of CDDs/CDFs from the pyrolysis of PCBs 

in a fluidized-bed system, indicating that PCBs in feeds may account for CDFs formed in MSW 

incineration. More combustion-related research needs to be conducted to firmly establish 

whether PCB contamination in feeds or post-combustion formation (or both) may explain the 

presence of PCBs in combustion flue gases. 

2.7. SUMMARY AND CONCLUSIONS 

2.7.1. Mechanisms of Formation of Dioxin-Like Compounds 

There are three primary mechanisms for CDD/CDF emissions from combustion sources. 

Mechanism 1 (pass through). This mechanism involves CDDs/CDFs contained in the 

feed passing through the combustor intact and being subsequently released into the environment. 

For most systems, this is not thought to be a major contributor to CDD/CDF emissions for three 

reasons. First, for commercial systems with good combustion controls, the temperatures and 

residence times should result in the destruction of most CDDs/CDFs in the feed. Second, mass 

balance studies of a number of combustion systems show that more CDDs/CDFs can be detected 

in the cool-down region downstream of the furnace than in the feed. Third, the CDD/CDF 

congener profile in the feed differs from the congener profile in the stack emissions. 

Consequently, synthesis appears to be a more important mechanism than is pass through. The 

concentration of CDDs/CDFs in the flue gases of any particular combustion system will 

ultimately be derived as a result of the balance between reactions leading to formation and 

reactions leading to destruction of these compounds. 

Mechanism 2 (precursor). This mechanism involves the formation of CDDs/CDFs 

from the thermal breakdown and molecular rearrangement of aromatic precursors either 

originating in the feed or forming as a product of incomplete combustion. Actual synthesis of 

CDDs/CDFs occurs in the post-combustor environment. Gaseous benzene is the most abundant 

aromatic compound associated with products of incomplete combustion of waste. Benzene 

2-31
 
reacts with ClA within the combustion gas plasma, causing aromatic H abstraction and the 

subsequent formation of chlorobenzenes and chlorophenols. Homogeneous gas-phase formation 

of CDDs/CDFs occurs from these precursor compounds at temperatures >500°C, catalyzed by 

the presence of copper compounds. In addition, the CDDs/CDFs can form from gas-phase 

precursors as heterogeneous, catalytic reactions with reactive fly ash surfaces. This reaction has 

been observed to be catalyzed by the presence of a transition metal sorbed to the fly ash. The 

most potent catalyst is CuCl2. Relatively low temperatures—in the range of 200 to 450°C—have 

been identified as a necessary condition for these heterogeneous reactions to occur, with either 

lower or higher temperatures inhibiting the process. Because these reactions involve 

homogeneous gas-phase and heterogeneous solid-phase chemistry, the rate of emissions is less 

dependent on reactant concentration than on conditions that are favorable to formation, such as 

temperature, retention time, source and species of chlorine, and the presence of a catalyst. 

Mechanism 3 (de novo synthesis). This mechanism involves the heterogeneous solidphase formation of CDDs/CDFs in the post-combustion environment on the surface of fly ash. 

Such heterogeneous chemistry occurs in two ways: (1) directly from the oxidation of carbon 

within the fly ash and subsequent reactions with organic and inorganic chlorine, and (2) the 

oxidative breakdown of macromolecular carbon structures (e.g., graphite) and oxychlorination 

reactions of aromatic precursors (such as chlorobenzenes and chlorophenols) on fly ash surfaces, 

leading to CDD/CDF formation. In either case, formation kinetics is most favored at 

temperatures in the range of 200 to 450°C and is promoted by the catalytic properties of either 

the fly ash or the presence of a transition metal compound. 

Mechanisms 2 and 3 can occur simultaneously, share a number of common reaction 

pathways, and occur in the same physical environment, and they are controlled by many of the 

same physical conditions. In well-designed and well-operated combustion systems, the precursor 

species needed for mechanism 2 are reduced; consequently de novo synthesis (mechanism 3) can 

become the dominant pathway for formation. In systems with incomplete combustion, it is 

difficult to sort out the relative contribution of these two mechanisms to total emissions. Both 

mechanisms, however, can be curtailed if steps are taken to minimize the physical conditions 

needed to support formation (i.e., time, temperature, and reactive surface). 

The combustion formation chemistry of PCBs is less well studied than that of 

CDDs/CDFs, but it is reasonable to assume that these same three mechanisms would apply. For 

waste incineration, PCBs can exist in significantly higher concentrations in the feed than do 

CDDs/CDFs. Consequently, mechanism 1 may play a more prominent role in the origin of PCB 

emissions than of CDD/CDF emissions. 

2-32
 
2.7.2. Role of Chlorine 

From the various analyses on the role and relationship of chlorine to CDD/CDF formation 

and emissions, the following observations and conclusions are made. 

1. Although chlorine is an essential component in the formation of CDDs/CDFs in 

combustion systems, the empirical evidence indicates that, for commercial-scale incinerators, 

chlorine levels in feed are not the dominant controlling factor for the amount of CDDs/CDFs 

released in stack emissions. Important factors that can affect the rate of CDD/CDF formation 

include overall combustion efficiency, post-combustion flue gas temperatures and residence 

times, and the types and designs of air pollution control devices employed on combustion 

systems. Data from bench-, pilot-, and commercial-scale combustors indicate that CDD/CDF 

formation can occur by three principal mechanisms. Some of these data, primarily from benchand pilot-scale combustors, have shown direct correlation between chlorine content in fuels and 

rates of CDD/CDF formation. Other data, primarily from commercial-scale combustors, show a 

weaker relationship between the presence of chlorine in feed and fuels and rates of CDD/CDF 

released from the stacks of combustion systems. The conclusion that the amount of chlorine in 

feed is not a strong determinant of the magnitude of CDD/CDF stack emissions applies to the 

overall population of commercial-scale combustors. For any individual commercial-scale 

combustor, circumstances may exist in which changes in chlorine content in feed could affect 

CDD/CDF emissions. For uncontrolled combustion, such as open burning of household waste, 

chlorine content of wastes may play a more significant role in levels of CDD/CDF emissions 

than the one observed in commercial-scale combustors. 

2. Both organic and inorganic forms of chlorine in combustion fuels yield abundant 

gaseous HCl in the post-combustion region. It has been shown that ClA are the most potent 

chlorinating agent in the formation of chloro-organic compounds from combustion. HCl may be 

the dominant chemical progenitor of ClA participating in oxychlorination reactions to CDD/CDF 

synthesis. Formation of ClA from HCl occurs in the cool-down zone via the oxidation of HCl in 

the presence of a transition metal catalyst (the Deacon reaction). Although the preponderance of 

scientific evidence suggests that this is an important pathway for producing chlorinated 

compounds in emissions, it is still unclear whether HCl can also directly chlorinate aromatics or 

whether it must first be oxidized to yield Cl2. 

3. Laboratory-scale experiments have examined correlations between chlorine content of 

feeds and total CDDs/CDFs formed in combustion systems. These experiments suggest that for 

feeds containing <1% chlorine, the amount of CDDs/CDFs formed is independent of the chlorine 

2-33
 
content of the feed. For feeds with a chlorine content >1%, a direct correlation was observed. 

The existence of an apparent threshold to the chlorine content of waste has been verified in fullscale combustion systems. It has not been determined, however, whether these relationships are 

relevant to poorly controlled combustion of wastes and biomass, such as backyard barrel burning, 

landfill fires, and agricultural burning. 

4. The combustion of PVC can contribute to the formation of CDDs/CDFs in two ways. 

First, gaseous HCl is a primary product formed from the combustion of PVC. We have seen that 

HCl is a major contributor of Cl@ necessary for the formation of CDDs/CDFs. Thus, PVC 

indirectly contributes to dioxin synthesis. Second, the combustion of PVC directly forms 

benzene, which is followed by oxychlorination reactions that further form chlorinated benzenes 

and chlorinated phenols; these compounds then act as precursors to CDD/CDF formation. 

5. The most critical factors associated with minimizing CDD/CDF formation in 

combustion systems are (a) achieving nearly complete combustion of the feed through the 

application of good combustion practice (i.e., time, temperature, and turbulence), and (b) 

ensuring that combustion gases are quenched to below the temperature range for heterogeneous 

solid-phase formation chemistry in the post-combustion region of the system, i.e., reducing the 

temperature to below 200°C. 

2.7.3. General Conclusion 

Although the formation chemistry of CDDs/CDFs is more complicated and less 

understood than the relatively simple constructs described in this review, the current weight of 

evidence suggests that the formation mechanisms outlined above describe the principal pathways 

of most CDD/CDF formation and emission from combustion sources. 

2-34
 
3. COMBUSTION SOURCES OF CDDs/CDFs: WASTE INCINERATION 

Incineration is the destruction of solid, liquid, or gaseous wastes through the application 

of heat within a controlled combustion system. The purposes of incineration are to reduce the 

volume of waste that needs land disposal and to reduce the toxicity of the waste. In keeping with 

this definition, incinerator systems can be classified by the types of wastes incinerated: 

municipal solid waste (MSW) incineration (commonly referred to as municipal waste 

combustion), medical and pathological waste incineration, hazardous waste incineration, sewage 

sludge incineration, tire incineration, and biogas flaring. Each of these types of incineration is 

discussed in this chapter. The purpose of this chapter is to characterize and describe waste 

incineration technologies in the United States and to derive estimates of annual releases of CDDs 

and CDFs into the atmosphere from waste incineration facilities for reference years 1987, 1995, 

and 2000. 

As discussed in Chapter 2, combustion research has developed three theories on the 

mechanisms involved in the emission of CDDs/CDFs from combustion systems: (1) 

CDDs/CDFs can be introduced into the combustor with the feed and pass through the system 

unchanged (pass through mechanism), (2) CDDs/CDFs can be formed during combustion 

(precursor mechanism), and/or (3) CDDs/CDFs can be formed via chemical reactions in the postcombustion portion of the system (de novo synthesis). Total CDD/CDF emissions are likely to 

be the net result of all three mechanisms; however, the relative importance of each mechanism is 

often uncertain. 

To the extent practical with the available data, the combustors in each source category 

were divided into classes according to similarity of emission factors. This classification effort 

attempted to reflect the emission mechanisms described above. The emission mechanisms 

suggest that the aspects of combustor design and operation that could affect CDD/CDF emissions 

are (a) furnace design, (b) composition of the waste feed, (c) temperature in the post-combustion 

region of the system, and (d) the type of air pollution control device (APCD) used to remove 

contaminants from the flue gases. Therefore, incineration systems that are similar in terms of 

these factors should have similar CDD/CDF emissions. Accordingly, this chapter proposes 

classification schemes that divide combustors into a variety of design classes based on these 

factors. Design class, as used here, refers to the combination of furnace type and accompanying 

APCD. 

3.1. MUNICIPAL WASTE COMBUSTION 

As noted above, emissions can be related to several factors, including furnace design, 

composition of the waste feed, temperature in the post-combustion region of the system, and 

3-1
type of APCD used to remove contaminants from the flue gases. This section proposes a 

classification scheme that divides municipal waste combustors (MWCs) into a variety of design 

classes based on those factors. Because different APCDs are operated at different temperatures, 

operating temperature is used to define some design classes. To account for the influence of the 

waste feed, the proposed classification system distinguishes between refuse-derived fuel (RDF) 

and normal MSW. This section begins with a description of the MWC technology and then 

proposes the design classification scheme. Using this scheme, the municipal waste combustion 

industry is characterized for the reference years 1987, 1995, and 2000. Finally, the procedures 

for estimating emissions are explained and the results summarized. 

3.1.1. Description of Municipal Waste Combustion Technologies 

For the purposes of this report, municipal waste combustion furnace types are divided 

into three major categories: mass burn, modular, and RDF. Mass burn and RDF technologies 

dominate the large MWC category and modular technology dominates the small MWC category. 

Each of these furnace types is described below, followed by a description of the APCDs used 

with the system. 

3.1.1.1. Furnace Types 

Mass burn. This furnace type was so named because it burned MSW as received (i.e., 

no preprocessing of the waste was conducted other than removal of items too large to go through 

the feed system). Today, a number of other furnace types also burn unprocessed waste, as 

described below. Mass burn furnaces are distinguished from the other types because they burn 

the waste in a single stationary chamber. In a typical mass burn facility, MSW is placed on a 

grate that moves through the combustor. Three subcategories of mass burn technologies are 

described below. 

• Mass burn refractory-walled systems represent an older class of MWCs (generally 

built in the late 1970s to early 1980s) that were designed only to reduce the volume of 

waste disposed of by 70 to 90%. These facilities usually lack boilers to recover the 

combustion heat for energy purposes. In the mass burn refractory-walled design, the 

MSW is delivered to the combustion chamber by a traveling grate or a ram feeding 

system. Combustion air in excess of stoichiometric amounts (i.e., more oxygen is 

supplied than is needed for complete combustion) is supplied both below and above 

the grate. As of 2000, few mass burn refractory-walled MWCs remain; almost all 

have closed or been dismantled. 

• Mass burn waterwall (MB-WW) facilities represent enhanced combustion efficiency, 

as compared with mass burn refractory-walled incinerators. Although it achieves 

similar volume reductions, the MB-WW incinerator design provides a more efficient 

delivery of combustion air, resulting in higher sustained temperatures. Figure 3-1 is a 

3-2
Air 

Super- Generator Econo­ heater mizer 

Stack Induced Draft 

Fan 

Belt Total Riddling Vibrating Conveyor Ash Secondary Conveyor Conveyor Discharge Fan Quench 

Tank 

Combustion 

Grate 

Steam 

Drum Overhead Crane 

Feed 

Bin Water 

Wall 

Section 

Forced Draft 

Fan 

Drying

Grate 

Feed Pit 

Steam Pollution Control Device 

Burnout 

Grate 

Figure 3-1. Typical mass burn waterwall municipal solid waste combustor. 

Source: U.S. EPA (1997a). 

schematic of a typical MB-WW MWC. The term “waterwall” refers to a series of 

steel tubes that run vertically along the walls of the furnace and contain water. Heat 

from combustion produces steam, which is then used to drive an electrical turbine 

generator or for other industrial needs. This transfer of energy is called energy 

recovery. MB-WW incinerators are the dominant form of incinerator found at large 

municipal waste combustion facilities. 

• Mass burn rotary kilns use a water-cooled rotary combustor that consists of a rotating 

combustion barrel configuration mounted at a 15- to 20-degree angle of decline. The 

refuse is charged at the top of the rotating kiln by a hydraulic ram (Donnelly, 1992). 

Preheated combustion air is delivered to the kiln through various portals. The slow 

rotation of the kiln (10 to 20 rotations/hr) causes the MSW to tumble, thereby 

exposing more surface area for complete burnout of the waste. These systems are 

also equipped with boilers for energy recovery. Figure 3-2 is a schematic of a typical 

rotary kiln combustor. 

3-3
Superheater 

Steam 

Drum 

Convection 

Section 

Radiant 

Section Water Drum 

Economizer 

Branch 

Pipe Ring

Header 

Coaxial 

Line 

Rotary

Joint 

Afterburning

Grate 

Residue 

Conveyor 

Windbox 

Hopper 

Feed 

Chute 

Feed Throat 

Chute 

Resistance Shroud 

Door 

Tire 

Combustor Barrel 

Ram Feeding Ring 

System Header 

Windbox 

Support

Frame 

MSW 

Flue 

Gas 

Figure 3-2. Typical mass burn rotary kiln combustor. 

Source: U.S. EPA (1997a). 

Modular. This is a second general type of municipal waste combustion furnace used in 

the United States. As with the mass burn type, modular incinerators burn waste without 

preprocessing. Modular MWCs consist of two vertically mounted combustion chambers (a 

primary and a secondary chamber). In 1995, the combustion capacity of modular combustors 

ranged from 4 to 270 metric tons per day, i.e., they are predominately small MWCs. The two 

major types of modular systems, excess air and starved air, are described below. 

• The modular excess-air system consists of a primary and a secondary combustion 

chamber, both of which operate with air levels in excess of stoichiometric 

requirements (i.e., 100 to 250% excess air). Figure 3-3 illustrates a typical modular 

starved-air MWC. 

3-4
Waste Tipping Floor 

Ram Feeder 

Charge Hopper 

Ash 

Quench 

Primary Air 

Transfer Rams 

Primary Chamber 

Fire Door 

Secondary Chamber 

To Stack or Waste Heat Boiler 

Secondary Air 

Secondary

Gas Burner 

Feed Chute 

Primary 

Gas Burner 

Figure 3-3. Typical modular starved-air combustor with transfer rams. 

Source: U.S. EPA (1997a). 

• In the starved (or controlled) air type of modular system, air is supplied to the primary 

chamber at substoichiometric levels. The products of incomplete combustion entrain 

in the combustion gases that are formed in the primary combustion chamber and then 

pass into a secondary combustion chamber. Excess air is added to the secondary 

chamber, and combustion is completed by elevated temperatures sustained with 

auxiliary fuel (usually natural gas). The high, uniform temperature of the secondary 

chamber, combined with the turbulent mixing of the combustion gases, results in low 

levels of particulate matter (PM) and organic contaminants being formed and emitted. 

Therefore, many existing modular units lack post-combustion APCDs. Figure 3-4 is 

a schematic view of a modular excess-air MWC. 

Refuse-derived fuel (RDF). The third major type of MWC furnace technology is 

designed to combust RDF; this technology is generally used at very large MWC facilities. RDF 

is a general term that describes MSW from which relatively noncombustible items are removed, 

3-5
Process

Control Room

Room

Flue Gas Recirculation (FGR) 

Manifold 

Secondary 

Tertiary Chamber Underfire Chamber Air 

Tipping 

Room Hopper Heat Recovery

Plenum 

Hopper Overfire Door FGR Underfire 

FGR 

Fire Loader door 

Water 

Cooled 

RAM(s) 

Isolation 

Damper 

T 

T 

Relief 

Door 

Ash 

Trough 

and 

Conveyor 

FGR 

Mixing

Zone 

O2 

Ash Conveyor 

Figure 3-4. Typical modular excess-air combustor. 

Source: U.S. EPA (1997a). 

thereby enhancing the combustibility of the waste. RDF is commonly prepared by shredding, 

sorting, and separating out metals to create a dense MSW fuel in a pelletized form having a 

uniform size. Three types of RDF systems are described below. 

• The dedicated RDF system burns RDF exclusively. Figure 3-5 shows a typical 

dedicated RDF furnace using a spreader-stoker boiler. Pelletized RDF is fed into the 

combustor through a feed chute using air-swept distributors; this allows a portion of 

the feed to burn in suspension and the remainder to burn out after falling on a 

horizontal traveling grate. The traveling grate moves from the rear to the front of the 

furnace, and distributor settings are adjusted so that most of the waste lands on the 

rear two-thirds of the grate. This allows more time to complete combustion on the 

grate. Underfire and overfire air are introduced to enhance combustion, and these 

incinerators typically operate at 80 to 100% excess air. Waterwall tubes, a 

3-6
Forced 

Draft Fan 

Steam Coil Air Preheater 

Ash Removal 

Traveling Grate Underfire 

Air 

RDF Distributor 

Overfire Air 

Boiler Tubes 

Feed 

Chute 

Superheater 

Air 

Gas 

Flow 

Gas Burners 

Induced Draft Fan 

Overfire Air 

Economizer 

Stack 

RDF 

Ash Quench Chamber 

Pollution 

Control 

Device 

Figure 3-5. Typical dedicated refuse-derived fuel-fired spreader stoker boiler. 

Source: U.S. EPA (1997a). 

superheater, and an economizer are used to recover heat for production of steam or 

electricity. The 1995 inventory indicated that dedicated RDF facilities range from 

227 to 2,720 metric tons per day total combustion capacity. 

• Cofired RDF furnaces burn either RDF or normal MSW along with another fuel. 

The fluidized-bed RDF furnace burns the waste in a turbulent and semisuspended bed 

of sand. The MSW may be fed into the incinerator either as unprocessed waste or as 

a form of RDF. The RDF may be injected into or above the bed through ports in the 

combustor wall. The sand bed is suspended during combustion by introducing 

underfire air at a high velocity, hence the term “fluidized.” Overfire air at 100% of 

stoichiometric requirements is injected above the sand suspension. Waste-fired 

fluidized-bed RDF furnaces typically operate at 30 to 100% excess air levels and at 

3-7
bed temperatures around 815°C. A typical fluidized-bed RDF furnace is represented 

in Figure 3-6. The technology has two basic designs: a bubbling-bed incineration 

unit and a circulating-bed incineration unit. The 1995 inventory indicated that 

fluidized-bed MWCs have capacities ranging from 184 to 920 metric tons per day. 

These systems are usually equipped with boilers to produce steam. 

Figure 3-6. Fluidized-bed refuse-derived fuel furnace. 

Source: U.S. EPA (1997a). 

3-8
3.1.1.2. Air Pollution Control Devices 

MWCs are commonly equipped with one or more post-combustion APCDs to remove 

various pollutants such as PM, heavy metals, acid gases, and organic contaminants prior to 

release from the stack (U.S. EPA, 1992a). Types of APCDs include 

C Electrostatic precipitator 

C Fabric filter 

C Spray dry scrubbing system 

C Dry sorbent injection 

C Wet scrubber 

Electrostatic precipitator (ESP). The ESP is generally used to collect and control PM 

that evolves during MSW combustion by introducing a strong electrical field into the flue gas 

stream; this in turn charges the particles entrained in the combustion gases (Donnelly, 1992). 

Large collection plates receive an opposite charge to attract and collect the particles. CDD/CDF 

formation can occur within the ESP at temperatures in the range of 150°C to about 350°C. As 

temperatures at the inlet to the ESP increase from 150 to 300°C, CDD/CDF concentrations have 

been observed to increase by approximately a factor of 2 for each 30°C increase in temperature 

(U.S. EPA, 1994a). As the temperature increases beyond 300°C, formation rates decline. 

Although ESPs in this temperature range efficiently remove most particulates and the 

associated CDDs/CDFs, the CDD/CDF formation that does occur can result in a net increase in 

CDD/CDF emissions. This temperature-related formation of CDDs/CDFs within the ESP can be 

applied, for purposes of this report, to distinguish cold-sided ESPs, which operate at or below 

230°C, from hot-sided ESPs, which operate at an inlet temperature greater than 230°C. Most 

ESPs have been replaced with better-performing and lower-cost fabric filter technology. 

Fabric filter (FF). FFs are also PM control devices that remove dioxins associated with 

particles and any vapors that adsorb to the particles. The filters are usually 6- to 8-inch-diameter 

bags, 30 feet long, made from woven fiberglass material, and arranged in series. An induction 

fan forces the combustion gases through the tightly woven fabric. The porosity of the fabric 

allows the bags to act as filter media and retain a broad range of particle sizes (down to less than 

1 :m in diameter). The FF is sensitive to acid gas; therefore, it is usually operated in 

combination with spray dryer (SD) adsorption of acid gases. 

Spray dry scrubbing system (SDSS). Spray dry scrubbing, also called SD adsorption, 

involves the removal of both acid gas and PM from the post-combustion gases. By themselves, 

the units probably have little effect on dioxin emissions. In a typical SDSS, hot combustion 

gases enter a scrubber reactor vessel. An atomized hydrated lime slurry (water plus lime) is 

injected into the reactor at a controlled velocity (Donnelly, 1992). The slurry rapidly mixes with 

3-9
the combustion gases within the reactor. The water in the slurry quickly evaporates, and the heat 

of evaporation causes the combustion gas temperature to rapidly decrease. The neutralizing 

capacity of hydrated lime reduces the acid gas constituents of the combustion gas (e.g., HCl and 

SO2) by greater than 70%. A dry product consisting of PM and hydrated lime settles to the 

bottom of the reactor vessel. 

SDSS technology is used in combination with ESPs or FFs. SDSSs reduce ESP inlet 

temperatures to make a cold-sided ESP. In addition to acid gas, particulate, and metals control, 

SDSSs with FFs or ESPs achieve greater than 90% dioxin control (U.S. EPA, 1992a), and they 

typically achieve greater than 90% SO2 and HCl control. 

Dry sorbent injection (DSI). DSI is used to reduce acid gas emissions. As with SDSSs, 

these units by themselves probably have little effect on dioxin emissions. In this system, dry 

hydrated lime or soda ash is injected directly into the combustion chamber or into the flue duct of 

the hot post-combustion gases. In either case, the reagent reacts with and neutralizes the acid gas 

constituents (Donnelly, 1992). 

Wet scrubber (WS). WS devices are designed for acid gas removal and are more 

common to MWC facilities in Europe than in the United States. They should help reduce 

emissions of dioxin in both vapor and particle forms. The devices consist of two-stage 

scrubbers. The first stage removes HCl, and the second stage removes SO2 (Donnelly, 1992). 

Water is used to remove HCl, and caustic or hydrated lime is added to remove SO2 from the 

combustion gases. 

Other types of APCDs. In addition to the APCDs described above, some less common 

types are also used in some MWCs. An example is activated carbon injection (CI) technology. 

Activated carbon is injected into the flue gas prior to the gas reaching SDSSs with FFs (or an 

ESP). Dioxin (and mercury) are absorbed onto the activated carbon, which is then captured by 

the FFs or ESP. CI technology improves dioxin control technologies by an additional 75% and 

is commonly referred to as flue gas polishing. Many APCDs have been retrofitted to include CI, 

including more than 120 large MWCs. 

3.1.1.3. Classification Scheme 

Based on the array of municipal waste combustion technologies described above, a 

classification system for deriving CDD/CDF emission estimates was developed. Assuming that 

facilities with common design and operating characteristics have a similar potential for 

CDD/CDF emissions, the MWCs operating in 1987 and 1995 were divided into categories 

according to the eight furnace types and six APCDs described above. This resulted in 17 design 

classes in 1987 and 40 design classes in 1995. Because fewer types of APCDs were used in 1987 

3-10
than in 1995, fewer design classes are needed for estimating emissions. The MWCs operating in 

2000 were divided into three furnace types and 12 APCDs, resulting in 36 design classes. 

Design classes for all three reference years are summarized in Figures 3-7 through 3-9. 

3.1.2. Characterization of MWCs in Reference Years 2000, 1995, and 1987 

Table 3-1 lists, by design/APCD type, the number of facilities and activity level (kg 

MSW incinerated/yr) for MWCs in reference year 2000. Similar inventories are provided for 

reference years 1995 and 1987 in Tables 3-2 and 3-3, respectively. This information was 

derived from five reports: U.S. EPA (1987b), SAIC (1994), Taylor and Zannes (1996), Solid 

Waste Technologies (1994), and a memoran dum dated March 27, 2003, from Jason Huckaby, 

ERG, to Walt Stevenson, U.S. EPA. In general, the information was collected via telephone 

interviews with the plant operators. 

Figure 3-7. Municipal waste combustor design classes for 1987. 

DS/FF = Dry scrubber combined with a fabric filter 

EGB = Electro gravel bed 

FF = Fabric filter 

H-ESP = Hot-sided electrostatic precipitator (temperature at control device is >230°C) 

WS = Wet scrubber 

3-11
Figure 3-8. Municipal waste combustor design classes for 1995. 

C-ESP = Cold-sided electrostatic precipitator (temperature at control device is <230°C)

DS/CI/FF = Dry scrubber with carbon injection and fabric filter

DS/FF = Dry scrubber combined with a fabric filter

DSI/FF = Dry sorbent injection coupled with a fabric filter

EGB = Electro gravel bed

H-ESP = Hot-sided electrostatic precipitator (temperature at control device is >230°C)

WS = Wet scrubber

Using Tables 3-1, 3-2, and 3-3, a number of comparisons can be made between the 

reference years: 

• The number of facilities stayed about the same (105 in 1987, 130 in 1995, and 104 in 

2000), but the amount of MSW incinerated more than doubled from 1987 to 1995 

(from 13.4 billion kg in 1987 to 29 billion kg in 1995); it remained constant from 

1995 to 2000 (30.6 billion kg in 2000). 

3-12
Municipal Solid Waste Incinerator Design Classes for 2000 

SD/FF/CI 

SD/FF/CI/SNCR 

SD/FF/SNCR 

SD/ESP/CI 

SD/FF/SNCR 

SD /FF/CI/SNCR 

SD/FF 

DSI/FF 

ESP 

FF 

Waterwall 

SD/FF 

DSI/FF 

SD/FF/CI/SNCR 

Rotary Kiln Combuster 

Mass Burn 

ESP 

FF 

SD/FF 

WS/ESP 

Modular, Excess air 

DSI/ESP 

DSI/FF 

ESP 

FF 

SD/FF/CI 

WS 

Modular, Starved air 

DSI/FF 

SD/ESP 

SD/FF 

SD/FF/CI/SNCR 

DSI/FF/H2O/SNCR 

SD/ESP/FF/CI 

SD/FF /SNCR 

SD/FF/SNCR 

ESP 

Refuse-Derived Fuel 

b 

c 

a 

Figure 3-9. Municipal waste combustor design classes for 2000. 

a

Also equipped with furnace dry sorbent injection system. b

Also equipped with flue gas cooling (280–290°F). c

Also equipped with compact hybrid particulate collector system.

 CI = Carbon injection

 DSI = Dry sorbent injection

 ESP = Electrostatic precipitator

 FF = Fabric filter

 H2O = Water scrubber

 SD = Spray dryer

 SNCR = Selective noncatalytic reduction

 WS = Wet scrubber 

• In terms of number of facilities, the dominant furnace technology shifted from 

modular in 1987 (49 units and 1.1 billion kg) to MB-WW facilities in 1995 (57 units 

and 16.8 billion kg) and 2000 (140 units and 19 billion kg). 

• The dominant APCD technology shifted from hot-sided ESPs in 1987 (46 units and 

11 billion kg) to SDs with FFs (SDs/FFs) in 1995 (41 units and 12.8 billion kg) and 

SDs/FFs with CI and selective noncatalytic reduction (88 units and 4.6 billion kg), 

and SDs/FFs (27 units and 3.1 billion kg) in 2000. 

• The use of hot-sided ESPs dropped from 46 facilities in 1987 (11 billion kg) to 12 

facilities in 1995 (1.3 billion kg). No hot-sided ESPs were used in 2000. 

• The number of uncontrolled facilities dropped from 38 in 1987 (0.6 billion kg) to 10 

in 1995 (0.2 billion kg) and 6 in 2000 (0.08 billion kg). 

3-13
Table 3-1. Inventory of municipal waste combustors (MWCs) in 2000 by technology, air pollution c

device (APCD), sizea, and annual activity level (kg/yr) 

3-14

APCDb 

MWC type 

MB/RC MB MB/WW/RC MB/REF MB/WW MOD/EA MOD/SA FB/RDF 

Size 

(N) 

Activity 

level 

Size 

(N) 

Activity 

level 

Size 

(N) 

Activity 

level 

Size 

(N) 

Activity 

level 

Size 

(N) 

Activity 

level 

Size 

(N) 

Activity 

level 

Size 

(N) 

Activity 

level 

Size 

(N) 

Activ

leve

DSI/FFc S(4) 2.78e+08 S(2) 6.27e+07 S(3) 1.13e+08 S(3) 4.7e+07 S(2) 9.41e+

DSI/ESP S(2) 1.25e+08 S(7) 1.49e+08 DSI/FF/ H2O/SNCR 

ESP S(4) 2.61e+08 S(6) 2.79e+08 S(6) 9.22e+07 S(8) 1.14e+08 FF S(1) 1.76e+07 S(2) 3.14e+07 WS S(4) 5.46e+07 WS/ESP S(3) 1.13e+08 SDd/FF/CI/ 

SNCR 

L(3) 3.97e+08 L(75) 1.34e+10 

SD/ESP SD/ESP/CI L(4) 3.74e+08 

SD/ESP/ CI/SNCR 

L(15) 2.79e+09 

SD/ESP/ FFe/CI 

SD/FF L(6) 1.11e+09 S(2) 6.27e+07 S(4) 1.25e+08 S(3) 1.32e+08 SD/FFe/ SNCR 

L(13) 2.99e+09 

SD/FF/CI S(2) 1.25e+08 S(4) 6.27e+07 

L(5) 9.51e+08 

Unc S(2) 3.14e+07 S(4) 5.17e+07 TOTAL L(9) 

S(8) 

1.51e+09 4.94e+08 

S(2) 6.27e+07 S(1) 1.76e+07 S(2) 1.25e+08 L(124) 

S(16) 

2.05e+10 6.24e+08 

S(15) 4.50e+08 S(31) 5.10e+08 S(2) 9.41e+
3-15

Table 3-1. Inventory of municipal waste combustors (MWCs) in 2000 by technology, air pollution cdevice (APCD), sizea, and annual activity level (kg/yr) (continued) 

aFor size, S = small; L = large.bSlash(es) indicates devices used in co njunction.cAlso equipped with flue gas cooling ( 138 to 143°C).dAlso equipped with furnace dry sorbe nt injection system.eAlso equipped with compact hybrid p articulate collector system.

APCD: CI = Carbon injection DSI = Dry sorbent injection ESP = Electrostatic precipitator FF = Fabric filter H 2O = Water scrubber

 SD = Spray dryer SNCR = Selective noncatalytic reduction Unc = Uncontrolled WS = Wet scrubberMWC type: FB/RDF = Fluidized-bed refuse-derived fuel MB = Mass burn MB/RC = Mass burn rotary kiln MB/REF = Mass burn refractory walled  MB/WW/RC = Mass burn waterwalled/refractory walled MOD/EA = Modular excess air RDF = Refuse-derived fuel
Table 3-2. Inventory of municipal waste combustors (MWCs) in 1995 by technology, air pollution c

(APCD), and annual activity level (kg/yr) 

3-16

APCDa 

MWC type 

MB/RC MB/REF MB/WW FB/RDF RDF/ded MOD/SA MOD/

N 

Activity 

level N Activity level N Activity level N Activity level N Activity level N Activity level N Activ

 Unc 9 1.87e+08 1 1.4 H-ESP 6 1.04e+09 1 4.22e+07 4 1.82e+08 1 1.9 C-ESP 2 2.00e+08 1 1.69e+08 8 2.81e+09 4 1.81e+09 4 1.25e+08 3 8.2 DSI/H-ESP 1 4.22e+08 1 2.00e+08 1 1.4 DS/FF 2 1.14e+09 2 2.68e+08 28 8.57e+09 1 1.69e+08 7 2.51e+09 1 1.1 DS/CI/FF 3 1.17e+09  DS/FF/C-ESP 1 5.63e+08  WS/FF 1 2.82e+07  WS/C-ESP 1 6.7 DS/C-ESP 1 4.22e+08 8 2.31e+09 4 1.75e+09  DS/DSI/C-ESP 1 7.60e+07  DSI/CI/H-ESP 1 2.75e+08  DSI/C-ESP 6 5.07e+08  DSI/FF 2 2.59e+08 1 1.13e+08 2 1.97e+08 1 8.45e+07 1 4.22e+08 1 3.42e+07 1 1.0 DSI/EGB 1 1.13e+08  WS 2 2.04e+08 3 4.90e+07  TOTAL 12 2.10e+09 7 1.18e+09 57 1.68e+10 3 3.67e+08 19 7.30e+09 23 6.81e+07 9 4.1
3-17

Table 3-2. Inventory of municipal waste combustors (MWCs) in 1995 by technology, air pollution c(APCD), and annual activity level (kg/yr) (continued) 

aSlash(es) indicates devices used in conjunction.

 APCD: C-ESP = Cold-sided electrostatic precipitator CI = Carbon injection DS = Dry scrubber DSI = Dry sorbent injection EGB = Electro gravel bed FF = Fabric filter H-ESP = Hot-sided electrostatic precipitator SD = Spray dryer Unc = Uncontrolled WS = Wet scrubber MWC type: FB/RDF = Fluidized-bed refuse-derived fuel MB/RC = Mass burn rotary kiln MB/REF = Mass burn refractory walled MB/WW = Mass burn waterwalled MOD/EA = Modular excess air MOD/SA = Modular starved air RDF/ded = Refuse-derived fuel/dedicated 
Table 3-3. Inventory of municipal waste combustors (MWCs) in 1987 by technology, air pollution c

device (APCD), and annual activity level (kg/yr) 

3-18

APCDa 

MWC type 

MB/RC MB/REF MB/WW RDF/ded RDF/cofired MOD/SA MOD/

N 

Activity 

level N Activity level N 

Activity 

level N 

Activity 

level N Activity level N Activity level N Acti

 Unc 36 5.73e+08 2 4. H- ESP 3 3.94e+08 12 2.00e+09 19 5.20e+09 7 3.01e+09 3 2.53e+08 2 1.17e+08  DS/FF 1 1.41e+07 1 1.55e+08  FF 1 1.58e+07 3 1.43e+08  EGB 1 6. WS 7 9.01e+08 2 3.38e+08 4 5.30e+07 1 1. TOTAL 4 4.10e+08 20 2.92e+09 20 5.36e+09 9 3.35e+09 3 2.53e+08 45 8.9e+08 4 2.

aSlash indicates devices used in conjunction.

 APCD: DS = Dry scrubber EGB = Electro gravel bed FF = Fabric filter H-ESP = Hot-sided electrostatic precipitator Unc = Uncontrolled WS = Wet scrubber MWC type: MB/RC = Mass burn rotary kiln MB/REF = Mass burn refractory walled MB/WW = Mass burn waterwalled MOD/EA = Modular excess air MOD/SA = Modular starved air RDF/cofired = Refuse-derived fuel/cofired RDF/ded = Refuse-derived fuel/dedicated 
3.1.3. Estimation of CDD/CDF Emissions from MWCs 

Compared with other CDD/CDF source categories, MWCs have been more extensively 

evaluated for CDD/CDF emissions. In 2000, due to new regulations, EPA’s Office of Air 

Quality Planning and Standards (OAQPS) obtained emission test reports for all large MWCs. 

3.1.3.1. Estimating CDD/CDF Emissions from MWCs in Reference Year 2000 

EPA regulates dioxin emissions from MWCs on the basis of whether a facility is large or 

small (see Section 3.1.7). Large facilities are those with a total design combustion capacity >250 

tons/day; small facilities are those with a combustion capacity <250 tons/day). Combustion 

capacity is determined on the basis of a single incineration unit. Facilities may comprise more 

than one incinerator unit. In 2000, all large MWC facilities were source tested for stack 

emissions of dioxin, as required by EPA regulations. Therefore, actual emissions from large 

facilities were used to represent facility-specific dioxin emissions in 2000. More than 70% of the 

small MWCs have been tested for dioxin emissions, and the available data were used to represent 

facility-specific dioxin emissions in 2000. For small MWCs that were not source tested, dioxin 

emissions were estimated on the basis of emission factors. 

Using the test reports, concentrations and emissions were calculated for each of the 17 

named dioxin/furan congeners and the remainder of the congener groups (homologues), making 

up total dioxin/furan emissions (for 27 congeners/groups) for each of the MWC units 

(Memorandum dated March 27, 2003, from Jason Huckaby, ERG, to Walt Stevenson, U.S. 

EPA). The calculations were based on the individual CDD/CDF congener/group concentrations 

for the MWC, the flue gas flow rate and MWC steam generation rate during the test, and the 

annual steam generation at the MWC. Table 3-4 presents the average CDD/CDF congenerspecific stack gas concentrations used to derive the emission factors for the nontested facilities. 

Table 3-4 shows concentrations for three detection limit (DL) assumptions: (1) a value of zero 

for concentrations below the DL, (2) a value of one-half the DL for concentrations below the DL, 

and (3) a value of the DL for concentrations below the DL. 

Table 3-5 shows facility-specific dioxin emissions for all MWCs operating in 2000 

(because of its size, Table 3-5 is placed at the end of this chapter). Note that the facilities are 

divided into large and small MWCs. In total, 83.8 g TEQDF-WHO98 (76.3 g I-TEQ) were emitted 

from all 104 MWCs in 2000. Of this total amount, 13.8 g TEQDF-WHO98 (12.7 g I-TEQ) were 

emitted from large MWCs and 69.9 g TEQDF-WHO98 (63.6 g I-TEQ) were emitted from small 

MWCs. Although 91% of the MSW combusted in 2000 was incinerated in large MWCs, the 

large MWCs accounted for only 17% of total dioxin emissions. 

3-19
Table 3-4. National average CDD/CDF congener concentrations for 

large municipal waste combustors (ng/dscm @ 7% O2) 

Congener Nondetect set to zeroa 

Nondetect set to 

1⁄2 detection limita 

Nondetect set to 

detection limita

 TrCDD 

 2,3,7,8-TCDD 

 1,2,3,7,8-PeCDD 

 1,2,3,4,7,8-HxCDD 

 1,2,3,6,7,8-HxCDD 

 1,2,3,7,8,9-HxCDD 

 1,2,3,4,6,7,8-HpCDD 

 OCDD 

 Other TCDDs 

 Other PeCDDs 

 Other HxCDDs 

 Other HpCDDs 

0.031 

0.005 

0.016 

0.016 

0.037 

0.032 

0.219 

0.345 

0.232 

0.323 

0.494 

0.220 

0.031 

0.006 

0.016 

0.016 

0.036 

0.032 

0.219 

0.345 

0.239 

0.334 

0.502 

0.220 

0.031

0.006

0.017

0.016

0.037

0.032 

0.219

0.345

0.246

0.344 

0.510

0.220

 OCDF 

 Other TCDF 

 Other PeCDF 

 Other HxCDF 

 Other HpCDF 

 TrCDF 

 2,3,7,8-TCDF 

 1,2,3,7,8-PeCDF 

 2,3,4,7,8-PeCDF 

 1,2,3,4,7,8-HxCDF 

 1,2,3,6,7,8-HxCDF 

 1,2,3,7,8,9-HxCDF 

 2,3,4,6,7,8-HxCDF 

 1,2,3,4,6,7,8-HpCDF 

 1,2,3,4,7,8,9-HpCDF 

0.090 

1.081 

0.747 

0.326 

0.079 

0.037 

0.072 

0.050 

0.069 

0.082 

0.059 

0.013 

0.066 

0.156 

0.024 

0.092 

1.083 

0.758 

0.329 

0.079 

0.037 

0.072 

0.051 

0.069 

0.083 

0.059 

0.013 

0.066 

0.157 

0.024 

0.094

1.085 

0.770

0.332

0.079

0.037 

0.073

0.052

0.069

0.083

0.060

0.014

0.067

0.159

0.024 

 TOTAL 4.92 4.97 5.02 a

Values incorporating use of the detection limit when the laboratory report indicated “not detected” for individual

 CDD/CDF congeners. 

Source: Memorandum dated March 27, 2003, from Jason Huckaby, ERG, to Walt Stevenson, U.S. EPA. 

3-20
_____________ 

3.1.3.2. Estimating CDD/CDF Emissions from MWCs in Reference Years 1995 and 1987 

Within the context of this report, adequate emissions testing for CDDs/CDFs was 

available for 11 of the 113 facilities in the 1987 inventory and 27 of the 130 facilities in the 1995 

inventory. Nationwide CDD/CDF air emissions from MWCs for reference years 1987 and 1995 

were estimated using the three-step process described below. 

Step 1. Estimation of emissions from all stack-tested facilities. The EPA stack testing 

method (EPA Method 23) produces a measurement of CDDs/CDFs in units of mass 

concentration of CDD/CDF (nanograms per dry standard cubic meter [ng/dscm] of combustion 

gas) at standard temperature and pressure (20°C and 1 atmosphere [atm]) and adjusted to a 

measurement of 7% oxygen in the flue gas (U.S. EPA, 1995a). This concentration is assumed to 

represent conditions at the point of release from the stack into the air. Equation 3-1 was used to 

derive annual emission estimates for each tested facility: 

C × V × CF × H 

ETEQ = (3-1) 

109 ng/g 

where: 

ETEQ = annual TEQ emissions (g/yr) 

C = combustion flue gas TEQ concentration (ng/dscm) (20oC, 1 atm; adjusted to 7% O2) 

V = volumetric flow rate of combustion flue gas (dscm/hr) (20oC, 1 atm; adjusted to 

7% O2) 

CF = capacity factor; fraction of time that the MWC operates (0.85) 

H = total hours in a year (8,760 hr) 

After calculating annual emissions for each tested facility, the emissions were summed 

across all tested facilities for each reference year. (Many of the emission tests do not correspond 

exactly to these two years. In these cases, the equipment conditions present at the time of the test 

were compared with those during the reference year to determine their applicability.) 

Step 2. Estimation of emissions from all nonstack-tested facilities. This step involves 

multiplying the emission factor and the annual activity level for each MWC design class and then 

summing across classes. The activity levels for reference years 1995 and 1987 are summarized 

in Tables 3-2 and 3-3, respectively. The emission factors were derived by averaging the emission 

factors across each tested facility in a design class. The emission factor for each facility was 

calculated using the following equation: 

3-21
_____ __________ _________ 

 C × Fv ______ EFMWC = (3-2) 

Iw 

where: 

EFMWC = emission factor; average ng TEQ/kg of waste burned 

C = TEQ or CDD/CDF concentration in flue gases (ng TEQ/dscm) (20°C, 

 1 atm; adjusted to 7% O2) 

Fv = volumetric flue gas flow rate (dscm/hr) (20°C, 1 atm; adjusted to 7% O2) 

Iw = average waste incineration rate (kg/hr) 

Using an MB-WW MWC equipped with a cold-sided ESP as an example, and given: 

C = 10 ng TEQ/dscm (20°C, 1 atm; adjusted to 7% O2)

Fv = 40,000 dscm/hr (20°C, 1 atm; adjusted to 7% O2)

Iw = 10,000 kg MSW/hr, then

10 ng 40,000 dscm hr 

EFMB-WW = × × 

dscm hr 10,000 kg

 40 ng TEQ 

= _____________ 

kg MSW burned 

EPA was not able to obtain engineering test reports of CDD/CDF emissions for a number 

of design classes. In these cases, the above procedure could not be used to derive emission 

factors. Instead, the emission factors of the tested design class that was judged most similar in 

terms of dioxin control was assumed to apply to the untested class. The following logic was used 

to make this decision: 

1. The tested APCDs for the furnace type of the untested class were reviewed to see 

whether any operated at a similar temperature. 

2. If any operated at similar temperatures, the one with the most similar technology was 

assumed to apply. 

3-22
3. If none operated at a similar temperature, then the most similar furnace type with the 

same control device was assumed to apply. 

Table 3-6 lists all design categories with no tested facilities and shows the class with 

tested facilities that was judged to be most similar. The emission factors for each design class 

are the same for both reference years because the emission factor is determined only by the 

design and operating conditions and is independent of the year of the test. 

Step 3. Summation of emissions from tested and untested facilities. This step involves 

summing emissions from all tested and untested facilities. This process is shown in Tables 3-7a 

and 3-7b and 3-8a and 3-8b for the reference years 1995 and 1987, respectively. The tables are 

organized by design class and show the emission estimates for the tested and untested facilities 

separately. The calculation of emissions from untested facilities is broken out to show the 

activity level and emission factor for each design class. 

3.1.4. Summary of CDD/CDF (TEQ) Emissions from MWCs for 2000, 1995, and 1987 

The activity level estimates (i.e., the amount of MSW that is annually combusted by the 

various municipal waste combustion technologies) are given a high confidence rating for 1987 

(13.4 billion kg of waste), 1995 (29.3 billion kg of waste), and 2000 (30.6 billion kg of waste). 

For all three years, independent sources conducted comprehensive surveys of activity levels for 

virtually all facilities (U.S. EPA, 1987b; Solid Waste Technologies, 1994; SAIC, 1994; Taylor 

and Zannes, 1996; Memorandum dated March 27, 2003, from Jason Huckaby, ERG, to Walt 

Stevenson, U.S. EPA). 

The emission factor estimates are given a high confidence rating for 2000 and a medium 

confidence rating for both 1995 and 1987. A large number of MWC facilities were tested in 

2000, whereas a moderate fraction of the facilities were tested in 1995 and 1987: 27 of 130 

facilities (21%) in 1995 and 11 of 104 facilities in 1987 (10%). The tested facilities represented 

27 and 21% of the total activity level of operating MWCs in 1995 and 1987, respectively. These 

tests represent most of the design categories identified in this report. The emission factors were 

developed from emission tests that followed standard EPA protocols, used strict QA/QC 

procedures, and were well documented in engineering reports. Because all tests were conducted 

under normal operating conditions, some uncertainty exists about the magnitude of emissions 

that may have occurred during other conditions (e.g., upset conditions, start up, and shut down). 

In summary, this report estimates the following dioxin releases to air from MWCs operating in 

2000, 1995 and 1987. 

3-23
Table 3-6. CDD/CDF TEQ emission factors (ng TEQ/kg waste) for

municipal solid waste incineration 

Incinerator 

design 

Air pollution

control device 

(APCD)a 

Average ITEQDF 

emission factor 

Average

TEQDF-WHO98 

emission factor Basis and rationale 

MOD/SA C-ESP 

DS/DSI/

C-ESP 

DSI/FF 

FF 

H-ESP 

UNC 

WS 

WS/FF 

16.2 

16.2 

0.025 

16.2 

79 

0.025 

16.2 

16.2 

17 

17 

0.024 

17 

85.7 

0.024 

17 

17 

Based on MOD/EA; C-ESP, similar furnace

(modular design) and same APCD 

Based on MOD/EA; C-ESP, similar furnace

(modular design) and similar emission control 

Based on direct tests 

Based on MOD/EA; C-ESP, similar furnace

(modular design) and similar emission control 

Based on direct tests 

Based on MOD/SA; DSI/FF, same furnace and

most similar expected emissions 

Based on MOD/EA; C-ESP, similar furnace

(modular design) and similar APCD 

temperature 

Based on MOD/EA; C-ESP, similar furnace

(modular design) and similar APCD 

temperature 

MOD/EA C-ESP 

DS/FF 

DSI/FF 

DSI/H-ESP 

EGB 

H-ESP 

Unc 

WS 

WS/C-ESP 

16.2 

16.2 

0.025 

118 

0.025 

118 

0.025 

16.2 

16.2 

17 

17 

0.024 

119 

0.024 

119 

0.024 

17 

17 

Based on direct tests 

Based on MOD/EA; C-ESP, same furnace and

similar temperature in APCD; may

overestimate emissions 

Based on MOD/SA; DSI/FF, similar (modular

design) furnace and same APCD 

Based on MOD/EA; H-ESP, same furnace and

similar emissions 

Based on MOD/SA; DSI/FF, same furnace and

most similar expected emissions 

Based on direct tests 

Based on MOD/SA; DSI/FF, same furnace and

most similar expected emissions 

Based on MOD/EA; C-ESP, same furnace and

similar APCD temperature 

Based on MOD/EA; C-ESP, same furnace and

similar APCD 

FB/RDF DS/FF 

DSI/EGB 

DSI/FF 

0.63 

0.63 

0.63 

0.72 

0.72 

0.72 

Based on MB/WW; DS/FF similar furnace and

same APCD 

Based on MB/WW; DS/FF similar furnace;

may underestimate emissions 

Based on MB/WW; DS/FF similar furnace;

may underestimate emissions a

Slash indicates devices used in conjunction.

 APCD: MWC type:

 C-ESP = Cold-sided electrostatic precipitator FB/RDF = Fluidized-bed refuse-derived fuel

 DS = Dry scrubber MB/WW = Mass burn waterwalled

 DSI = Dry sorbent injection MOD/EA = Modular excess air

 EGB = Electro gravel bed MOD/SA = Modular starved air

 FF = Fabric filter

 H-ESP = Hot-sided electrostatic precipitator

 Unc = Uncontrolled

 WS = Wet scrubber 

3-24
Table 3-7a. Annual I-TEQDF emissions from municipal waste combustors 

(MWCs) operating in 1995 

MWC 

type 

Air pollution 

control device 

(APCD)a 

I-TEQDF 

emissions from

tested facilities 

(g TEQ/yr) 

Average 

 I-TEQDF 

emission factor 

(ng/kg) 

Activity 

level 

nontested 

facilities 

(kg/yr) 

I-TEQDF 

emissions 

from 

nontested 

facilities 

(g TEQ/yr) 

Total 

I-TEQDF 

emissions 

from all 

facilities 

(g TEQ/yr) 

MB/WW C-ESP 

 DS/C-ESP 

 DS/CI/FF 

DS/FF 

 DSI/CI/H-ESP 

DSI/FF 

 DSI/H-ESP 

 H-ESP 

0 

2.1 

0.64 

2 

2.1 

0.3 

0 

163 

6.1 

6.1 

1.5 

0.63 

– 

– 

7.74 

473 

2.81e+09 

1.88e+09 

7.44e+08 

5.98e+09 

0 

0 

4.22e+08 

1.79e+08 

17.1 

11.5 

1.1 

3.8 

0.0 

0.0 

3.3 

84.7 

17.1

13.6

1.7

5.8

2.1

0.3

3.3

247.6

 Subtotal 170.1 1.20e+10 121.4 291.5 

MB/REF C-ESP 

 DS/C-ESP 

DS/FF 

DSI/FF 

WS 

39.8 

21.6 

0 

0 

0 

– 

– 

0.63 

1.91 

236 

0 

0 

2.68e+08 

1.13e+08 

2.04e+08 

0 

0 

0.2 

0.2 

48.1 

39.8

21.6

0.2

0.2

48.1

 Subtotal 61.4 5.85e+08 48.5 109.9 

MB/RC C-ESP 

DS/FF 

 DSI/C-ESP 

DSI/FF 

0 

0.25 

0 

5.3 

47.0 

0.65 

47.0 

47.0 

2.00e+08 

7.57e+08 

5.07e+08 

1.46e+08 

9.4 

0.5 

23.8 

6.9 

9.4

0.8

23.8

12.2

 Subtotal 5.6 1.61e+09 40.6 46.2 

RDF/ded C-ESP 

 DS/C-ESP 

DS/FF 

DSI/FF 

 DSI/H-ESP 

 H-ESP 

 DS/FF/C-ESP 

32.5 

0.3 

0.1 

0 

0 

0 

0 

231 

0.53 

0.24 

231 

231 

1,492 

0.24 

1.67e+09 

1.14e+09 

1.58e+09 

4.22e+08 

2.00e+08 

4.22e+07 

5.63e+08 

385.8 

0.6 

0.4 

97.5 

46.2 

63.0 

0.1 

418.3

0.9

0.5

97.5

46.2

63

0.1

 Subtotal 32.9 5.62e+09 593.5 626.4 

MOD/SA C-ESP 

DSI/FF 

 H-ESP 

 Unc 

WS 

WS/FF 

 DS/DSI/C-ESP 

0 

0 

8 

0 

0 

0 

0 

16.2 

– 

79 

0.025 

16.2 

16.2 

16.2 

1.25e+08 

0 

8.03e+07 

1.87e+08 

4.90e+07 

2.82e+07 

7.60e+07 

2.0 

0.0 

6.3 

0.0 

0.8 

0.5 

1.2 

2

0

14.3

0.005

0.8

0.5

1.2

 Subtotal 8 5.46e+08 10.9 18.9 

3-25
Table 3-7a. Annual I-TEQDF emissions from municipal waste combustors 

(MWCs) operating in 1995 (continued) 

MWC 

type 

Air pollution 

control device 

(APCD)a 

I-TEQDF 

emissions from

tested facilities 

(g TEQ/yr) 

Average 

 I-TEQDF 

emission factor 

(ng/kg) 

Activity 

level 

nontested 

facilities 

(kg/yr) 

I-TEQDF 

emissions 

from 

nontested 

facilities 

(g TEQ/yr) 

Total 

I-TEQDF 

emissions 

from all 

facilities 

(g TEQ/yr) 

MOD/EA C-ESP 

DS/FF 

DSI/FF 

 DSI/H-ESP 

 H-ESP 

 Unc 

 WS/C-ESP 

0.1 

0 

0 

0 

2.3

0 

0 

16.2 

16.2 

0.025 

118 

b 

0.025 

16.2 

6.25e+07 

1.18e+08 

1.01e+08 

1.41e+07 

0 

1.41e+07 

6.76e+07 

1 

1.9 

0.003 

1.7 

0 

0.0003 

1.1 

1.1

1.9

0.003

1.7

2.3

0.0003

1.1

 Subtotal 2.4 3.77e+08 5.7 8.1 

FB/RDF DS/FF 

DSI/EGB 

DSI/FF 

0 

0 

0 

0.63 

0.63 

0.63 

1.69e+08 

1.13e+08 

8.45e+07 

0.1 

0.1 

0.1 

0.1

0.1

0.1

 Subtotal 0 3.67e+08 0.3 0.3 

TOTAL 280.4 2.11e+10 820.9 1,101.3 a

Slash indicates devices used in conjunction. b

Value could not be calculated. 

– = Emissions not developed

 APCD:

 C-ESP = Cold-sided electrostatic precipitator

 CI = Carbon injection

 DS = Dry scrubber

 DSI = Dry sorbent injection

 EGB = Electro gravel bed

 FF = Fabric filter

 H-ESP = Hot-sided electrostatic precipitator

 Unc = Uncontrolled

 WS = Wet scrubber

 MWC type:

 FB/RDF = Fluidized-bed refuse-derived fuel

 MB/RC = Mass burn rotary kiln

 MB/REF = Mass burn refractory walled

 MB/WW = Mass burn waterwalled

 MOD/EA = Modular excess air

 MOD/SA = Modular starved air

 RDF/ded = Refuse-derived fuel/dedicated 

3-26
Table 3-7b. Annual TEQDF-WHO98 emissions from municipal waste 

combustors (MWCs) operating in 1995 

MWC 

type 

Air pollution 

control device 

(APCD)a 

TEQDF ­

WHO98 

emissions 

from tested 

facilities 

(g TEQ/yr) 

Average 

TEQDF ­

WHO98 

emission 

factor 

(ng/kg) 

Activity level 

nontested 

facilities 

(kg/yr) 

TEQDF ­

WHO98 

emissions 

from 

nontested 

facilities 

(g TEQ/yr) 

Total TEQDF ­

WHO98 

emissions 

from all 

facilities 

(g TEQ/yr) 

MB/WW C-ESP

 DS/C-ESP

 DS/CI/FF

 DS/FF

 DSI/CI/H-ESP

 DSI/FF

 DSI/H-ESP

 H-ESP

 0

 2.24

 0.68

 2.1

 2.26

 0.3

 0

 183

 6.54 

6.54 

1.61 

0.72 

– 

– 

8.22 

535.0 

2.81e+09 

1.88e+09 

7.44e+08 

5.98e+09 

0.0 

0.0 

4.22e+08 

1.79e+08 

18.4 

12.3 

1.2 

4.3 

0.0 

0.0 

3.5 

95.8 

18.4

14.54

1.88

6.4

2.26

0.3

3.5

278.8

 Subtotal 190.6 1.20e+10 135.4 326.0 

MB/REF C-ESP

 DS/C-ESP

 DS/FF

 DSI/FF

 WS

 43

 22.5

 0

 0

 0

 –

 – 

0.72 

2.07 

254.0 

0.0 

0.0 

2.68e+08 

1.13e+08 

2.04e+08 

0.0 

0.0 

0.2 

0.2 

51.8 

43.0

22.5

0.2

0.2

51.8

 Subtotal 65.5 5.85e+08 52.2 117.7 

MB/RC C-ESP

 DS/FF

 DSI/C-ESP

 DSI/FF

 0

 0.265

 0

 10.5

 93.1 

0.68 

93.1 

93.1 

2.00e+08 

7.57e+08 

5.07e+08 

1.46e+08 

18.6 

0.5 

47.2 

13.6 

18.6

0.8

47.2

24.1

 Subtotal 10.8 1.61e+09 79.9 90.7 

RDF/ded C-ESP

 DS/C-ESP

 DS/FF

 DSI/FF

 DSI/H-ESP

 H-ESP

 DS/FF/C-ESP

 35.6

 0.34

 0.1

 0

 0

 0

 0

 253.0 

0.56 

0.26 

253.0 

253.0 

1,679.0 

253.0 

1.67e+09 

1.14e+09 

1.58e+09 

4.22e+08 

2.00e+08 

4.22e+07 

5.63e+08 

422.5 

0.6 

0.4 

106.8 

50.6 

70.9 

142.4 

458.1

1.0

0.5

106.8

50.6

70.9

142.4

 Subtotal 36 5.62e+09 794.2 830.2 

MOD/SA C-ESP

 DSI/FF

 H-ESP

 Unc

 WS

 WS/FF

 DS/DSI/C-ESP

 0

 0.0008

 8.69

 0

 0

 0

 0

 17.0 

– 

85.7 

0.024 

17.0 

17.0 

17.0 

1.25e+08 

0.0 

8.03e+07 

1.87e+08 

4.90e+07 

2.82e+07 

7.60e+07 

2.1 

0.0 

6.9 

0.0 

0.8 

0.5 

1.3 

2.1

0.001

15.6

0.005

0.8

0.5

1.3

 Subtotal 8.7 5.46e+08 11.6 20.3 

3-27
Table 3-7b. Annual TEQDF-WHO98 emissions from municipal waste 

combustors (MWCs) operating in 1995 (continued) 

MWC 

type 

Air pollution 

control device 

(APCD)a 

TEQDF ­

WHO98 

emissions 

from tested 

facilities 

(g TEQ/yr) 

Average 

TEQDF ­

WHO98 

emission 

factor 

(ng/kg) 

Activity level 

nontested 

facilities 

(kg/yr) 

TEQDF ­

WHO98 

emissions 

from 

nontested 

facilities 

(g TEQ/yr) 

Total TEQDF ­

WHO98 

emissions 

from all 

facilities 

(g TEQ/yr) 

MOD/EA C-ESP

 DS/FF

 DSI/FF

 DSI/H-ESP

 H-ESP

 Unc

 WS/C-ESP

 0.068

 0

 0

 0

 2.35

 0

 0

 17.0 

17.0 

0.024 

119.0 

– 

0.024 

17.0 

6.25e+07 

1.18e+08 

1.01e+08 

1.41e+07 

0.0 

1.41e+07 

6.76e+07 

1.06

2.01

0.002

1.68

0.0

0.003

1.15

 1.1

 2.0

 0.002

 1.7

 2.4

 0.003

 1.2

 Subtotal 2.4 3.77e+08 5.9 8.3 

FB/RDF DS/FF

 DSI/EGB

 DSI/FF

 0

 0

 0

 0.72 

0.72 

0.72 

1.69e+08 

1.13e+08 

8.45e+07 

0.114

0.076

0.057

 0.1

 0.1

 0.1

 Subtotal 0 3.67e+08 0.3 0.3 

TOTAL 314 2.11e+10 1,079.5 1,393.5 a

Slash indicates devices used in conjunction.

 – = Emissions not developed

 APCD:

 C-ESP = Cold-sided electrostatic precipitator

 CI = Carbon injection

 DS = Dry scrubber

 DSI = Dry sorbent injection

 EGB = Electro gravel bed

 FF = Fabric filter

 H-ESP = Hot-sided electrostatic precipitator

 Unc = Uncontrolled

 WS = Wet scrubber

 MWC type:

 FB/RDF = Fluidized-bed refuse-derived fuel

 MB/RC = Mass burn rotary kiln

 MB/REF = Mass burn refractory walled

 MB/WW = Mass burn waterwalled

 MOD/EA = Modular excess air

 MOD/SA = Modular starved air

 RDF/ded = Refuse-derived fuel/dedicated 

3-28
Table 3-8a. Annual I-TEQDF emissions to the air from municipal waste 

combustors (MWCs) operating in 1987 

MWC type 

Air pollution 

control 

device 

(APCD)a 

I-TEQDF 

emissions 

from tested 

facilities 

(g TEQ/yr) 

Average 

I-TEQDF 

emission 

factor 

(ng/kg) 

Activity level 

nontested 

facilities 

(kg/yr) 

I-TEQDF 

emissions 

from 

nontested 

facilities 

(g TEQ/yr) 

Total ITEQDF 

emissions 

from all 

facilities 

(g TEQ/yr) 

MB/WW DS/FF

 H-ESP

 0.0373

 433.0

 – 

473.0 

1.55e+08

3.27e+09 

0.0

1,546.7 

0.0373

1,979.7

 Subtotal 433.0 3.43e+09 1,546.7 1,979.7 

MB/REF DS/FF 0.0 0.63 1.41e+08 0.09 0.09

 H-ESP 0.0 473.0 2.00e+09 946.0 946.0

 WS 0.0 236.0 9.01e+08 212.6 212.6

 Subtotal 0.0 3.04e+09 1,158.7 1,158.7 

MB/RC FF 0.0 47.0 1.58e+07 0.7 0.7

 H-ESP 48.2 285.0 2.25e+08 64.1 112.3

 Subtotal 48.2 2.41e+08 64.8 113.0 

RDF/ded H-ESP

 WS

 840.0

 0.0

 1,492.0

 231.0

 2.45e+09 

3.38e+08 

3,655.4

 78.1

 4,495.4

 78.1

 Subtotal 840.0 2.79e+09 3,733.5 4,573.5 

RDF/cofired H-ESP 0.0 231.0 2.53e+08 58.4 58.4 

MOD/SA FF

 H-ESP

 Unc

 WS

 0.0

 0.0643

 0.0

 0.0

 16.2

 79.0

 0.025

 16.2

 1.43e+08 

3.61e+08 

5.73e+08 

5.30e+07 

2.3

 28.5

 0.01

 0.86

 2.3

 28.6

 0.01

 0.86

 Subtotal 0.0643 1.13e+09 31.7 31.8 

MOD/EA EGB 0.0 0.025 6.76e+07 0.002 0.002

 Unc 0.0 0.025 4.17e+07 0.001 0.001

 WS 0.0 16.2 1.27e+08 2.057 2.057

 Subtotal 0.0 2.36e+08 2.1 2.1 

TOTAL 1,321.3 1.11e+10 6,537.5 7,858.8 a

Slash indicates devices used in conjunction.

 APCD:

 DS = Dry scrubber

 EGB = Electro gravel bed

 FF = Fabric filter

 H-ESP = Hot-sided electrostatic precipitator

 Unc = Uncontrolled

 WS = Wet scrubber

 MWC type:

 MB/RC = Mass burn rotary kiln

 MB/REF = Mass burn refractory walled

 MB/WW = Mass burn waterwalled

 MOD/EA = Modular excess air

 MOD/SA = Modular starved air

 RDF/cofired = Refuse-derived fuel/cofired

 RDF/ded = Refuse-derived fuel/dedicated 

3-29
Table 3-8b. Annual TEQDF-WHO98 emissions to the air from municipal 

waste combustors (MWCs) operating in 1987 

MWC type 

Air pollution 

control device 

(APCD)a 

TEQDF ­

WHO98 

emissions 

from tested 

facilities 

(g TEQ/yr) 

Average 

TEQDF ­

WHO98 

emission 

factor 

(ng/kg) 

Activity level 

nontested 

facilities 

(kg/yr) 

TEQDF ­

WHO98 

emissions 

from 

nontested 

facilities 

(g TEQ/yr) 

Total TEQDF ­

WHO98 

emissions 

from all 

facilities 

(g TEQ/yr) 

MB/WW DS/FF 0.039 – 0 0.0 0.039

 H-ESP 485.0 535.0 3.27e+09 1,749.5 2,234.5

 Subtotal 485.0 1,749.5 2,234.5 

MB/REF DS/FF 0.0 0.72 1.41e+08 0.1 0.1

 H-ESP 0.0 535.0 2.00e+09 1,070.0 1,070.0

 WS 0.0 254.0 9.01e+08 228.9 228.9

 Subtotal 0.0 1,299.0 1,299.0 

MB/RC FF 0.0 93.1 1.58e+07 1.47 1.47

 H-ESP 53.4 316.0 2.25e+08 71.1 124.5

 Subtotal 53.4 72.6 126.0 

RDF/ded H-ESP 946.0 1,679.0 2.45e+09 4,113.6 5,059.6

 WS 0.0 253.0 3.38e+08 85.5 85.5

 Subtotal 946.0 4,199.1 5,145.1 

RDF/cofired H-ESP 0.0 253.0 2.53e+08 64.0 64.0 

MOD/SA FF

 H-ESP

 0.0

 0.068

 17.0 

 85.7 

1.43e+08 

3.61e+08 

2.4 

30.9 

2.4

31.0

 Unc 0.0 0.024 5.73e+08 0.01 0.01

 WS 0.0 17.0 5.30e+07 0.9 0.9

 Subtotal 0.068 34.2 34.3 

MOD/EA EGB 0.0 0.024 6.76e+07 0.0016 0.0016

 Unc 0.0 0.024 4.17e+07 0.001 0.001

 WS 0.0 17.0 1.27e+08 2.16 2.16

 Subtotal 0.0 2.2 2.2 

TOTAL 1,484.5 3.04e+09 7,420.6 8,905.1 

a

Slash indicates devices used in conjunction.

 APCD:

 DS = Dry scrubber

 EGB = Electro gravel bed

 FF = Fabric filter

 H-ESP = Hot-sided electrostatic precipitator

 Unc = Uncontrolled

 WS = Wet scrubber

 MWC type:

 MB/RC = Mass burn rotary kiln

 MB/REF = Mass burn refractory walled

 MB/WW = Mass burn waterwalled

 MOD/EA = Modular excess air

 MOD/SA = Modular starved air

 RDF/cofired = Refuse-derived fuel/cofired

 RDF/ded = Refuse-derived fuel/dedicated 

3-30
Reference year g TEQDF-WHO98 g I-TEQ Confidence rating 

2000 83.8 76.3 A 

1995 1,393.5 1,101.3 B 

1987 8,905.1 7,858.8 B 

As noted, a high confidence rating (A) is assigned to the estimate of dioxin emissions 

from MWCs in 2000 because a large number of facilities were tested, providing a highly certain 

and representative emissions estimate. Table 3-5 displays the CDD/CDF TEQ emissions by each 

MWC facility operating in 2000. Moreover, the activity level of MWCs in 2000 is known and is 

very representative of this source. A confidence rating B is assigned to the overall dioxin air 

releases for years 1995 and 1987. Although the activity level for those years is known and is 

considered very representative, fewer facilities were stack tested in those years relative to the 

total number of MWCs in operation. Therefore, there was a reliance on the use of emission 

factors to estimate releases in 1995 and 1987. However, estimates for these years are considered 

representative of MWCs operating in those years. 

3.1.5. Congener Profiles of Municipal Waste Combustion Facilities 

The air emissions from MWCs contain a mixture of CDD and CDF congeners. These 

mixtures can be translated into what are called “congener profiles,” which represent the 

distribution of total CDDs and total CDFs present in the mixture. A congener profile may serve 

as a signature of the types of CDDs/CDFs associated with a particular MWC technology and 

APCD. Figure 3-10 is a congener profile of an MB-WW MWC equipped with an SDSS and an 

FF (the most common type of MWC and APCD design in use today). This congener profile 

indicates that OCDD dominates CDD/CDF emissions and that every toxic CDD/CDF congener 

is detected in the emissions. Figures 3-11 and 3-12 present 2,3,7,8-TCDD frequency distribution 

and 1,2,3,7,8-PeCDD frequency distribution, respectively. According to a memorandum dated 

March 27, 2003, from Jason Huckaby, ERG, to Walt Stevenson, U.S. EPA, the distribution of 

these two congeners varies little from MWC to MWC. Although these two congeners represent 

less than 1% of total dioxin/furan emissions, they contribute approximately 13 to 23% of the ITEQDF emissions, depending on which TEF system is used. 

3.1.6. Estimated CDDs/CDFs in MWC Ash 

Ash from MWCs is required to be disposed of in permitted landfills from which releases 

to the general environment are controlled. For background purposes, however, some information 

is presented below about the quantities of CDDs/CDFs in ash from MWCs. 

3-31
 Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.02 0.04 0.06 0.08 0.1 0.12 

Ratio (congener group emission factor/total CDD/CDF emission factor) 

0 0.08 0.1 0.15 0.2 0.25 

Figure 3-10. Congener and congener group profiles for air 

emissions from a mass burn waterwall municipal waste 

combustor equipped with a dry scrubber and fabric filter. 

3-32
 ND = 0 ND = 1⁄2 detection limit ND = full detection 

Arithmetic average (ng/dscm @ 7%O2) 0.00545 0.00578 0.0061 

Arithmetic standard deviation 0.01542 0.01535 0.0153 

3-33

Figure 3-11. 2,3,7,8-TCDD frequency distribution (negative natural log concentration
ND = 0 ND = 1⁄2 detection limit ND = full detection 

Arithmetic average (ng/dscm @ 7%O2) 0.01589 0.0163 0.01669 

Arithmetic standard deviation 0.03375 0.03364 0.0336 

3-34 

Figure 3-12. 1,2,3,7,8-PeCDD frequency distribution (negative natural log concentratio
An estimated 7 million metric tons of total ash (bottom ash plus fly ash) were generated by 

MWCs in 1992 (telephone conversation between J. Loundsberry, U.S. EPA Office of Solid 

Waste, and L. Brown, Versar, Inc., February 24, 1993). EPA indicated that 2 to 5 million metric 

tons of total ash were produced annually in the late 1980s from MWCs, with fly ash comprising 

5 to 15% of the total (U.S. EPA, 1991a) . 

EPA reported the results of analyses of MWC ash samples for CDDs/CDFs (U.S. EPA, 

1990a). 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/CDFs. The TEQ levels in the ash (fly ash mixed 

with bottom ash) ranged from 106 to 466 ng I-TEQDF/kg, with a mean value of 258 ng ITEQDF/kg. CDD/CDF levels are generally much higher in fly ash than in bottom ash. For 

example, Fiedler and Hutzinger (1992) reported levels of 13,000 ng I-TEQDF/kg in fly ash. 

In another study (Washington State Department of Ecology, 1998), CDD/CDF congener 

data were reported for ash and other solid residuals from three municipal incinerators (Fort 

Lewis, Bellingham [municipal plus medical wastes], and Spokane). The data were compiled and 

evaluated to determine a total I-TEQ concentration and loading. Nondetect values were included 

as either zero or one-half the DL or at the DL. The results were as follows, assuming that 

nondetect values were at zero concentration: 

Location Type of Residual I-TEQ (:g/kg) I-TEQ (mg/day) 

Ft. Lewis Bottom ash 0.0 0.0 

Fly ash 4.98 0.76 

Bellingham Mixed ash 

(average of three tests) 0.038 1.14 

Spokane Mixed ash 0.163 38.0 

Fly ash 0.510 24.3 

Bottom ash 0.0001 0.02 

In Shane et al. (1990), ash from five municipal incinerators was analyzed for a number of 

constituents, including CDDs (but not CDFs) and PCBs. For dioxins, three of the incinerators 

were at nondetectable levels (DL of 1 :g/kg). The other two incinerators had detectable levels of 

five CDD congener groups (no analyses were reported for individual congeners), and the 

averages for the two units were 26, 59, 53, 25, and 12 :g/kg for TCDD, PeCDD, HxCDD, 

HpCDD, and OCDD, respectively. These levels were much higher that those reported by EPA 

(U.S. EPA, 1990a). 

For PCBs, the five sets of ashes were analyzed for 10 congener groups. All groups were 

detected for one of the incinerators. However, the other four incinerators contained little or no 

3-35
octa, nona, or deca congeners. The average PCB concentration (all congener groups) for the five 

incinerators was 216 :g/kg, with a range of 99 to 322 :g/kg. 

No generation rates of the ashes were given (Shane et al., 1990); therefore, the measured 

concentrations cannot be readily converted to quantities of CDDs or PCBs. The ashes from each 

of the five incinerators were disposed of in various ways. For two of the incinerators, the ash 

was sent to metal recovery and also landfilled. For a third, the fly ash was sold. For a fourth, the 

ashes were landfilled only. For the fifth, the ashes were used in road building and also landfilled. 

For those incinerators with more than one ash disposition, no breakdown was given of how much 

went to each location. Fifteen other incinerators were discussed in Shane et al. (1990). Thirteen 

of them disposed of their ash exclusively in landfills, and the other two partially disposed of their 

ash in landfills. 

Table 7 in Clement et al. (1988) presents 13 data sets for CDD/CDF congener groups for 

municipal incinerator ash. The average data for each congener group and the ranges of each 

group are given in Table 3-9. No data were presented for individual congeners or for ash 

quantities. 

Table 3-9. Average and range of CDD/CDF congener groups in fly ash from 

a municipal incinerator (:g/kg) 

Congener group Average concentration Concentration range 

TCDD 

PeCDD 

HxCDD 

HpCDD 

OCDD 

3.7 

6.4

9.1 

2.3 

1.5

1.6–12 

2–25 

1.5–42 

0.5–9.2 

0.2–6 

TOTAL CDDs 23.0 6.2–94 

TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF 

12.0 

17.0 

14.0 

2.9 

1.2 

5.1–36 

8.3–40 

3.9–40 

0.8–9.2 

ND–2.1 

TOTAL CDFs 47.1 22–110 

Source: Clement et al. (1988). 

3-36
Ash from three incinerators (one in North America, one in Europe, and one in Japan) had 

mean CDD concentrations of 363, 588, and 2.6 :g/kg, respectively (Table 3-3 in U.S. EPA, 

1987a). The values ranged from less than 0.5 to 3.537 :g/kg. For CDFs, the respective mean 

concentrations for the first two incinerators were 923 and 288 :g/kg. Data for the third 

incinerator were not reported. The CDF range for the two incinerators was from less than 0.5 to 

1,770 :g/kg. No data were given for individual congeners or for quantities of ashes. 

In Table 1 in Lahl et al. (1991), data are presented for concentrations of total CDDs and 

total CDFs in the ash from an ESP from a municipal incinerator. Total CDDs were 140.46 :g/kg 

in the summer samples and 86 :g/kg in the winter samples. Total CDFs were 54.97 :g/kg in the 

summer samples and 73.85 :g/kg in the winter samples. No data were given for individual 

congeners, nor was there information about the quantity of precipitator ash generated. It was 

assumed that the data were not for TEQs. 

A wire reclamation incinerator was reported to have 0.41 :g/kg of CDDs and 11.6 :g/kg 

of CDFs in fly ash from its stack emissions (Table 3-11 in U.S. EPA, 1987a). For the same 

incinerator, the furnace ash concentrations were reported as 0.58 :g/kg CDDs and 0.73 :g/kg 

CDFs. Again, no data were given for individual congeners or for quantities of the ashes. 

Data from the aforementioned sources are compiled in Table 3-10 of this document for 

comparison purposes. Annual TEQ amounts were estimated by multiplying the mean TEQ total 

ash concentration by the estimated amount of MWC ash generated annually (approximately 7 

million metric tons in 1995 and 5 million metric tons in 1987). Where possible, ash quantities 

were broken down into fly ash or bottom ash. Fly ash was assumed to be 10% of the total ash, 

and bottom ash was assumed to be 90% of the total ash. 

Imagawa and Lee (2001) analyzed samples collected from eight Japanese MSW 

incinerators to determine dioxin levels in the fly ash (Table 3-11). Specific congener data were 

not available, so TEQ calculations could not be performed. 

Kobylecki et al. (2001) analyzed the reduction of dioxins in fly ash by pelletizing the ash 

and reburning the pellets in a laboratory-scale bubbling fluidized-bed furnace. Fly ash for the 

test input material was collected from a fly ash filter vessel during 4 days of MWC operation. 

The concentrations of the dioxin collected and composited congeners are shown in Table 3-12. 

The total TEQ value derived by Kobylecki was 862 ng I-TEQDF/kg of fly ash. 

Sakai et al. (2001) analyzed the levels of dioxins and PCBs in fly ash and bottom ash 

from a newly constructed MWC in Japan (Table 3-13). TEQ values derived from the data give a 

total of 423 ng I-TEQDF/kg for fly ash and 10.5 ng I-TEQDF/kg for bottom ash for dioxins and 

31.6 ng I-TEQDF/kg for fly ash and 0.85 ng I-TEQDF/kg for bottom ash for PCBs. 

3-37
Table 3-10. Comparison of the amount of TEQs generated annually in 

municipal waste combustor ash 

Data source 

Type of 

ash 

Mean total 

CDD/CDF 

concentrati 

on 

(ng/kg) 

Mean 

I-TEQDF 

(ng/kg) 

Annual 

TEQ 

amount 

1995 valuea 

(g ITEQDF/yr) 

Annual TEQ 

amount 

1987 valuea 

(g ITEQDF/yr) 

U.S. EPA (1990a) Mixed 12,383 258 1,806 1,290 

Washington State 

Department of Ecology 

(1998)

 Ft. Lewis

 Bellingham

 Spokane 

Bottom 

Fly 

Mixed 

Mixed 

Fly 

Bottom 

0 

71,280 

1,884 

1,414 

10,320 

100 

0 

4,980 

38 

163 

510 

0.1 

0 

3,486 

266 

1,141 

357 

1 

0 

2,490 

190 

815 

255 

0.05 

Shane et al. (1990) Fly 175,000 – – – 

Clement et al. (1988) Fly 70,000 – – – 

U.S. EPA (1987a)

 North America

 Europe

 Japan

 Wire reclamation 

Fly 

Fly 

Fly 

Fly 

Bottom 

1,286,000 

876,000 

2,600 

12,010 

1,310 

– 

– 

– 

– 

– 

– 

– 

– 

– 

– 

– 

– 

– 

–

 Lahl et al. (1991) Mixed 177,640 – – – a

In calculating the annual TEQ amounts, fly ash and bottom ash were considered to be 10% and 90% of the total ash,

 respectively. 

– = Value could not be calculated 

Each of the five facilities sampled by EPA had companion ash disposal facilities 

equipped with leachate collection systems or some means of collecting leachate samples (U.S. 

EPA, 1990a). Leachate samples were collected and analyzed for each of these systems. 

Detectable levels were found in the leachate at only one facility (3 ng I-TEQDF/L); the only 

detectable congeners were HpCDDs, OCDD, and HpCDFs. 

3-38
Table 3-11. Concentration of CDD/CDF congener groups (ng/kg) in fly ash 

samples from combustion of municipal solid waste in eight Japanese 

incineratorsa 

Congener 

group 

Stoker incinerators Fluidized-bed incinerators 

B C D E F I J L

 TCDD 

 PeCDD 

HxCDD 

HpCDD 

OCDD 

5,000 

20,000 

45,000 

70,000 

125,000 

200,000 

340,000 

440,000 

340,000 

110,000 

80,000 

200,000 

250,000 

230,000 

160,000 

75,000 

105,000 

90,000 

37,000 

15,000 

6,000 

10,000 

12,000 

8,000 

7,000 

10,000 

28,000 

41,000 

40,000 

25,000 

10,000 

37,000 

100,000 

200,000 

187,000 

5,000 

10,000 

30,000 

40,000 

50,000 

TCDF 

 PeCDF 

HxCDF 

HpCDF 

OCDF 

25,000 

50,000 

65,000 

75,000 

40,000 

210,000 

410,000 

400,000 

230,000 

20,000 

330,000 

320,000 

300,000 

200,000 

40,000 

50,000 

45,000 

22,000 

10,000 

1,000 

13,000 

14,000 

21,000 

17,000 

10,000 

18,000 

32,000 

34,000 

33,000 

13,000 

50,000 

125,000 

210,000 

225,000 

150,000 

70,000 

120,000 

200,000 

270,000 

120,000 a

 Incinerators are designated by letters because they remained anonymous. 

Source: Imagawa and Lee (2001) (numbers estimated from Figure 2 of report). 

Table 3-12. Concentration of CDD/CDF congener groups in fly ash samples 

from municipal solid waste 

Congener group Concentration (ng/kg) TEQ (ng/kg) 

TCDD 

PeCDD 

HxCDD 

HpCDD 

OCDD 

8,000 

9,000 

40,000 

10,800 

8,000 

15 

45 

100 

50 

1 

TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF 

8,000 

10,000 

9,500 

8,500 

8,000 

10 

300 

300 

40 

1 

TOTAL 119,800 862 

Source: Kobylecki et al. (2001) (estimated from values in Figure 4 of “Before Incineration”). 

3-39
Table 3-13. CDD/CDF concentrations in municipal solid waste ash from a 

newly constructed municipal waste combustor in Japan 

Congener 

Concentration (ng/kg) I-TEQs (ng/kg) 

Fly ash Bottom ash Fly ash Bottom ash 

2,3,7,8-TCDD 

1,2,3,7,8-PeCD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

19 

78 

92 

210 

130 

1,300 

2,800 

1.6 

3.1 

2.6

5.6 

3.6 

33.0 

110.0

19.0 

39.0 

9.2 

21.0 

13.0 

13.0 

2.8 

1.6 

1.65 

0.26 

0.56 

0.36 

0.33 

0.11 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

150 

290 

320 

310 

310 

21 

400 

1,100 

110 

320 

4.8 

5.3 

5.9 

4.4 

4.9 

0.36

6.7 

23.0 

1.6

9.3

15.0 

14.5 

160.0 

31.0 

31.0 

2.1 

40.0 

11.0 

1.1 

0.32 

0.48 

0.265 

2.95 

0.44 

0.49 

0.036 

0.67 

0.23 

0.016 

0.0093 

TOTAL 7,960 226.0 423.0 10.5 

Source: Sakai et al. (2001). 

3.1.7. Recent EPA Regulatory Activities 

As part of the 1990 Clean Air Act mandates, EPA promulgated CDD/CDF emission 

standards for all existing and new MWC units at facilities with aggregate combustion capacities 

greater than 35 metric tons per day (Federal Register, 1995a). These standards, established under 

Section 129 of the Clean Air Act, required facilities to use “maximum achievable control 

technology” (MACT) at MWC units and emission control retrofit for large MWC units (units 

with capacities greater than 225 metric tons per day) by December 2000. In response to a court 

remand, the regulations were subsequently amended to remove small MWC units (units with 

capacities ranging from 35 to 225 metric tons per day) (Federal Register, 1995a). 

The specific emission standards for large MWCs (expressed as ng/dscm of total 

CDD/CDF, based on standard dry gas corrected to 7% oxygen) are a function of the size, APCD 

configuration, and age of the facility, as listed below. 

3-40
1995 emission standards 

(ng total CDD/CDF/dscm) Facility age, size, and APCD 

60 Existing; >225 metric tons/day; ESPbased APCD 

30 Existing; >225 metric tons/day; nonESP-based APCD 

13 New; >225 metric tons/day 

EPA reestablished emission standards for small MWCs in December 2000. These 

standards contain two dioxin emission limits: one for small MWCs at plants with an aggregate 

capacity greater than 250 tons/day (Class I MWCs) and another for small MWCs at plants with 

an aggregate capacity less than 250 tons/day (Class II MWCs). The limits for the Class I MWCs 

were the same as the 1995 limits for large MWCs. The limit for the smaller Class II MWCs is 

125 ng/dscm. These small MWCs were on schedule to comply with the standards by December 

2005. Small MWC emissions were estimated to be 63 g/yr I-TEQ in 2000 and should be less 

than 2 g/yr in 2005, when all control retrofits are completed (Federal Register, 2003). 

3.2. HAZARDOUS WASTE INCINERATION 

Hazardous waste incineration is the controlled pyrolysis and/or oxidation of potentially 

dangerous liquid, gaseous, and solid waste. It is one of the technologies used to manage 

hazardous waste under the Resource Conservation and Recovery Act (RCRA) and the 

Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA, or 

Superfund). 

Hazardous wastes are burned in a variety of situations and are covered in a number of 

different sections in this report. 

• Much hazardous waste is burned in facilities dedicated to burning this type of waste. 

Most of these dedicated facilities are located on-site at chemical manufacturing 

facilities and burn only the waste associated with their on-site industrial operations. 

Hazardous waste is also burned at dedicated facilities located off-site. These facilities 

accept waste from multiple sources. On- and off-site hazardous waste burning 

facilities are addressed in Sections 3.2.1 through 3.2.4. 

• Hazardous waste is also burned in industrial boilers and furnaces that are permitted to 

burn the waste as supplemental fuel. These facilities have significantly different 

furnace designs and operations than those of dedicated hazardous waste incinerators 

(HWIs). They are discussed in Section 3.2.6. 

• Hazardous waste is also burned in halogen acid furnace s (HAFs), in which halogen

acids (such as HCl) may be produced from halogenatedsecondary materials. These 

facilities are discussed in Section 3.2.7. 

3-41
• A number of cement kilns and lightweight aggregate kilns are also permitted to burn 

hazardous waste as auxiliary fuel. These are discussed separately in Section 5.1. 

• Mobile HWIs are typically used for site cleanup at Superfund sites. These units can 

be transported from one location to another and operate for a limited duration at any 

given location. Because these facilities are transitory, they are not included in this 

inventory at this time. 

The following sections review the types of hazardous waste incineration technologies 

commonly in use in the United States and present the CDD/CDF emission estimates from all 

facilities operating in 1987, 1995, and 2000. 

3.2.1. Furnace Designs for HWIs 

The four principal furnace designs employed for the combustion of hazardous waste in 

the United States are rotary kiln, liquid injection, fixed-hearth, and fluidized-bed (Dempsey and 

Oppelt, 1993). The majority of commercial operations use rotary kiln incinerators. On-site 

(noncommercial) hazardous waste incineration technologies use an equal mix of rotary kiln and 

liquid injection furnaces, along with some fixed-hearth and fluidized-bed operations (U.S. EPA, 

1996a). These HWI technologies are discussed below. 

Rotary kiln. Rotary kiln incinerators consist of a rotary kiln coupled with a hightemperature afterburner. Because rotary kilns are excess-air units designed to combust hazardous 

waste in any physical form (i.e., liquid, semisolid, or solid), they are the most common type of 

HWI used by commercial off-site operators. The rotary kiln is a horizontal cylinder lined with 

refractory material. Rotation of the cylinder on a slight slope provides for gravitational transport 

of the hazardous waste through the kiln (Buonicore, 1992a). The tumbling action of the rotating 

kiln causes mixing and exposure of the waste to the heat of combustion, thereby enhancing 

burnout. 

Solid and semisolid wastes are loaded into the top of the kiln by an auger or rotating 

screw. Fluid and pumpable sludges and wastes are typically introduced into the kiln through a 

water-cooled tube. Liquid hazardous waste is fed directly into the kiln through a burner nozzle. 

Auxiliary fuel (natural gas or oil) is burned in the kiln chamber at startup to reach elevated 

temperatures. The typical heating value of hazardous waste (8,000 British thermal units 

[Btu]/kg) is sufficient to sustain combustion without auxiliary fuel (U.S. EPA, 1996a). The 

combustion gases emanating from the kiln are passed through a high-temperature afterburner 

chamber to more completely destroy organic pollutants entrained in the flue gases. Rotary kilns 

can be designed to operate at temperatures as high as 2,580°C, but they more commonly operate 

at about 1,100°C. 

3-42
Liquid injection. Liquid injection incinerators are designed to burn liquid hazardous 

waste. These wastes must be sufficiently fluid to pass through an atomizer for injection as 

droplets into the combustion chamber. The incinerator consists of a refractory-lined steel 

cylinder mounted in either a horizontal or a vertical alignment. The combustion chamber is 

equipped with one or more waste burners. Because of the rather large surface area of the 

atomized droplets of liquid hazardous waste, the droplets quickly vaporize. The moisture 

evaporates, leaving a highly combustible mix of waste fumes and combustion air (U.S. EPA, 

1996a). Secondary air is added to the combustion chamber to complete the oxidation of the 

fume and air mixture. 

Fixed-hearth. Fixed-hearth incinerators are starved-air or pyrolytic incinerators. Waste 

is ram-fed into the primary chamber and incinerated at about 50 to 80% of stoichiometric 

requirements. The resulting smoke and pyrolytic combustion products are then passed through 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. Other types of 

hearths include roller hearths and rotary hearths. Roller hearths use a conveyor system to move 

waste from the kiln entrance to the exit. In rotary hearths, waste enters and exits through the 

same gate, and the hearth rotates inside a circular tunnel kiln. 

Fluidized-bed. The fluidized-bed incinerator is similar in design to the incinerators used 

in MSW incineration (see Section 3.1). In fluidized-bed HWIs, a layer of sand is placed on the 

bottom of the combustion chamber. The bed is preheated by underfire auxiliary fuel at startup. 

The hot gases channel through the sand at relatively high velocity, and the turbulent mixing of 

combustion gases and combustion air causes the sand to become suspended (Buonicore, 1992a) 

and take on the appearance of a fluid medium; hence the term “fluidized-bed” combustor. The 

incinerator is operated at temperatures below the melting point of the bed material (typical 

temperatures are within a range of 650 to 940°C). A constraint on the types of waste burned is 

that the solid waste particles must be capable of being suspended within a furnace. When the 

liquid or solid waste is combusted in the fluid medium, the exothermic reaction causes heat to be 

released into the upper portion of the combustion chamber. The upper portion typically has 

much larger volume than the lower portion, and temperatures can reach 1,000°C (Buonicore, 

1992a). This high temperature is sufficient to combust volatilized pollutants emanating from the 

combustion bed. 

3-43
3.2.2. APCDs for HWIs 

Most HWIs use APCDs to remove undesirable components from the flue gases that 

evolve during the combustion of the hazardous waste. These unwanted pollutants include 

suspended ash particles (PM), acid gases, metals, and organic pollutants. The APCD controls 

collect these pollutants and reduce the amount discharged from the incinerator stack to the 

atmosphere. The levels and types of these combustion byproducts are highly site specific, 

depending on factors such as waste composition and incinerator system design and operating 

parameters (e.g., temperature and exhaust gas velocity). The APCD typically comprises a series 

of different devices that work together to clean the combustion exhaust flue gas. Unit operations 

usually include exhaust gas cooling followed by PM and acid gas control. 

Exhaust gas cooling may be achieved by using a waste heat boiler or heat exchanger, 

mixing with cool ambient air, or injecting a water spray into the exhaust gas. A variety of types 

of APCDs are used to remove PM and acid gases. Such devices include WSs (such as venturi, 

packed bed, and ionizing systems), ESPs, and FFs (sometimes used in combination with dry acid 

gas scrubbing). In general, the control systems can be grouped into the following three 

categories: wet, dry, and hybrid wet/dry systems. The controls for acid gases (either dry or wet 

systems) cause temperatures to be reduced before the gases reach the control device. This 

impedes the formation of CDDs/CDFs in the post-combustion area of the typical HWI. It is not 

unusual for stack concentrations of CDDs/CDFs at a particular HWI to be in the range of 1 to 

100 ng/dscm (Helble, 1993), which is low when compared with concentrations from other waste 

incineration systems. However, the range of total CDD/CDF flue gas concentrations measured 

in the stack emissions of HWIs during trial burns across the class of HWI facilities spans four 

orders of magnitude, ranging from 0.1 to 1,600 ng/dscm (Helble). 

The three categories of APCD systems are described below: 

• Wet system. A WS is used for both particulate and acid gas control. Typically, a 

venturi scrubber and a packed-bed scrubber are used in a back-to-back arrangement. 

Ionizing WSs, wet ESPs, and innovative venturi-type scrubbers may be used for more 

efficient particulate control. WSs generate a wet effluent liquid wastestream 

(scrubber blowdown). They are relatively inefficient at fine particulate control when 

compared with dry control techniques, and they have equipment corrosion concerns. 

However, WSs provide efficient control of acid gases and have lower operating 

temperatures (compared with dry systems), which may help control the emissions of 

volatile metals and organic pollutants. 

• Dry system. In SDSSs, an FF or ESP is used for particulate control, frequently in 

combination with dry scrubbing for acid gas control. Compared with WSs, SDSSs 

are inefficient in controlling acid gases. 

3-44
• Hybrid system. In hybrid systems, a dry technique (ESP or FF) is used for 

particulate control, followed by a wet technique (WS) for acid gas control. Hybrid 

systems have the advantages of both wet and dry systems (lower operating 

temperature for capture of volatile metals, efficient collection of fine particulates, 

efficient capture of acid gases) while avoiding many of the disadvantages. In some 

hybrid systems, known as “zero discharge systems,” the WS liquid is used in the dry 

scrubbing operation, thus minimizing the amount of liquid byproduct waste. 

Facilities that do not use any APCDs fall under a separate and unique category. These are 

primarily liquid waste injection facilities, which burn wastes with low ash and low chlorine 

content; therefore, they are low emitters of PM and acid gases. 

3.2.3. Estimation of CDD/CDF Emission Factors for HWIs 

To estimate emission factors, EPA’s Office of Research and Development (ORD) 

generally subdivides the combustors in each source category into design classes judged to have 

similar potential for CDD/CDF emissions. However, as explained below, dedicated HWIs have 

not been subdivided. 

Total CDD/CDF emissions are likely the net result of all three of the mechanisms 

described above (pass through, precursor, and de novo synthesis); however, the relative 

importance of each mechanism can vary among source categories. In the case of HWIs, the third 

mechanism (post-combustion formation) is likely to dominate, because HWIs are typically 

operated at high temperatures and with long residence times, and most have sophisticated realtime monitoring and controls to manage the combustion process. Therefore, any CDDs/CDFs 

present in the feed or formed during combustion are likely to be destroyed before exiting the 

combustion chamber. Consequently, for purposes of generating emission factors, it was decided 

not to subdivide this class on the basis of furnace type. 

Emissions resulting from the post-combustion formation of CDDs/CDFs in HWIs can be 

minimized using a variety of technologies: 

• Rapid flue gas quenching. The use of wet and dry scrubbing devices to remove acid 

gases usually results in the rapid reduction of flue gas temperatures at the inlet to the 

APCD. If the temperature is reduced below 200°C, the low-temperature catalytic 

formation of CDDs/CDFs is substantially retarded. 

• Use of PM APCDs. PM control devices can effectively capture condensed and 

adsorbed CDDs/CDFs that are associated with the entrained PM (in particular, those 

adsorbed on unburned carbon-containing particulates). 

3-45
• Use of activated carbon. Activated CI is used at some HWIs to collect (sorb) 

CDDs/CDFs from the flue gas. This may be achieved using carbon beds or by 

injecting carbon and collecting it in a downstream PM APCD. 

All of these approaches appear to be very effective in controlling dioxin emissions at 

dedicated HWIs; emissions data are insufficient to generalize about any minor differences. 

Consequently, for purposes of generating emission factors, ORD decided not to subdivide this 

class on the basis of APCD type. 

EPA’s Office of Solid Waste (OSW) compiled a database summarizing the results of 

stack testing for CDDs/CDFs at a number of HWIs between 1993 and 2000 (U.S. EPA, 2002b). 

The CDD/CDF emission factors for HWIs in 1995 are based on data from 17 HWIs tested 

between 1993 and 1996; emissions of HWIs in 2000 are based on data from 22 HWIs tested in 

2000. The furnaces at the 22 HWI facilities tested in 2000 were 11 rotary kiln incinerators, 6 

liquid injection incinerators, 2 rotary hearth units, 1 fluidized-bed incinerator, and 1 roller hearth. 

Rather than classifying the dedicated HWI designs to derive an emission factor, ORD 

decided to derive the emission factor as an average across all tested facilities. First, an average 

emission factor was calculated using eq 3-3. 

C × Fv __________ EFHWI = (3-3) 

Iw 

where: 

EFHWI = emission factor (average ng TEQ per kg of waste burned) 

C = TEQ or CDD/CDF concentration in flue gases (ng TEQ/dscm) (20°C,

 1 atm; adjusted to 7% O2) 

Fv = volumetric flue gas flow rate (dscm/hr) (20°C, 1 atm; adjusted to 7% O2) 

Iw = average waste incineration rate (kg/hr) 

Although 22 HWIs were tested in 2000, the OSW database contained values for flue gas 

flow rates for only 12 of these incinerators. Therefore, only 12 HWIs could be used to develop 

an emission factor. After developing an average emission factor for each HWI, the overall 

average congener-specific emission factor was derived using eq 3-4. 

EFavgHWI

n=1–17 = (EFHWI1

 + EFHWI2 + EFHWI3 + ........ + EFHWI17 ) / N (3-4)

3-46
where: 

EFavgHWI = average emission factor for the tested HWIs (ng/kg) 

N = number of tested facilities 

Tables 3-14a and 3-14b present the average emission factors developed for specific 

congeners, total CDDs and total CDFs, and TEQs for the HWIs tested from 1993 to 1996 and in 

2000, respectively. The average congener emission profile for the 17 HWIs tested from 1993 to 

1996 are presented in Figure 3-13. The average emission factor for the 17 HWIs was 3.88 ng 

TEQDF-WHO98/kg (3.83 ng I-TEQDF/kg) of waste feed (assuming nondetect values were zero). 

The average emission factor for the 22 HWIs tested in 2000 was 2.13 ng TEQDF-WHO98/kg (2.12 

ng I-TEQDF/kg) of waste feed (assuming nondetect values were zero). The emission factor 

developed for reference year 1995 was used as a surrogate for reference year 1987. 

3.2.4. Emission Estimates for HWIs 

Although emissions data were available for 10% of the HWIs operating in 1995 and 17% 

of the HWIs operating in 2000 in the United States (i.e., 22 of the 132 HWIs operating in 2000 

have been tested), the emission factor estimates are assigned a medium confidence rating because 

of uncertainties resulting from the following: 

• Variability of the waste feeds. The physical and chemical composition of the waste 

can vary from facility to facility and even within a facility. Consequently, CDD/CDF 

emissions measured for one feed may not be representative of those of other feeds. 

• Trial burns. Much of the CDD/CDF emissions data were collected during trial 

burns, which are required as part of the RCRA permitting process and are used to 

establish the destruction rate efficiency of principal hazardous organic constituents in 

the waste. During trial burns, a prototype waste is burned that is intended to 

maximize the difficulty in achieving good combustion. For example, chlorine, 

metals, and organics may be added to the waste. The HWI may also be operated 

outside normal operating conditions. The temperature of both the furnace and the 

APCD may vary by a wide margin (high and low temperatures), and the waste feed 

system may be increased to maximum design load. Accordingly, it is uncertain how 

representative the CDD/CDF emissions measured during the trial burn will be of 

emissions during normal operating conditions. 

Dempsey and Oppelt (1993) estimated that up to 1.3 million metric tons of hazardous 

waste were combusted in HWIs during 1987. A confidence rating of medium is assigned to this 

estimate. EPA estimated that 1.5 million metric tons of hazardous waste were combusted in 

HWIs each year in the early 1990s (Federal Register, 1996a). The activity level estimate for 

3-47
Table 3-14a. CDD/CDF emission factors for hazardous waste incinerators 

and boilers tested from 1993 to 1996 

Congener 

Incinerator average mean 

emission factor (17 facilities) 

(ng/kg feed) 

Hot-sided ESP boilers mean 

emission factor (2 facilities) 

(ng/kg feed) 

Nondetect set 

to 

1⁄2 detection 

limit 

Nondetect set to 

zero 

Nondetect set to 

1⁄2 detection 

limit 

Nondetect set to 

zero 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.44 

0.18 

0.22 

0.32 

0.49 

1.77 

4.13 

0.14 

0.14 

0.18 

0.28 

0.48 

1.74 

3.74 

0.1 

0.11 

0.15 

0.2 

0.22 

1.17 

5.24 

0.0 

0.04 

0.08 

0.18 

0.2 

1.17 

5.24 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF

1,2,3,4,7,8,9-HpCDF 

OCDF

2.96 

2.36 

2.56 

9.71 

3.95 

0.31 

2.7 

16.87 

1.74 

13.79 

2.69 

2.33 

2.51 

9.71 

3.95 

0.29 

2.7 

16.68 

1.71 

13.46 

0.81 

0.38 

0.52 

0.83 

0.37 

0.08 

0.56 

1.04 

0.18 

0.7 

0.81 

0.38 

0.52 

0.83 

0.37 

0.02 

0.56 

0.93 

0.16 

0.7 

Total I-TEQDF 

Total TEQDF-WHO98 

4.22 

4.29 

3.83 

3.88 

0.78 

0.83 

0.64 

0.65 

Total CDD/CDF 153 153 28.83 28.39 

ESP = Electrostatic precipitator 

3-48
Table 3-14b. CDD/CDF emission factors for hazardous waste incinerators 

and boilers tested in 2000 

Congener/congener 

group 

Incinerator average mean 

emission factor (12 facilities) 

(ng/kg feed)

a 

Hot-sided ESP boilers mean 

emission factor (1 facility) 

(ng/kg feed)

a 

Nondetect set to 

1⁄2 detection limit 

Nondetect set 

to zero 

Nondetect set to 

1⁄2 detection limit 

Nondetect set to 

zero 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.0615 

0.6141 

0.2347 

0.5408 

0.3037 

2.729 

5.211 

0.036 

0.0907 

0.1395 

0.4351 

0.2178 

2.699 

5.17 

0.0346 

0.0488 

0.1149 

0.1715 

0.3361 

1.406 

1.554 

0.0 

0.0 

0.0789 

0.1228 

0.231 

1.4055 

1.5541 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.6931 

0.9406 

0.88 

4.085 

3.031 

2.667 

1.218 

28.74 

5.056 

36.270 

0.6399 

0.8375 

0.735 

4.045 

3.001 

2.637 

1.121 

28.71 

5.021 

36.23 

0.9531 

0.4599 

0.8836 

3.611 

0.69 

0.038 

1.3272 

4.6345 

0.1895 

0.7841 

0.9531 

0.3862 

0.8836 

3.6108 

0.561 

0.0 

1.3272 

4.6345 

0.1257 

0.7841 

Total I-TEQDF 

Total TEQDF-WHO98 

2.54 

2.809 

2.119 

2.127 

1.313 

1.335 

1.214 

Total CDD/CDF 195.70 194.10 17.24 16.66 

a

Values incorporating use of the detection limit when the laboratory report indicated “not detected” for individual 

CDD/CDF congeners. 

ESP = Electrostatic precipitator 

NR = Not reported 

Source: U.S. EPA (2002a). 

3-49
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 

Figure 3-13. Congener profile for air emissions from 17 hazardous waste 

incinerators tested from 1993 through 1996. 

1995 is assigned a high confidence rating because it is based on a review by EPA of the various 

studies and surveys conducted in the 1990s to assess the quantity and types of hazardous wastes 

being managed by various treatment, storage, and disposal facilities. Because of a lack of data 

regarding the amount of waste burned in 2000, the 1995 estimate (1.5 million metric tons) was 

also used for determining TEQ emissions for 2000. 

The annual TEQ emissions for reference years 1987, 1995, and 2000 were estimated 

using eq 3-5. 

EHWI = EFHWI × AHWI (3-5) 

where: 

EHWI

 = annual emissions from all HWIs, tested and nontested (g TEQ/yr) 

EFHWI

 = mean emission factor for HWIs (ng TEQ/kg of waste burned) 

AHWI = annual activity level of all operating HWIs (million metric tons/yr) 

3-50
Applying the average TEQ emission factor for dedicated HWIs (3.88 ng TEQDFWHO98/kg waste [3.83 ng I-TEQDF/kg waste]) to these production estimates yields estimated 

emissions of 5 g TEQ (TEQDF-WHO98 or I-TEQDF) in 1987 and 5.8 g TEQDF-WHO98 (5.7 g ITEQDF) in 1995. For 2000, applying the average TEQ emission factors for dedicated HWIs (2.13 

ng TEQDF-WHO98/kg waste [2.12 ng I-TEQDF/kg waste]) to a production estimate of 1.5 million 

metric tons yields estimated emissions of 3.2 g TEQDF-WHO98 (3.18 g I-TEQDF). Medium 

confidence rating is assigned to these estimates because the emission factor was given a medium 

confidence rating. 

3.2.5. Recent EPA Regulatory Activities 

CDD/CDF emissions from HWIs are regulated by EPA (Federal Register, 1999a, 2004). 

The regulations are specific to the I-TEQ concentration in the combustion gases leaving the 

stack. Existing HWIs equipped with waste heat boilers and dry scrubbers (as APCDs) cannot 

emit more than 0.28 ng I-TEQ/dscm. All other existing HWIs are limited to 0.4 ng I-TEQ/dscm 

of stack gas. Regulatory requirements are more strict for newly built HWIs: those equipped with 

waste heat boilers and dry scrubbers (as APCDs) cannot emit more than 0.11 ng I-TEQ/dscm, 

and all others are limited to 0.2 ng I-TEQ/dscm of stack gas. 

3.2.6. Industrial Boilers and Furnaces Burning Hazardous Waste 

In 1991, EPA established rules that allow the combustion of some liquid hazardous waste 

in industrial boilers and furnaces (Federal Register, 1991b). These facilities typically burn oil or 

coal for the primary purpose of generating electricity. Liquid hazardous waste can be burned 

only as supplemental (auxiliary) fuel, and the rule limits use to no more than 5% of the primary 

fuels. These facilities typically use an atomizer to inject the waste as droplets into the 

combustion chamber. They are equipped with particulate and acid gas emission controls and in 

general are sophisticated, well-controlled facilities that achieve good combustion. 

The national OSW database contains congener-specific emission concentrations for two 

boilers burning liquid hazardous waste as supplemental fuel tested from 1993 to 1996. The 

average congener and congener group emission profiles for the industrial boiler data set are 

presented in Figure 3-14. The database also contains congener-specific emission concentrations 

for four boilers tested in 2000. Of the boilers tested in 2000, sufficient data to calculate average 

TEQ emissions were available for only one boiler. The average congener and TEQ emission 

factors are presented in Tables 3-14a and 3-14b. The limited set of emissions data prevented 

subdividing this class to derive an emission factor. The equation used to derive the emission 

factor is the same as eq 3-4. The TEQ emission factors for the industrial boiler are 0.65 ng 

3-51
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 

Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 0.25 

Figure 3-14. Congener and congener group profiles for air emissions from 

boilers and industrial furnaces burning hazardous waste. 

3-52
TEQDF-WHO98/kg (0.64 ng I-TEQDF/kg) of waste feed for 1993 to 1996 and 1.212 ng TEQDFWHO98/kg (1.214 ng I-TEQDF/kg) of waste feed for 2000. These emission factors are assigned a 

low confidence rating because they reflect testing at only 2 of 136 hazardous waste boilers and 

furnaces operating from 1993 to 1996 and only 1 of the 114 hazardous waste boilers and furnaces 

operating in 2000. 

Dempsey and Oppelt (1993) estimated that approximately 1.2 billion kg of hazardous 

waste were combusted in industrial boilers/furnaces in 1987. EPA estimated that in each year in 

the early 1990s approximately 0.6 billion kg of hazardous waste were combusted in industrial 

boilers/furnaces (Federal Register, 1996a). It is possible that cement kilns and light-weight 

aggregate kilns burning hazardous waste were included in the estimate by Dempsey and Oppelt 

for 1987; the estimate for 1995 does not appear to include these hazardous waste-burning kilns. 

A confidence rating of low is assigned to the estimated activity level for 1987, which was largely 

based on a review of state permits (Dempsey and Oppelt, 1993). The activity level estimate for 

1995 is assigned a medium confidence rating because it was based on a review by EPA of the 

various studies and surveys conducted in the 1990s to assess the quantity and types of hazardous 

wastes being managed by various treatment, storage, and disposal facilities. Because of a lack of 

data regarding the amount of waste burned in 2000, the 1995 estimate (1.5 million metric tons) 

was used as a surrogate for 2000. 

Equation 3-5, which was used to calculate annual TEQ emissions for dedicated HWIs, 

was also used to calculate annual TEQ emissions for industrial boilers/furnaces. Multiplying the 

average TEQ emission factors by the total estimated kg of liquid hazardous waste burned in 

1987, 1995, and 2000 yields annual emissions in g-TEQ/yr. From this procedure, the emissions 

from all industrial boilers/furnaces burning hazardous waste as supplemental fuel are estimated 

as 0.78 g TEQDF-WHO98 (0.77 g I-TEQDF) in 1987, 0.39 g TEQDF-WHO98 (0.38 g I-TEQDF) in 

1995, and 1.82 g TEQ (TEQDF-WHO98 or I-TEQDF) in 2000. Because of the low confidence 

rating for the emission factor, the overall confidence rating is low for the emission estimates for 

all three reference years. 

3.2.7. Halogen Acid Furnaces Burning Hazardous Waste 

Hazardous waste can be used in the production of halogen acids using an HAF. 

According to EPA rules, products that qualify as hazardous waste, as defined in 40 CFR 261.2 

must be regulated as such, even if the products are used in the production of halogen acids using 

an HAF (Federal Register, 1991c). 

The national OSW database contains congener-specific emission concentrations for two 

HAFs burning liquid hazardous waste as supplemental fuel tested in 2000. Data from these two 

facilities were used to calculate an emission factor for HAFs. The average congener and TEQ 

3-53
emission factors are presented in Table 3-15. The equation used to derive the emission factor is 

the same as eq 3-4. The average TEQ emission factor for HAFs is 0.836 ng TEQDF-WHO98/kg 

(0.803 ng I-TEQDF/kg) of waste feed for reference year 2000. This emission factor is assigned a 

low confidence rating because it reflects testing at only 12.5% of all HAFs operating in 2000 (2 

out of 16). 

Table 3-15. CDD/CDF emission factors for halogen acid furnaces tested in 

2000 

Congener/congener 

group 

Incinerator average mean emission factor (12 facilities) (ng/kg feed) 

Nondetect set to 

1⁄2 detection limit Nondetect set to zero 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.0274 

0.1164 

0.0979 

0.1663 

0.1686 

0.9868 

1.4944 

0.0208 

0.112 

0.0913 

0.1594 

0.1293 

0.9868 

1.4944 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.3821 

0.583 

0.5689 

1.1244 

0.7172 

0.4412 

0.2685 

3.4914 

1.0429 

25.015 

0.3821 

0.583 

0.5689 

1.1244 

0.7172 

0.4412 

0.2685 

3.4914 

1.0429 

25.015 

Total I-TEQDF 

Total TEQDF-WHO98 

0.8176 

0.8519 

0.8034 

0.8356 

Total CDD/CDF 62.4773 62.4607 

Source: U.S. EPA (2002a). 

The amount of hazardous waste combusted using HAFs in 2000 was conservatively 

estimated to be 375,600 metric tons. This estimate is based on data provided by OSW that 

described activity levels for each individual HAF in 2000. Activity data were available for 14 of 

the 16 facilities. By assuming that plants operate continuously throughout the year, that they are 

always running at 80% of maximum capacity, and that the activity levels represent the maximum 

capacity, a conservative estimate for the annual quantity burned per HAF was derived (23,480 

3-54
kg/yr). This quantity multiplied by the total universe of 16 facilities yields the final estimate of 

375,600 metric tons. This was assigned a low confidence rating because the data was possibly 

nonrepresentative. 

Equation 3-5, which was used to calculate annual TEQ emissions from dedicated HWIs, 

was also used to calculate annual TEQ emissions from HAFs. Multiplying the average TEQ 

emission factors by the total estimated kilograms of liquid hazardous waste burned in 2000 yields 

annual emissions in g I-TEQDF. From this procedure, the emissions from all industrial 

boilers/furnaces burning hazardous waste as supplemental fuel are estimated as 0.31 g TEQDFWHO98 (0.3 g I-TEQDF). Because of the low confidence rating for the emission factor, the 

overall confidence rating is low for the emission estimates. 

3.2.8. Solid Waste from Hazardous Waste Combustion 

U.S. EPA (1987a) contains limited data on ash generated from hazardous waste 

incineration. The study indicates that the mean concentrations of CDDs and CDFs from an HWI 

with an afterburner were 538 :g/kg and 2,853 :g/kg, respectively (Table 3-8 in U.S. EPA, 

1987a). Specific data for congeners and for ash quantities were not provided. 

3.3. MEDICAL WASTE INCINERATION 

Medical waste incineration is the controlled burning of solid wastes generated primarily 

by hospitals, veterinary facilities, and medical research facilities. EPA defines medical waste as 

any solid waste generated in the treatment, diagnosis, or immunization of humans or animals or 

research pertaining thereto or in the production or testing of biologicals (Federal Register, 

1997a). The primary purposes of medical waste incineration are to reduce the volume and mass 

of waste in need of land disposal and to sterilize the infectious materials. The following sections 

review the basic types of medical waste incinerator (MWI) designs used to incinerate medical 

waste and the distribution of APCDs used on MWIs and summarize the derivation of dioxin TEQ 

emission factors for MWIs and the national dioxin TEQ emission estimates for reference years 

1987, 1995, and 2000. 

3.3.1. Design Types of MWIs Operating in the United States 

For purposes of this document, EPA has classified MWIs into three broad technology 

categories: modular furnaces using controlled air, modular furnaces using excess air, and rotary 

kilns. Of the MWIs in use today, the vast majority are believed to be modular furnaces using 

controlled air. EPA has estimated that 97% are modular furnaces using controlled air, 2% are 

modular furnaces using excess air, and 1% are rotary kiln combustors (U.S. EPA, 1997a). 

3-55
Modular furnaces using controlled air. Modular furnaces have two separate 

combustion chambers mounted in series (one on top of the other). The lower chamber is where 

the primary combustion of the medical waste occurs. Medical waste is ram-fed into the primary 

chamber and underfire air is delivered beneath the incinerator hearth to sustain good burning of 

the waste. The primary combustion chamber is operated at below stoichiometric levels, hence 

the terms “controlled air” or “starved air.” With substoichiometric conditions, combustion 

occurs at relatively low temperatures (760 to 985°C). Under the conditions of low oxygen and 

low temperatures, partial pyrolysis of the waste occurs and volatile compounds are released. 

The combustion gases pass into a second chamber. Auxiliary fuel (such as natural gas) is 

burned to sustain elevated temperatures (985 to 1,095°C) in this secondary chamber. The net 

effect of exposing the combustion gases to an elevated temperature is more complete destruction 

of the organic contaminants entrained in the combustion gases emanating from the primary 

combustion chamber. Combustion air at 100 to 300% in excess of stoichiometric requirements is 

usually added to the secondary chamber. Gases exiting the secondary chamber are directed to an 

incinerator stack (U.S. EPA, 1991b, 1997a; Buonicore, 1992b). Because of its low cost and good 

combustion performance, this design has been the most popular choice for MWIs and has 

accounted for more than 95% of systems installed over the past two decades (U.S. EPA, 1990b, 

1991b; Buonicore, 1992b). 

Modular furnaces using excess air. These systems use the same modular furnace 

configuration as described above for the controlled-air systems. The difference is that the 

primary combustion chamber is operated at air levels of 100 to 300% in excess of stoichiometric 

requirements, hence the name “excess air.” A secondary chamber is located on top of the 

primary unit. Auxiliary fuel is added to sustain high temperatures in an excess-air environment. 

Excess-air MWIs typically have smaller capacity than do controlled-air units, and they are 

usually batch-fed operations. This means that the medical waste is ram-fed into the unit and 

allowed to burn completely before another batch of medical waste is added to the primary 

combustion chamber. Figure 3-4 shows a schematic of a typical modular furnace using excess 

air. 

Rotary kiln. In terms of design and operational features, the rotary kiln technology used 

in medical waste incineration is similar to that employed in both municipal and hazardous waste 

incineration (see description in Section 3.1). Because of their relatively high capital and 

operating costs, few rotary kiln incinerators are in operation for medical waste treatment (U.S. 

EPA, 1990b, 1991b; Buonicore, 1992b). 

MWIs can be operated in three modes: batch, intermittent, and continuous. Batch 

incinerators burn a single load of waste, typically only once per day. Waste is loaded and ashes 

are removed manually. Intermittent incinerators, which are loaded continuously and frequently 

3-56
with small waste batches, operate less than 24 hr/day, usually on a shift basis. Either manual or 

automated charging systems can be used, but the incinerator must be shut down for ash removal. 

Continuous incinerators are operated 24 hr/day and use automatic charging systems to charge 

waste into the unit in small, frequent batches. All continuous incinerators operate using a 

mechanism to automatically remove the ash from the incinerator (U.S. EPA, 1990b, 1991b). 

3.3.2. Characterization of MWIs for Reference Years 1987, 1995, and 2000 

Medical waste incineration remains a poorly characterized industry in the United States 

in terms of knowing the exact number of facilities in operation over time, the types of APCDs 

installed on these units, and the aggregate volume and weight of medical waste that is combusted 

in any given year (U.S. EPA, 1997a). The primary reason for this lack of information is that 

permits were not generally required for the control of pollutant stack emissions from MWIs until 

the early 1990s, when state regulatory agencies began setting limits on emissions of PM and 

other contaminants (Federal Register, 1997a). Prior to that, only opacity was controlled. 

The information available to characterize MWIs from 1987 and 1995 comes from 

national telephone surveys, stack emission permits, and data gathered by EPA during public 

hearings (Federal Register, 1997a). For 2000, information was also provided by a memorandum 

on emissions from MWIs (Strong and Hanks, 1999) and a limited telephone survey (McAloon, 

2003). Strong and Hanks provided information on MWIs in the United States, including the 

APCD being used by each facility. A telephone survey was conducted with the state agencies in 

each of these six states to obtain the number of MWIs that were operating in 2000. EPA was 

able to obtain an updated list from four of the six states, which are shown below, along with the 

dates they were contacted, the number of MWIs operating in 1999, the updated number of MWIs 

for that state in 2000, and the percent of facilities closed over this time period for each state. 

State Date contacted 

No. of MWIs 

1999 2000 

Percentage of facilities 

closed from 1999 to 2000 

Illinois Jan. 16, 2003 97 13 86.6 

Louisiana Jan. 16, 2003 92 24 73.91 

Maryland 

Michigan 

Dec. 2, 2002

Nov. 26, 2002 

36 

228 

30 

45 

16.67 

80.26 

The geometric mean of the closure percentages for the four states was determined to be 

54.09 and the arithmetic mean was 64.36. Maryland had the lowest closure percent from 1999 to 

2000; however, through discussions with representatives of Maryland state agencies, it was 

3-57
determined that close to 70% of the facilities operating in 1999 would be shut down as of 2003. 

It was therefore assumed that the average closure percent of 64.36 was a fairly good estimate for 

all states. This average was applied to the total number of facilities operating in 1999 from the 

Strong and Hanks (1999) memorandum to estimate the number of facilities operating in 2000. 

The information obtained from these sources suggests the following: 

• The number of MWIs in operation for each reference year was approximately 5,000 in 

1987 (U.S. EPA, 1987c), 2,375 in 1995 (Federal Register, 1997a), and 1,065 in 2000 

(Strong and Hanks, 1999; McAloon, 2003). 

• The amount of medical waste combusted annually in the United States was 

approximately 1.43 billion kg in 1987 (U.S. EPA, 1987c) and 0.77 billion kg in 1995 

(Federal Register, 1997a). 

These estimates indicate that between 1987 and 1995 the total number of operating MWIs 

and the total amount of waste combusted decreased by more than 50%. From 1995 to 2000, the 

total number of operating MWIs decreased by approximately 55%. A variety of factors probably 

contributed to the reduction in the number of operating facilities, including federal and state 

regulations and air pollution control requirements. In 1997, EPA adopted emission guidelines 

for existing MWIs (incinerators constructed on or before June 20, 1996) and New Source 

Performance Standards for new MWIs (incinerators constructed after June 20, 1996). The Clean 

Air Act requires that states implement the emission guidelines according to a state plan and that 

they submit the state plan to EPA within one year of EPA’s promulgation of the guidelines (i.e., 

by September 15, 1998). The compliance schedule, however, allows up to three years from EPA 

approval of the state plan for MWIs to comply, provided the plan includes enforceable 

increments of progress. All MWIs were required to be in compliance within three years of 

approval of their state plan or by September 15, 2002, whichever was earlier. 

Compliance is stated to be either completion of retrofit of air pollution controls or 

shutdown of the facility. As a result, many facilities have closed down and hospitals have 

switched to less expensive medical waste treatment technologies, such as autoclaving (Federal 

Register, 1997a). Autoclaving, or steam sterilization, is one of the most common waste 

management practices used today. This process involves placing bags of infectious waste into a 

sealed chamber, sometimes pressurized, and then heating it by direct contact with steam to 

sterilize the waste. 

The actual controls used on MWIs on a facility-by-facility basis in 1987 are unknown, 

and EPA generally assumes that MWIs were mostly uncontrolled (U.S. EPA, 1987c). However, 

the modular design does cause some destruction of organic pollutants within the secondary 

3-58
combustion chamber. Residence time within the secondary chamber is key to inducing the 

thermal destruction of the organic compounds. Residence time is the time that the organic 

compounds entrained within the flue gases are exposed to elevated temperatures in the secondary 

chamber. EPA has demonstrated with full-scale MWIs that increasing residence time from 1/4 

sec to 2 sec in the secondary chamber can reduce organic pollutant emissions, including 

CDDs/CDFs, by up to 90% (Federal Register, 1997a). In this regard, residence time can be 

viewed as a method of air pollution control. 

EPA estimates that about two-thirds of the medical waste burned in MWIs in 1995 went 

to facilities that had some method of air pollution control (Federal Register, 1997a).The types of 

APCDs installed and the methods used on MWIs include DSI, FFs, ESPs, WSs, and FFs 

combined with packed-bed scrubbers (composed of granular activated carbon). Some organic 

constituents in the flue gases can be adsorbed by the packed bed. Within the uncontrolled class 

of MWIs, about 12% of the waste was combusted in facilities with design capacities of less than 

200 lb/hr, with the majority of waste burned at facilities with capacities greater than 200 lb/hr. In 

controlled facilities, an estimated 70% of the aggregate activity level is associated with facilities 

equipped with either WSs, FFs, or ESPs; 29.9% is associated with facilities that use DSI 

combined with FFs; and less than 1% is associated with facilities that have an FF/packed-bed 

APCD (AHA, 1995; Federal Register, 1997a). 

Strong and Hanks (1999) provided information on the types of APCDs used by facilities 

operating in 1999. Ten types were included in the memorandum, which included residence time 

as a type of control technology. The 10 types were 1/4-sec combustion, 1-sec combustion, 2-sec 

combustion, low-efficiency WS, moderate-efficiency WS, high-efficiency WS, dry lime injectFF, dry lime inject-FF with CI, WS/dry lime inject-FF, and SD/FF with CI. Table 3-16 provides 

an estimated breakdown of these APCDs. 

3.3.3. Estimation of CDD/CDF Emissions from MWIs 

Emission tests reported for 22 MWIs (about 3% of the existing facilities operating in 

2000) were collected for use in this document; emission levels of dioxin-like compounds at most 

facilities are unmeasured. Because so few facilities have been evaluated, the estimation of 

annual air emissions of CDDs/CDFs from MWIs is quite dependent on extrapolations, 

engineering judgment, and assumptions. In addition, the information about the activity levels of 

these facilities is also quite limited. 

The analysis divided MWIs into three design types on the basis of mode of daily 

operation: batch, intermittent, or continuous. This was done using the information from the 

inventory on design-rated annual incineration capacity of each facility. The smaller capacity 

3-59
Table 3-16. Estimated breakdown of facilities by air pollution control device 

(APCD) 

APCD 

Number of 

facilities 

Percent of 

total 

1/4-sec combustion 229 17.5 

1-sec combustion 259 19.8 

2-sec combustion 455 34.8 

Low-efficiency wet scrubber 208 15.9 

Moderate-efficiency wet scrubber 75 5.7 

High-efficiency wet scrubber 16 1.2 

Dry lime inject fabric filter 44 3.4 

Dry lime inject fabric filter with carbon injection 7 0.5 

Wet scrubber/dry lime inject fabric filter 14 1.1 

Spray dryer fabric filter with carbon injection 1 0.1 

Source: Strong and Hanks (1999). 

units were assumed to be batch operations, and the others were classified as either intermittent or 

continuous, assuming a ratio of 3 to 1. 

The activity level of each facility was estimated by multiplying the design-rated annual 

incineration capacity of the MWI (kg/hr) by the hours of operation (hr/yr). The annual hours of 

operation were determined by assuming a capacity factor (defined as the fraction of time that a 

unit operates over the year) for each design type of MWI (Randall, 1995). Table 3-17 is a 

summary of the estimated annual operating hours for each MWI design type. 

In estimating dioxin emissions, the MWIs were divided into two classes: those having 

APCDs (controlled) and those lacking any APCD (uncontrolled). These two classes of MWIs 

are discussed below. 

For 1987, it is assumed that every MWI was uncontrolled. An EPA study of MWIs 

conducted at that time indicated that MWIs operating in 1987 did not need controls because they 

were not subject to state or federal limits on either PM or organic pollutant emissions (U.S. EPA, 

1987c). The activity level estimates were derived from data presented in that 1987 study. This 

approach resulted in the following activity level assumptions for 1987: (a) 15% of the activity 

level (0.22 billion kg) was incinerated annually by MWIs with capacities less than or equal to 

200 lb/hr, and (b) 85% of the activity level (1.21 billion kg) was incinerated annually by facilities 

with capacities greater than 200 lb/hr (see Table 3-18). For 1995, the activity levels were then 

summed across facilities for each APCD subclass (see Table 3-19). 

3-60
Table 3-17. Summary of annual operating hours for each medical waste 

incinerator (MWI) type 

MWI type 

Capacity 

range 

(lb/hr) 

Annual 

charging 

hours 

(hr/yr) 

Maximum 

annual 

charging 

hours 

(hr/yr) 

Capacity 

factor 

Continuous commercial >1,000 7,776 8,760 0.89 

Continuous onsite 501–1,000 

>1,000 

1,826 

2,174 5,475 

0.33 

0.40 

Intermittent #500 1,250 4,380 0.29 

Batch Case by case Case by case Case by case 

Source: U.S. EPA (1990c). 

In 1997, the amount of waste combusted by MWIs was estimated to be 0.8 million tons/yr 

(0.7 billion kg/yr) (NRC, 2000). This number represents a 9% decrease from 1995. If we 

assume that this decrease occurred every two years from 1997 to 2000, the estimated amount of 

waste combusted by MWIs for 2000 would be 0.6 billion kg/yr. This is a conservative estimate, 

considering the large number of facilities that have shut down or switched to less expensive 

medical waste treatment technologies. For 2000 activity level estimates, the same distributions 

among APCD classes were assumed as for 1995. These activity level estimates are presented in 

Table 3-20. For all years, these activity levels were assigned a rating of low confidence because 

the data were judged to be possibly nonrepresentative. 

The stack test results showing the air emissions of dioxin from 24 MWIs were obtained 

and used to calculate 1987 and 1995 emission estimates. After reviewing these test reports, EPA 

determined that 20 met the criteria for acceptability (see Section 3.1.3). In some cases, 

CDD/CDF congener-specific data were not reported or values were missing. In other cases, the 

protocols used in the laboratory analysis were not described; therefore, no determination of the 

adequacy of the laboratory methods could be made. For 2000, two additional test reports from 

facilities operating in that year were obtained and were included with the previously obtained test 

reports in order to calculate updated emission estimates. Each test report was included in its 

respective MWI subclass according to its APCD and was also included in the overall emission 

estimate. 

3-61
Table 3-18. TEQ emissions from medical waste incinerators (MWIs) for reference year 1987 

3-62

MWI classa 

No. of tested 

facilities 

Activity 

level 

(kg/yr) 

Total CDD/ 

CDF 

emission 

factorb 

(g/kg) 

I-TEQDF 

emission 

factor 

(g/kg) 

TEQDF ­

WHO98 

emission 

factor 

(g/kg) 

Annual 

CDD/ 

CDF 

emissions 

(g/yr) 

Ann

I-TE

emiss

(g/y

#200 lb/hr 3 2.19e+08 9.25e!05 1.86e!06 1.98e!06 2.02e+04 4.08e>200 lb/hr 5 1.21e+09 6.05e!05 1.68e!06 1.78e!06 7.32e+04 2.03eTOTAL 8 1.43e+09 9.34e+04 2.44e

Table 3-19. TEQ emissions from medical waste incinerators (MWIs) for reference year 1995 

MWI class 

(air pollution 

control device 

[APCD]) 

MWI 

subclass 

(capacity or 

APCDa) 

No. of 

tested

facilities 

Total 

CDD/CDF 

emission 

 factor 

(ng/kg) 

I-TEQDF 

emission 

factor 

(ng/kg)

TEQDF ­

WHO98 

emission 

 factor (ng/kg) 

Activity 

level 

(kg/yr) 

Annual 

CDD/CDF 

emissions 

(g/yr) 

A

I

em

Uncontrolled #200 lb/hr 3 9.25e+04 1.86e+03 1.98e+03 3.06e+07 2.83e+03 5

>200 lb/hr 5 6.05e+04 1.80e+03 1.78e+03 2.23e+08 1.35e+04 3

Controlled WS/FF/ESP 9 4.67e+04 7.22e+01 6.63e+01 3.71e+08 1.73e+03 2

DSI/FF 2 2.85e+02 6.78 4.61 1.46e+08 4.16e+01 9

FF/PBS 1 1.11e+05 1.35e+03 1.49e+03 6.99e+05 7.76e+01 9

TOTAL 7.71e+08 1.82e+04 4

aSlash(es) indicates devices used in conjunction. 

APCD (air pollution control device): 

DSI = Dry sorbent injection ESP = Electrostatic precipitator FF = Fabric filter PBS = Packed-bed scrubber 

WS = Wet scrubber 
Table 3-20. TEQ emissions from medical waste incinerators (MWIs) for reference year 2000 

3-63

MWI class 

(air pollution 

control device 

[APCD]) 

MWI 

subclass 

(capacity or 

APCDa) 

No. of 

tested 

facilities 

Total 

CDD/CDF 

emission 

factor 

(ng/kg) 

I-TEQDF 

emission 

factor 

(ng/kg) 

TEQDF ­

WHO98 

emission 

factor 

(ng/kg) 

Activity 

level 

(kg/yr) 

Annual 

CDD/CDF 

emissions 

(g/yr) 

An

I-T

em

(g

Uncontrolled #200 lb/hr 3 9.25e+04 1.86e+03 1.98e+03 2.40e+07 2.22e+03 4.4

>200 lb/hr 5 6.05e+04 1.68e+03 1.78e+03 1.74e+08 1.05e+04 3.1

Controlled WS/FF/ESP 9 4.67e+04 6.44e+01 6.63e+01 2.88e+08 1.34e+04 2.0

DSI/FF 2 2.85e+02 4.56 4.61 1.14e+08 3.25e+01 7.7FF/PBS 1 1.11e+05 1.35e+03 1.49e+03 5.40e+05 5.99e+03 7.2

TOTAL 6.01e+08 3.22e+04 3.5

aSlash(es) indicates devices used in conjunction.

 APCD:

 DSI = Dry sorbent injection ESP = Electrostatic precipitator FF = Fabric filter PBS = Packed-bed scrubber WS = Wet scrubber 
______ 

The EPA stack testing method (EPA Method 23) produces a measurement of CDDs/ 

CDFs in units of mass concentration (ng/dscm) at standard temperature and pressure and 1 atm 

and adjusted to a measurement of 7% oxygen in the flue gas (U.S. EPA, 1995a). This 

concentration is assumed to represent conditions at the point of release from the stack into the air 

and to be representative of routine emissions. The emission factors were derived by averaging 

the emission factors across each tested facility in a design class. The emission factor for each 

tested MWI was calculated using the following equation: 

C × F v 

EFMWI = (3-8) 

Iw 

where: 

EFMWI = emission factor per MWI (average ng TEQ per kg medical waste burned) 

C = average TEQ concentration in flue gases of tested MWIs (ng TEQ/dscm)

 (20°C, 1 atm; adjusted to 7% O2) 

Fv = average volumetric flue gas flow rate (dscm/hr) (20°C, 1 atm; adjusted

 to 7% O2) 

Iw = average medical waste incineration rate of the tested MWI (kg/hr) 

3.3.4. Summary of CDD/CDF Emissions from MWIs 

Annual dioxin emissions were estimated by multiplying the emission factor and activity 

level developed for each design class and then summing the calculated emissions for all classes. 

Tables 3-18, 3-19, and 3-20 summarize the resulting national TEQ air emissions for reference 

years 1987, 1995, and 2000, respectively. These tables also indicate the activity level and the 

TEQ emission factor used in estimating annual TEQ emissions. 

In estimating annual TEQ emissions for each reference year, a low confidence rating was 

assigned to the estimate of the activity level, primarily because very limited information is 

available on a facility-level basis for characterizing MWIs in terms of the frequency and duration 

of operation, the actual waste volume handled, and the level of pollution control. The 1987 

inventory of facilities was based on very limited information. Although the 1995 OAQPS 

inventory was more comprehensive than the 1987 inventory, it was still based on a fairly limited 

survey of operating facilities (approximately 6%). The 2000 inventory included only two 

additional facilities and estimated an activity level based on a 1997 value and the distribution 

among APCDs from the 1995 estimates. 

The emission factor estimates were given a low confidence rating because the reports of 

only 20 tested MWI facilities could be used to derive emission factors representing the 2,375 

facilities operating in 1995 (i.e., less than 1% of the estimated number of operating facilities) and 

only two additional test reports were obtained for 2000. Even fewer tested facilities could be 

3-64
used to represent the larger number of facilities operating in 1987 (8 tested facilities were used to 

represent 5,000 facilities). The limited emission tests available cover all design categories used 

here to develop emission factors. However, because of the large number of facilities in each of 

these classes, it is very uncertain whether the few tested facilities in each class capture the true 

variability in emissions. 

Table 3-20 shows the 2000 emissions estimate as being 378 g TEQDF-WHO98 (357 g ITEQDF). The TEQ emissions are estimated to have been 487 g TEQDF-WHO98 (459 g I-TEQDF) 

in 1995 (Table 3-19) and 2,590 g TEQDF-WHO98 (2,440 g I-TEQDF) in 1987 (Table 3-18). 

Because the activity level and emission factors had low confidence ratings, the emission 

estimates for all years were assigned a low confidence rating, i.e., a Category C. Figures 3-15 

and 3-16 display the congener and congener group profiles of MWIs without APCDs and those 

equipped with WSs and FFs, respectively. 

3.3.5. Recent EPA Regulatory Activities 

In September 1997, EPA promulgated final regulations under the Clean Air Act 

Amendments limiting CDD/CDF stack emissions from MWIs (Federal Register, 1997a). These 

emission limits are specific to the sum of CDD and CDF emissions (the sum of tetra- through 

octa-CDDs and CDFs). For either new or existing MWIs that were operational before or after 

June 20, 1996, EPA limits the total CDD/CDF concentration in the stack gases to 2.3 ng/dscm. 

This would require the application of WSs, DSI of activated carbon combined with FFs and/or 

SDs/FFs. EPA expects that many facilities that currently operate on-site incinerators will switch 

to less expensive methods of treatment and disposal of medical and infectious waste when faced 

with the compliance costs associated with the emission standards for MWIs. EPA projects that, 

following full compliance with these standards, annual emissions from MWIs will be 5 to 7 g ITEQDF/yr. 

3.4. CREMATORIA 

3.4.1. Human Crematoria 

3.4.1.1 Emissions Data 

Bremmer et al. (1994) measured CDD/CDF emissions at two crematoria in the 

Netherlands. The first, a “cold”-type furnace with direct, uncooled emissions, was calculated to 

yield 2,400 ng I-TEQDF per body. In the cold-type furnaces, the coffin is placed inside at a 

temperature of about 300°C. The temperature of the chamber is then increased to 800 to 900°C 

using a burner and kept there for 2 to 2.5 hr. The second furnace, a “warm” type in which flue 

gases are cooled to 220°C prior to discharge, was calculated to yield 4,900 ng I-TEQDF per body. 

In the warm-type furnace, the coffin is placed in a chamber preheated to 800°C or higher for 1.2 

to 1.5 hr. The chamber exhausts from both furnace types were incinerated in an afterburner at a 

3-65
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.01 0.02 0.03 0.04 0.05 0.06 0.07 0.08 0.09 

Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 0.25 

Figure 3-15. Congener and congener group profiles for air emissions from 

medical waste incinerators without air pollution control devices (nondetects 

set equal to zero). 

3-66
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.01 0.02 0.03 0.04 0.05 0.06 0.07 0.08 0.09 

Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 

Figure 3-16. Congener and congener group profiles for air emissions from 

medical waste incinerators equipped with a wet scrubber and fabric filter. 

3-67
temperature of about 850°C. The higher emission rate for the warm-type furnace was attributed 

by the authors to the formation of CDDs/CDFs during the intentional cooling of the flue gases to 

220°C. 

Jager et al. (1992) (as reported in Bremmer et al., 1994) measured an emission rate of 

28,000 ng I-TEQDF per body for a crematorium in Berlin, Germany. No operating process 

information was provided by Bremmer et al. for the facility. 

Mitchell and Loader (1993) reported even higher emission factors for two crematoria in 

the United Kingdom. The first facility tested was manually operated and had primary and 

secondary combustion chambers preheated to 650°C and a residence time of 1 sec in the 

secondary combustion chamber. The second tested facility was computer controlled and had 

primary and secondary combustion chambers heated to 850°C and a residence time of 2 sec in 

the secondary combustion chamber. The measured stack gas TEQ concentrations ranged from 42 

to 71.3 ng I-TEQDF/m3

 (at 11% oxygen) at the first facility and from 25.4 to 45.5 ng I-TEQ /m3 

DF 

(at 11% oxygen) at the second facility. Emission factors based on these test results and gas 

generation rates reported by Bremmer et al. (1994) were calculated to range from 70,000 to 

80,000 ng I-TEQDF/body (HMIP, 1995). 

Takeda et al. (1998) measured CDD/CDF emissions at 10 crematoria in Japan. Although 

there are more than 1,600 crematoria in Japan, the 10 tested facilities handle 4% of the 

cremations carried out in Japan annually. A wide range of CDD/CDF emissions were observed. 

When nondetect values were treated as zero, the emission factor range was 42 to 62,000 ng ITEQDF/body (mean of 9,200 ng I-TEQDF/body). When nondetect values were treated as one-half 

the DL, the range was 450 to 63,000 ng I-TEQDF/body (mean of 11,000 ng I-TEQDF/body). 

To obtain more data on CDD/CDF emissions from crematoria in Japan, Takeda et al. 

(2001) measured CDD/CDF emissions at 17 additional crematoria. In that study, all the 

crematoria except one had secondary combustion chambers. Additionally, one crematorium had 

a secondary combustion chamber but did not use it. One to four main chambers were connected 

to the secondary chambers, and the temperature of the main chambers ranged from 

approximately 650 to 1,150°C. In most cases, only one body was cremated at time. However, 

between two and four bodies were cremated at four sampling events. A coffin and any 

accompanying materials were combusted along with the body. Emission factors ranged from 120 

to 24,000 ng I-TEQDF/body. In general, as the average temperature in the main combustion 

chamber increased, CDD/CDF emissions decreased. However, the crematorium that had a 

secondary combustion chamber but did not use it had both high temperatures in the main 

combustion chamber and high CDD/CDF emissions. Additionally, with the rise of the average 

temperature in the secondary combustion chamber of the eight crematoria without dust 

collectors, CDD/CDF emissions decreased. For crematoria with dust collectors, the relationship 

3-68
between the average temperature in the secondary combustion chamber and CDD/CDF emissions 

was not clear. 

EPA obtained test data from two crematoria for humans operating in the United States, 

one at Camellia Memorial Lawn in California (CARB, 1990a) and one at Woodlawn Cemetery in 

New York (U.S. EPA, 1999a). Additionally, EPA obtained test data from one crematorium for 

animals operating in the United States: University of Georgia Veterinary School (U.S. EPA, 

2000a); however, it is not appropriate to use the emission factors from this facility to characterize 

emissions associated with human cremation. 

Testing at the Camellia Memorial Lawn crematorium, which is classified as a warm-type 

facility using the criteria of Bremmer et al. (1994), was conducted in 1990 (CARB, 1990a). The 

combusted material at this facility consisted of the body, as well as 4 lb of cardboard, up to 6 lb 

of wood, and an unquantified amount of unspecified plastic wrapping. The three emissions tests 

conducted at this facility, which operates using an afterburner, yielded an average emission factor 

of 543 ng TEQDF-WHO98/body (501 ng I-TEQDF/body). Table 3-21 presents the congenerspecific emission factors for this facility. 

Testing at Woodlawn Cemetery, which has a crematorium with a primary combustion 

chamber, a secondary combustion chamber, and a scrubber APCD, was conducted in 1995. Tests 

were run at three secondary combustion chamber temperatures: 675, 870, and 980°C (U.S. EPA, 

1999a). The combusted material consisted of the body, as well as a 10- to 100-lb casket 

constructed of fiberboard, particle board, or wood and various body wrappings and articles such 

as a plastic sheet, a cloth sheet, or clothes. For this facility, average emission factors of 362 and 

709 ng TEQDF-WHO98/body cremated (348 and 638 ng I-TEQDF/body cremated) were calculated, 

based on emissions collected at the scrubber inlet and outlet, respectively. The congener-specific 

emission factors for this facility are shown in Table 3-22. 

In 1995, 1,155 crematoria were reported to be operating in the United States; this number 

had decreased to approximately 1,060 by 2000. To determine whether the emissions data 

collected at the Woodlawn Cemetery facility are representative of a typical crematorium 

operating in the United States, representatives from the Cremation Association of North America 

(CANA) were contacted to identify the typical operating conditions at U.S. crematoria. 

According to the CANA representatives, all crematoria operating in the United States have 

primary and secondary combustion chambers. Additionally, crematoria with operating 

conditions that indicate the presence of an afterburner are considered to contain secondary 

combustion chambers. The primary and secondary combustion chambers at U.S. crematoria 

typically operate at between 675 and 870°C, but many operate at 980°C, as required by their 

respective states. 

3-69
Table 3-21. Congener-specific profile for Camellia Memorial Lawn 

crematorium 

Congener/congener 

group 

Mean facility emission factor (ng/body) 

Assuming nondetect 

set to zero 

Assuming nondetect set to 

1⁄2 detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

28.9 

89.6 

108 

157 

197 

1,484 

2,331 

28.9 

89.6 

108 

157 

197 

1,484 

2,331 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

206 

108 

339 

374 

338 

657 

135 

1,689 

104 

624 

206 

117 

349 

374 

338 

657 

135 

1,813 

112 

624 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

4,396 

4,574 

501.8 

544.1 

4,396 

4,725 

508.6 

550.9 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

554 

860 

2,224 

3,180 

2,331 

4,335 

2,563 

4,306 

2,030 

624 

554 

860 

2,224 

3,180 

2,331 

4,335 

2,563 

4,306 

2,154 

624 

Total CDD/CDF 23,007 23,131 

Source: CARB (1990a). 

3-70
Table 3-22. Congener-specific profile for the Woodlawn Cemetery 

crematorium 

Congener 

Mean emission factor, 

scrubber inlet (ng/body) 

Mean emission factor, 

scrubber outlet (ng/body) 

Nondetect set 

to zero 

Nondetect set to 1⁄2 

detection limit 

Nondetect 

set to zero 

Nondetect set to 1⁄2 

detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

11 

31 

74 

115 

83 

724 

1,120 

12 

44 

74 

115 

83 

724 

1,120 

39 

168 

239 

565 

524 

1,253 

10,698 

45 

364 

258 

603 

553 

1,302 

1,154 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

106 

116 

285 

263 

278 

146 

466 

962 

165 

435 

106 

116 

285 

264 

278 

146 

466 

963 

165 

435 

256 

150 

409 

252 

253 

139 

429 

872 

142 

3,499 

279 

170 

463 

280 

282 

148 

474 

948 

148 

363 

Total I-TEQDF 

Total TEQDF-WHO98 

348 

362 

356 

376 

638 

709 

780 

961 

Source: U.S. EPA (1999a). 

Only one or two facilities in the United States incorporate the use of an APCD, such as a 

scrubber. Therefore, the inlet dioxin emission factors rather than the outlet dioxin emission 

factors at the Woodlawn crematorium would be representative of a typical crematorium operating 

in the United States (telephone conversation between Allen Krobath, CANA, and K. Riley, 

Versar, Inc., February 12, 2003, and telephone conversation between Dale Walter, Matthews 

Cremation, and K. Riley, Versar Inc., February 13, 2003). 

In the previous inventory, an average emission factor of 17,000 ng I-TEQDF/body 

(assuming nondetect values were zero) was developed, based on emission factors measured for 

16 of the tested facilities, including the one at Camellia Memorial Lawn (CARB, 1990a), the 10 

Japanese facilities (Takeda et al., 1998), the two Dutch facilities (Bremmer et al., 1994), the one 

German facility (Jager et al., 1992), and the two British facilities (Mitchell and Loader, 1993). 

The more recent data provided by Takeda et al. (2001) for the 17 Japanese facilities support the 

emission factor of 17,000 ng I-TEQDF/body. However, an average emission factor developed 

3-71
using the data reported for the two U.S. crematoria (i.e., the outlet values for the Camellia 

Memorial Lawn facility and the inlet values for the Woodlawn Cemetery facility) is 453 ng 

TEQDF-WHO98/body (425 I-TEQDF/body cremated), assuming nondetect values were zero. 

These values are two orders of magnitude less than the overall average calculated above. An 

examination of the differences in U.S. and foreign operating practices may provide a rationale 

for the large discrepancies. 

Bremmer et al. (1994) reported an emission factor of 2,400 ng I-TEQDF/body for a Dutch 

facility with a cold-type furnace and an emission factor of 4,900 ng I-TEQDF/body for another 

Dutch facility with a warm-type furnace where flue gases were cooled to 220°C. Neither of the 

U.S. facilities are considered to have cold-type furnaces. Additionally, the flue gases at the 

Camellia Memorial Lawn crematorium were not cooled prior to exiting the furnace. At the 

Woodlawn Cemetery facility, the flue gases were cooled from 681 to 860°C prior to entering the 

scrubber to 271 to 354°C by the time they exited the scrubber and the furnace. The emissions 

were higher at the scrubber outlet than at the inlet (961 vs. 325 ng TEQDF-WHO98/body [780 vs. 

319 I-TEF/body]); however, the emissions were not of the same magnitude as those reported by 

Bremmer for the warm-type facility (4,900 ng I-TEQDF/body). The Jager et al. (1992) report did 

not include operating process information; therefore, the German facility could not be compared 

with the U.S. facilities. Additionally, the emission values derived from the Mitchell and Loader 

(1993) emission concentrations were calculated using gas generation rates from the Bremmer et 

al. report and, as such, may not be indicative of crematoria in the United States. 

In the Takeda et al. (1998, 2001) reports, the burn time for the cremations varied from 47 

to 117 min. The average burn time in the U.S. studies was 120 min. This shorter burn time may 

not be optimal for dioxin reduction, resulting in higher dioxin emissions. Also, the secondary 

combustion chamber temperatures ranged from 250 to 950°C in the Takeda studies, again 

resulting in higher emission rates. In fact, in Takeda et al. (2001) two of the three runs that had 

the highest TEQ concentrations per body came from a crematorium that did not use a secondary 

combustion chamber. Of the 31 crematoria sampled in Takeda et al. (2001), 26 had lower than 

5,000 ng I-TEQDF/body. 

Because the Woodlawn facility is unique in that it incorporates an APCD, the sample data 

for the air stream entering the scrubber versus the stream exiting the scrubber should be analyzed. 

A comparison of the dioxin concentrations of these air streams shows a significant increase in 

dioxin concentrations in the stream exiting the scrubber. This increase can be attributed to the 

decrease in temperature that occurred in the scrubber. Upon exiting the scrubber, the flue gas 

temperatures were in the range of 271 to 354°C, compared with temperatures of between 681 and 

860°C at the scrubber inlet. As discussed in Section 2, these exit flue gas temperatures lie in the 

3-72
optimum temperature range for dioxin formation; therefore, an increase in dioxin concentrations 

would be expected. 

An analysis of scrubber inlet dioxin data indicates that the average dioxin concentrations 

increased with temperature (189, 445, and 503 ng TEQDF-WHO98/body at 681, 772, and 860°C, 

respectively). Because the operating temperatures are outside the temperature range for the 

formation of dioxin (200 to 400°C), dioxin concentrations should decrease as temperatures 

increase. Further analysis of the data shows that as temperatures at the scrubber inlet increased, 

so did concentrations of PM, HCl, and lead (Table 3-23). The data also indicate that oxygen 

levels decreased as the temperature increased (U.S. EPA, 1999a). Given these data, one could 

speculate that as the temperature increased, incomplete combustion conditions arose, leading to 

an increase in dioxin formation. 

Table 3-23. Operational data for the WoodlawnCemetery crematorium, 

scrubber inlet 

Parameter 

Mean value 

675°C 870°C 980°C 

Particulate matter (gr/dscf @ 7% O2) 

Hydrochloric acid (lb/hr) 

Lead (g/hr) 

Oxygen (%) 

0.015 

0.053 

0.1 

9.9

0.033 

0.14 

0.32 

8.6 

0.068 

0.26 

0.59 

7.5 

 Source: U.S. EPA (1999a). 

Using data from U.S. crematoria, EPA recommends an average emission factor of 453 ng 

TEQDF-WHO98/body (425 ng I-TEQDF/body). This is derived from the scrubber inlet dioxin 

concentrations from the Woodlawn Cemetery study and the results from the Camellia Memorial 

Lawn study. These average congener-specific emission are presented in Table 3-24, and the 

CDD/CDF congener and congener group emission profiles are presented in Figure 3-17. 

Because the emission factor was derived using emissions data from only 2 of 1,060 crematoria, 

the average emission factor is assigned a low confidence rating. 

3.4.1.2. Activity Level Information 

A total of 323,371 cremations were performed in reference year 1987, 488,224 in 1995, 

and 629,362 in 2000. A high confidence rating is assigned to these activity level estimates 

because they are based on comprehensive data provided by CANA (CANA, 2006). 

3-73
Table 3-24. Congener-specific profile for the Camellia Memorial Lawn 

crematorium and the Woodlawn Cemetery crematorium 

Congener/congener 

group 

Mean facility emission factor (ng/body) 

Nondetect set to zero Nondetect set to 1⁄2 detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD

1,2,3,7,8,9-HxCDD

1,2,3,4,6,7,8-HpCDD 

OCDD 

20 

60.3 

91 

136 

140 

1,104 

1,725.5 

20.5 

66.8 

91 

136 

140 

1,104 

1,725.5 

2,3,7,8-TCDF

1,2,3,7,8-PeCDF

2,3,4,7,8-PeCDF

1,2,3,4,7,8-HxCDF

1,2,3,6,7,8-HxCDF

1,2,3,7,8,9-HxCDF

2,3,4,6,7,8-HxCDF

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF

OCDF

 156 

112 

312 

318.5 

308 

401.5 

300.5 

1,325.5 

134.5

 529.5

156 

116.5 

317 

319 

308 

401.5 

300.5 

1,388 

138.5 

529.5 

Total I-TEQDF

Total TEQDF-WHO98

 424.8 

452.9

431.9 

463.3 

Total TCDD

Total PeCDD

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF

 330

 488

1,254.5 

1,721.5 

1,304.5 

2,240.5 

1,514.5 

2,634 

1,097.5 

529.5

 330 

488 

1,254.5 

1,721.5 

1,304.5 

2,240.5 

1,514.5 

2,634 

1,097.5 

529.5 

Total CDD/CDF 13,114.5 13,114.5 

Sources: CARB (1990a); U.S. EPA (1999a). 

3-74
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.01 0.02 0.03 0.04 0.05 0.06 0.07 0.08 0.09 0.1

Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2

4D 2378 

5D 12378 

6D 123478 

6D 123678 

6D 123789 

7D 1234678 

8D 

4F 2378 

5F 12378 

5F 23478 

6F 123478 

6F 123678 

6F 123789 

6F 234678 

7F 1234678 

7F 1234789 

8F 

4D Total 

5D Total 

6D Total 

7D Total 

8D 

4F Total 

5F Total 

6F Total 

7F Total 

8F 

Figure 3-17. Congener and congener group profiles for air emissions from 

the Camellia Memorial Lawn crematorium and Woodlawn Cemetery 

crematorium. 

Sources: CARB (1990a); U.S. EPA (1999a). 

3-75
3.4.1.3. Emission Estimates 

Combining the average emission rate of 453 ng WHO-TEQ98/body (425 ng ITEQDF/body) with the number of cremations in 1987, 1995, and 2000 (323,371; 488,224; and 

629,362, respectively) yields an estimated annual release of 0.15 g TEQDF-WHO98 (0.14 g ITEQDF) in 1987, 0.22 g TEQDF-WHO98 (0.2 g I-TEQDF) in 1995, and 0.29 g TEQDF-WHO98 (0.27 

g I-TEQDF) in 2000. An overall confidence rating of low was assigned to the emissions because 

the emission factor had a low rating. 

3.4.2. Animal Crematoria 

3.4.2.1. Emissions Data 

Only one study that measured CDD/CDF emissions from animal cremation could be 

located. In 1999, CDD/CDF emissions from a newly installed animal incineration unit located at 

the University of Georgia Veterinary School were measured (U.S. EPA, 2000a). The 

incineration unit, which consists of a primary and a secondary combustion chamber, is used to 

dispose of animals (mostly cows and horses) used in experimentation. Emissions are 

uncontrolled, with the exception of an NFPA spark screen located at the stack outlet. Based on 

four test runs, the average TEQ emission factor was 0.12 TEQDF-WHO98/kg (0.11 ng I-TEQDF/kg) 

of animal cremated. The average emission factors for these test runs are provided in Table 3-25 

and a congener-specific profile based on these data is provided as Figure 3-18. 

3.4.2.2. Activity Level Information 

As part of the 2000 inventory, OAQPS calculated a national animal cremation activity 

level estimate of 81.9 million kg/yr for reference year 2000. This estimate was scaled from the 

1999 activity level estimate by applying the ratio of the 2000 national human population 

(281,421,906) to the 1999 national human population (249,440,000). The 1999 national activity 

level was based on 1990 data provided by OAQPS’ Emission Standards Division. The 1999 and 

2000 activity level estimates assume that animal mortality and cremation rates are constant and 

that the animal population is directly proportional to human population. 

3.4.2.3. Emission Estimates 

Applying the TEQ emission factor of 0.12 ng TEQDF-WHO98/kg (0.11 ng I-TEQDF/kg) of 

animal combusted to the activity level estimated by OAQPS (81.9 million kg/yr) yields estimated 

annual emissions of 0.0098 g TEQDF-WHO98 (0.009 g I-TEQDF) in 2000. This estimate does not 

include events such as the mass burning of animals affected by mad cow disease. These 

estimates are based on extremely limited data and should be regarded as preliminary indications 

of possible emissions from this source; further testing is needed to confirm the true magnitude of 

the emissions. 

3-76
Table 3-25. Congener-specific profile for the University of Georgia 

Veterinary School 

Congener/congener group 

Mean facility emission factor (ng/kg of animal) 

Nondetect set to zero Nondetect set to 1⁄2 detection limit

 2,3,7,8-TCDD 

 1,2,3,7,8-PeCDD 

 1,2,3,4,7,8-HxCDD 

 1,2,3,6,7,8-HxCDD 

 1,2,3,7,8,9-HxCDD 

 1,2,3,4,6,7,8-HpCDD 

OCDD 

7.51e!03 

2.13e!02 

4.46e!03 

8.86e!03 

7.17e!03 

5.03e!03 

1.01e!03 

7.51e!03

2.13e!02

4.46e!03

8.86e!03

7.17e!03

5.03e!03

1.01e!03

 2,3,7,8-TCDF 

 1,2,3,7,8-PeCDF 

 2,3,4,7,8-PeCDF 

 1,2,3,4,7,8-HxCDF 

 1,2,3,6,7,8-HxCDF 

 1,2,3,7,8,9-HxCDF 

 2,3,4,6,7,8-HxCDF 

 1,2,3,4,6,7,8-HpCDF 

 1,2,3,4,7,8,9-HpCDF 

OCDF 

1.79e!02 

6.70e!03 

1.41e!01 

2.93e!02 

1.85e!02 

7.44e!02 

2.35e!02 

4.20e!03 

3.16e!03 

2.00e!04 

1.79e!02

6.70e!03

1.41e!01

2.93e!02

1.85e!02

7.44e!02

2.35e!02

4.20e!03

3.16e!03

2.00e!04

 Total CDD/CDF 0.37 0.37 

 Total I-TEQDF 

 Total TEQDF-WHO98 

0.11 

0.12 

0.11 

0.12 

 Source: U.S. EPA (2000a). 

3.5. SEWAGE SLUDGE INCINERATION 

The three principal combustion technologies used to incinerate sewage sludge in the 

United States are multiple-hearth incineration, fluidized-bed incineration, and electric furnace 

incineration (Brunner, 1992; U.S. EPA, 1995a). All of these technologies are excess-air 

processes (i.e., they combust sewage sludge with oxygen in excess of theoretical requirements). 

Approximately 80% of operating sludge incinerators are multiple-hearth design, about 20% are 

fluidized-bed incinerators, and fewer than 1% are electric incinerators. Other types of furnaces 

not widely used in the United States are single-hearth cyclones, rotary kilns, and high-pressure, 

wet-air oxidation units (U.S. EPA, 1997a; e-mail dated July 13, 1998, from K. Maw, Pacific 

Environmental Services, to G. Schweer, Versar, Inc.). 

Multiple-hearth incinerators. These types of furnaces consist 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 

3-77
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 0.25 0.3 0.35 0.4 

4D 2378 

5D 12378 

6D 123478 

6D 123678 

6D 123789 

7D 1234678 

8D 

4F 2378 

5F 12378 

5F 23478 

6F 123478 

6F 123678 

6F 123789 

6F 234678 

7F 1234678 

7F 1234789 

8F 

Figure 3-18. Congener profile for air emissions from the University of 

Georgia animal crematorium. 

Source: U.S. EPA (2000a). 

the furnace. Moisture is evaporated from the sludge cake in the upper hearths. The center 

hearths are the burning zone, 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 venturi scrubber, an 

impingement tray scrubber, or a combination of both. Wet cyclones and dry cyclones are also 

used (U.S. EPA, 1995a). 

Fluidized-bed incinerators. A 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 exits the furnace with the combustion gases; therefore, air pollution control systems 

3-78
typically consist of high-energy venturi scrubbers or venturi/impingement tray combinations 

(U.S. EPA, 1995a). 

Electric furnaces. Also called infrared furnaces, these consist 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 located just above the traveling belt within the furnace. Electric power 

is required to initiate and sustain combustion. Emissions are usually controlled with a venturi 

scrubber or some other WS (Brunner, 1992; U.S. EPA, 1995a). 

3.5.1. Emissions Estimates from Sewage Sludge Incinerators 

EPA measured CDD/CDF emissions at three multiple-hearth incinerators as part of Tier 4 

of the National Dioxin Survey (U.S. EPA, 1987a). During the pretest surveys, two of the 

facilities were judged to have “average” potential and one facility was judged to have “high” 

potential for CDD/CDF emissions with respect to other sewage sludge incinerators. The results 

of these tests include congener group concentrations in stack gas but lack measurements for 

specific congeners other than 2,3,7,8-TCDD and 2,3,7,8-TCDF. The results show a wide 

variability in the emission factors at these three facilities; total CDD/CDF emission factors 

ranged from 90 to 3,400 ng/kg (average of 1,266 ng/kg). Total TEQ emissions could not be 

determined for these facilities because of the lack of congener-specific data. 

In 1990, EPA measured CDD/CDF emissions (including all 17 toxic congeners) at 

another multiple-hearth incinerator and also at a fluidized-bed incinerator (U.S. EPA, 1990d). 

Assuming nondetects were zero, the total CDD/CDF emission factors for these two facilities 

were 79 and 846 ng/kg, and the total average TEQ emission factors were 3.6 and 43.2 ng TEQDFWHO98/kg (2.4 and 43.3 ng I-TEQDF/kg) of dry sludge. In 1995, the Association of Metropolitan 

Sewerage Agencies (AMSA) submitted to EPA the results of stack tests conducted at an 

additional 13 sewage sludge incinerators (Green et al., 1995). Two of these data sets were 

considered not usable by EPA because either DLs or feed rates and stack flow rates were not 

provided. As with the EPA-tested facilities (U.S. EPA, 1987a, 1990d), wide variability was 

observed in the emission factors for the 11 AMSA facilities. Assuming nondetects were zero, 

total CDD/CDF emission factors ranged from 0 to 1,392 ng/kg (average of 217 ng/kg), and total 

average TEQ emission factors ranged from 0 to 16 ng TEQDF-WHO98/kg (average, 3.47 ng) (3.46 

ng I-TEQDF/kg) of dry sludge. 

In 1999, stack tests were conducted at a multiple-hearth incinerator equipped with a 

venturi scrubber and a three-tray impingement conditioning tower (U.S. EPA, 2000b). Four test 

runs were conducted; however, the first test run was aborted, and the CDD/CDF results from the 

fourth test run were determined to be statistical outliers (p>0.05). The back-half emission 

3-79
concentrations for test run 4 were 50 to 60% lower than those for test runs 2 and 3. Overall, total 

CDD/CDF emissions measured during test run 4 were 48 ng/kg, whereas total CDD/CDF 

emissions measured during test runs 2 and 3 were 120 and 116 ng/kg, respectively. It could not 

be determined whether the lower concentrations associated with test run 4 were due to analyte 

loss or whether they represented an accurate reflection of a change in incinerator emission 

releases. 

The average TEQ emission factor, excluding test run 4, was 3.28 ng TEQDF-WHO98/kg 

(3.23 ng I-TEQDF/kg). The average TEQ emission factor based on the data for the 11 AMSA 

facilities (Green et al., 1995) and the three facilities reported by EPA (U.S. EPA, 2000b, 1990d) 

is 6.74 ng TEQDF-WHO98/kg (6.65 ng I-TEQDF/kg) of dry sludge, assuming nondetect values are 

zero. Figure 3-19 presents the average congener and congener group profiles based on these 

data. Additionally, Table 3-26 presents the average congener and congener-specific group 

emission factors and the average TEQ emission factors for these facilities. Table 3-26 also 

presents 2,3,7,8-TCDD, 2,3,7,8-TCDF, and congener-specific group emission factors for the 

three facilities reported by EPA (U.S. EPA, 1987a). 

Studies from other countries have reported similar results. Bremmer et al. (1994) 

reported an emission rate of 5 ng I-TEQDF/kg for a fluidized-bed sewage sludge incinerator in the 

Netherlands that was equipped with a cyclone and a WS. Cains and Dyke (1994) measured 

CDD/CDF emissions at two sewage sludge incinerators in the United Kingdom. The emission 

rate at an incinerator equipped with an ESP and a WS ranged from 2.75 to 28 ng I-TEQDF/kg. 

The emission rate measured at a facility equipped with only an ESP was 43 ng I-TEQDF/kg. 

In 1988, approximately 199 sewage sludge incineration facilities combusted abo 0.865 

million metric tons of dry sewage sludge (Federal Register, 1993a). In 1995, approximately 257 

sewage sludge incinerators (some of which were backup or alternate incinerators) combusted 

about 2.11 million dry metric tons of sewage sludge (e-mail dated July 13, 1998, from K. Maw, 

Pacific Environmental Services, to G. Schweer, Versar, Inc.). Using trends in wastewater flow 

rates from the 1988 National Sewage Sludge Survey and the 1984 to 1996 Needs Surveys, EPA 

estimated that in 2000 approximately 6.4 million metric tons of dry sewage sludge would be 

generated (U.S. EPA, 1999b). Of this amount, EPA projected that 22% (1.42 million metric 

tons) would be incinerated. 

According to EPA, sewage sludge generation would increase to 6.9 million dry tons in 

2005 and 7.4 million dry tons in 2010; however, the percentage of sewage sludge incinerated 

will decrease slightly, to 20% in 2005 and 19% in 2010. EPA estimates that approximately 1.38 

million metric tons of dry sewage sludge would be incinerated in 2005 and 1.41 million metric 

tons will be incinerated in 2010. EPA believes that incineration as a disposal method for sewage 

sludge will decrease as a result of increasing costs and public concerns about the environmental 

and health impacts associated with incineration. 

3-80
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 0.25 0.3

Ratio (congener group emission factor/total CDD/CDF emission factor) 

0 0.1 0.2 0.3 0.4 

4D 2378 

5D 12378 

6D 123478 

6D 123678 

6D 123789 

7D 1234678 

8D 12346789 

4F 2378 

5F 12378 

5F 23478 

6F 123478 

6F 123678 

6F 123789 

6F 234678 

7F 1234678 

7F 1234789 

8F 12346789 

4D Total 

5D Total 

6D Total 

7D Total 

8D 12346789 

4F Total 

5F Total 

6F Total 

7F Total 

8F 12346789 

Figure 3-19. Congener and congener group profiles for air emissions from 

sewage sludge incinerators. 

3-81
Table 3-26. CDD/CDF emission factors for sewage sludge incinerators 

Congener 

Mean emission factor 

(ng/kg) 

for U.S. EPA (1987a) 

(3 facilities) 

Mean emission factor (ng/kg) for 

Green et al. (1995) (11 facilities) 

U.S. EPA (1990d) (2 facilities) 

U.S. EPA (1999b) (1 facility) 

Nondetect 

set to zero 

Nondetect set to 

1⁄2 detection 

limit 

Nondetect set 

to zero 

Nondetect set to 

1⁄2 detection limit

 2,3,7,8-TCDD

 1,2,3,7,8-PeCDD

 1,2,3,4,7,8-HxCDD

 1,2,3,6,7,8-HxCDD

 1,2,3,7,8,9-HxCDD

 1,2,3,4,6,7,8-HpCDD

 OCDD

 0.39 

NR 

NR 

NR 

NR 

NR 

46.2 

0.44 

NR 

NR 

NR 

NR 

NR 

46.2 

0.16

 0.22

 0.04

 0.12

 0.29

 2.46

 12.78

 0.26

 0.3 

0.11

 0.17

 0.35

 2.59

 13.16

 2,3,7,8-TCDF

 1,2,3,7,8-PeCDF

 2,3,4,7,8-PeCDF

 1,2,3,4,7,8-HxCDF

 1,2,3,6,7,8-HxCDF

 1,2,3,7,8,9-HxCDF

 2,3,4,6,7,8-HxCDF

 1,2,3,4,6,7,8-HpCDF

 1,2,3,4,7,8,9-HpCDF

 OCDF 

179 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

109 

179 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

109 

25.41

 1.92

 6.47

 2.11

 0.77

 0.03

 1.22

 1.46

 0.17

 1.17

 25.41

 1.92

 6.47

 2.11

 0.77

 0.03

 1.22

 1.46

 0.17

 1.17

 Total TCDD

 Total PeCDD

 Total HxCDD

 Total HpCDD

 Total OCDD

 Total TCDF

 Total PeCDF

 Total HxCDF

 Total HpCDF

 Total OCDF 

37.6 

2.66 

16.6 

53.9 

46.2 

528 

253 

75.4 

144 

109 

37.7 

2.81 

16.9 

54 

46.2 

528 

253 

75.9 

144 

109 

35.8 

1.11

 1.74

 4.39

 12.78 

123.85

 59.94

 12.69

 2.63

 1.17

 37.81

 1.63

 2.25

 5.03

 13.16 

124.1 

60.16

 13.5 

3.12

 1.55

 Total I-TEQDF

 Total TEQDF-WHO98 

NR 

NR 

NR 

NR

 6.65

 6.74 

6.87

 7.01

 Total CDD/CDF 1,266 1,268 256 262 

NR = Not reported 

3-82
A medium confidence rating is assigned to the average TEQ emission factor because it 

was derived from stack testing at 14 U.S. sewage sludge incinerators. The 1988 activity level 

estimate (used as a surrogate for the 1987 activity level) and the 2000 activity level estimate are 

assigned a high confidence rating because they are based on extensive EPA surveys to support 

rule-making activities. The 1995 activity level estimate is assigned a medium confidence rating 

because assumptions were made for numerous facilities concerning hours of operation, operating 

capacity, and design capacity. 

Using the above estimated amounts of sewage sludge incinerated per year and the average 

TEQ emission factor of 6.74 ng TEQDF-WHO98/kg (6.65 ng I-TEQDF/kg), the estimate of TEQ 

emissions to air is 5.8 g TEQDF-WHO98 (5.8 g I-TEQDF) in 1987, 14.2 g TEQDF-WHO98 (14 g ITEQDF) in 1995, and 9.6 g TEQDF-WHO98/kg (9.4 g I-TEQDF/kg) in 2000. Because the emission 

factor had a medium confidence rating, the overall emission estimates were assigned a medium 

confidence rating for all years. 

3.5.2. Solid Waste from Sewage Sludge Incinerators 

In Table 5-16 of U.S. EPA (1987a), data are presented indicating that 2,3,7,8-TCDD was 

not detected in the bottom ash or scrubber water filtrate from three sewage sludge incinerators. 

However, total CDDs for the three incinerators and the filtrate were nondetects, 20 ng/kg, 10 

ng/kg, and 0.3 ng/kg, respectively. For total CDFs, the respective values were nondetects, 70 

ng/kg, 50 ng/kg, and 4 ng/kg. No data were given for any congeners (other than 2,3,7,8-TCDD), 

nor were there any data on the quantities of ash or filtrate. 

3.6. TIRE COMBUSTION 

Most discarded tires are combusted in dedicated tire incinerators or cement kilns. Some 

are combusted as auxiliary fuel in industrial boilers and in pulp and paper mill combustion 

facilities. Additionally, tires may be unintentionally burned in an uncontrolled fashion at 

landfills (open burning). This section addresses the total TEQ emissions that may result from the 

combustion of tires in dedicated tire incinerators, industrial boilers, and pulp and paper mill 

combustion facilities, but excludes cement kilns (addressed in Section 5.1). The open burning of 

tires is not discussed in this report due to the lack of information. 

Emissions of CDDs/CDFs from the incineration of discarded automobile tires were 

measured at a dedicated tire incinerator tested by the California Air Resources Board (CARB, 

1991). The facility consists of two excess air furnaces equipped with steam boilers to recover the 

energy from the heat of combustion. Whole tires were fed to the incineration units at rates 

ranging from 2,800 to 5,700 kg/hr during the three test days. The facility was equipped with a 

DS and an FF for the control of emissions prior to exiting the stack. Table 3-27 presents the 

congener-specific emission factors for this facility. Figure 3-20 presents CDD/CDF congener 

3-83
Table 3-27. CDD/CDF air emission factors for a tire combustion facility 

Congener/congener 

group 

Mean facility emission factor (ng/kg) 

Assuming nondetect set to zero 

Assuming nondetect set to 

1⁄2 detection limit

 2,3,7,8-TCDD

 1,2,3,7,8-PeCDD

 1,2,3,4,7,8-HxCDD

 1,2,3,6,7,8-HxCDD

 1,2,3,7,8,9-HxCDD

 1,2,3,4,6,7,8-HpCDD

 OCDD

 0.149

 0.006

 0.018

 0.055

 0.036

 0.379

 4.156

 0.149

 0.026

 0.023

 0.062

 0.048

 0.379

 4.156

 2,3,7,8-TCDF

 1,2,3,7,8-PeCDF

 2,3,4,7,8-PeCDF

 1,2,3,4,7,8-HxCDF

 1,2,3,6,7,8-HxCDF

 1,2,3,7,8,9-HxCDF

 2,3,4,6,7,8-HxCDF

 1,2,3,4,6,7,8-HpCDF

 1,2,3,4,7,8,9-HpCDF

 OCDF

 0.319

 0.114

 0.086

 0.103

 0.059

 0.036

 0.1

 0.0

 0.027

 0.756

 0.319

 0.118

 0.091

 0.111

 0.09

 0.068

 0.148

 0.166

 0.095

 0.756

 Total 2,3,7,8-CDD

 Total 2,3,7,8-CDF

 Total I-TEQDF

 Total TEQDF-WHO98

 4.799

 1.6

 0.282

 0.281

 4.843

 1.962

 0.312

 0.320

 Total TCDD

 Total PeCDD

 Total HxCDD

 Total HpCDD

 Total OCDD

 Total TCDF

 Total PeCDF

 Total HxCDF

 Total HpCDF

 Total OCDF

 0.153

 0.032

 0.391

 0.695

 4.156

 1.204

 0.737

 0.71

 0.119

 0.802

 0.153

 0.032

 0.391

 0.695

 4.156

 1.204

 0.737

 0.71

 0.186

 0.802

 Total CDD/CDF 8.999 9.067

 Source: CARB (1991). 

3-84
Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.1 0.2 0.3 0.4 0.5 

Ratio (congener emission factor/total CDD/CDF emission factor) 

0 0.1 0.2 0.3 0.4 0.5 

Figure 3-20. Congener and congener group profiles for air emissions from a 

tire combustor. 

3-85
and congener group profiles based on these TEQ emission factors. From these data, the average 

emission factor is estimated to be 0.281 ng TEQDF-WHO98/kg (0.282 ng I-TEQDF/kg) of tires 

incinerated (when all nondetect values are treated as zero). This emission factor was used to 

estimate annual TEQ releases from the tire combustion source category for the years 1987, 1995, 

and 2000. 

EPA assigned a low confidence rating to the estimated TEQ emission factor because it is 

possible that it is not representative of TEQ emissions from all tire combustion facilities. It is 

also possible that this emission factor is an underestimation of emissions from this source 

category because it was derived from the emissions of a facility equipped with very advanced air 

pollution control technology specific for the control of dioxin emissions. These devices (DS/FF) 

are capable of greater than 95% reduction and control of dioxin-like compounds prior to 

discharge from the stack into the air. Because other facilities may not be equipped with similar 

air pollution control systems, the TEQ emissions could be higher than the estimates shown 

above. For example, Cains and Dyke (1994) reported much higher emission rates for two tire 

incinerators in the United Kingdom that were equipped with only simple grit arrestors. These 

emissions produced emission factors of 188 and 228 ng I-TEQDF/kg of tires combusted. 

EPA estimated that approximately 500 million kg of tires were combusted in 1990 

(U.S. EPA, 1992b). Of this total, 23% (115 million kg) were combusted in cement kilns, and it 

is assumed that the remaining 385 million kg were combusted in dedicated tire combustion 

facilities, industrial boilers, and pulp and paper mill combustion facilities. This activity level 

was adopted for the years 1987 and 1995 and is assigned a medium confidence rating. 

The Rubber Manufacturers Association (2002) reported that 281 million scrap tires 

weighing approximately 5.68 million metric tons were generated in the United States in 2001. 

Approximately 115 million of these scrap tires were combusted as tire-derived fuel, or roughly 

2.32 million metric tons (2.32 billion kg) of tires. Subtracting the 23% of the tires burned in 

cement kilns yields a total of 1.8 billion kg of tires estimated to have been combusted in facilities 

other than cement kilns in 2001. This figure is used to represent the activity level for tire 

combustion in 2000. This activity level is assigned a medium confidence rating. 

Annual emissions for the reference years were estimated by multiplying the activity level 

times the TEQ emission factor. The TEQ emission factor of 0.281 ng TEQDF-WHO98/kg (0.282 

ng I-TEQDF /kg) of tires combusted was used to estimate annual emissions for all years. 

Multiplying the emission factor by the activity level (385 million kg of tires) yields an estimate 

of 0.11 g TEQDF-WHO98/yr (0.11 g I-TEQDF/yr) emitted to the air in 1987 and 1995. Using the 

same emission factor multiplied by the estimated activity level of 1.8 billion kg tires combusted 

in 2000 gives an estimate of 0.51 g TEQDF-WHO98/yr (0.51 g I-TEQDF/yr). The estimated TEQ 

3-86
emissions to air from tire combustion for 1987, 1995, and 2000 are given a low confidence rating 

because of the low confidence rating of the emission factor. 

3.7. COMBUSTION OF WASTEWATER SLUDGE AT BLEACHED CHEMICAL PULP 

MILLS 

Approximately 20.5% of the wastewater sludges generated at bleached chemical pulp 

mills are dewatered and burned in bark boilers at the mills. These sludges can contain 

CDDs/CDFs and elevated levels of chloride. However, the level of heat input from sludge in the 

mixed feed to bark boilers rarely exceeds 10% (NCASI, 1995). 

NCASI (1995) provided congener-specific test results for four wood residue/sludge 

boilers tested between 1987 and 1993. Sludge comprised 6 to 10% of the solids in the feed. The 

average congener-specific emission factors derived from the stack test results obtained from 

these facilities are presented in Table 3-28. The average TEQ emission factor derived from the 

test results is 0.062 ng I-TEQDF-WHO98 (0.061 ng I-TEQDF/kg) of feed (i.e., sludge and wood 

residue), assuming nondetect values were zero. The range in facility-specific emission factors 

was wide (0.0004 to 0.118 ng I-TEQDF/kg, assuming nondetect values are zero). 

NCASI (1995) also presented stack emission test results for five other bark boilers. 

These boilers combusted only bark during the tests even though the boilers normally fire bark in 

combination with sludge and coal. These boilers are discussed in Section 4.2.2 as industrial 

facilities burning wood scrap/residues. The average TEQ emission factor for these facilities was 

0.4 ng I-TEQDF/kg of feed. The emissions test data presented in NCASI (1995), and discussed 

above, indicate that the CDD/CDF emission factors for bark/sludge combustors are similar to the 

emission factor developed in Section 4.2.2 for industrial facilities burning only wood 

residues/scrap. Based on the fact that wood residues comprise a far greater fraction of the feed to 

these bark/sludge burners than does sludge, the national TEQ emission estimates derived in 

Section 4.2.2 for industrial wood-burning facilities are assumed to include emissions from these 

bark/sludge combustion units. 

3.8. BIOGAS COMBUSTION 

Using a specially developed sampling apparatus, Schreiner et al. (1992) measured the 

CDD/CDF content of a flare combusting exhaust gases from an anaerobic sewage sludge digestor 

in Germany. The nozzle of the apparatus was moved through three cross-sections of the flame 

and cooling zone. The CDD/CDF content was 1.4 pg I-TEQDF/standard cubic meter (Nm3) at the 

bottom of the flare, 3.3. pg I-TEQDF/Nm3 at the top of the flare, and 13.1 pg I-TEQDF/Nm3 in the 

middle of the flare. Congener-specific results were not reported. Using the theoretical ratio of 

3-87
Table 3-28. CDD/CDF emission factors for combustion of bleached-kraft 

mill sludge in wood residue boilers 

Congener 

Mean emission factors 

(ng/kg feed) 

Nondetect set to zero 

Nondetect set to 

1⁄2 detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD

OCDD

 0.005

 0.005

 0.012

 0.05

 0.035

 0.301

 1.189

 0.013

 0.012

 0.022

 0.056

 0.043

 0.302

 1.192 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF

1,2,3,4,7,8,9-HpCDF

OCDF

 0.104

 0.022

 0.019

 0.069

 0.043

 0.036

 0.004

 0.274

 0.081

 0.187

 0.107

 0.029

 0.027

 0.071

 0.046

 0.041

 0.012

 0.275

 0.083

 0.188 

Total TCDD

Total PeCDD

Total HxCDD

Total HpCDD

Total OCDD

Total TCDF

Total PeCDF

Total HxCDF

Total HpCDF

Total OCDF

 0.101

 0.03

 0.599

 0.956

 1.189

 0.56

 0.469

 0.748

 1.102

 0.187

 0.108 

0.109 

0.6 

0.958

 1.192 

0.56 

0.47 

0.748 

1.102

 0.188 

Total I-TEQDF 

Total TEQDF-WHO98

 0.061

 0.062

 0.082

 0.087 

Total CDD/CDF 5.941 6.035 

Source: NCASI (1995). 

flare gas volume to digestor gas volume combusted, 78.6:1, and the average CDD/CDF content 

of the three measurements, 5.9 pg I-TEQDF/Nm3, yields an emission rate of 0.46 ng I-TEQ DF/Nm3 

of digestor gas combusted. 

During 1996, publicly owned treatment works (POTWs) in the United States treated 

approximately 122 billion L of wastewater daily (U.S. EPA, 1997b). Although reliable data are 

3-88
not readily available on the amount of sewage sludge generated by POTWs that is subjected to 

stabilization by anaerobic digestion, a reasonable approximation is 25% of the total sludge 

generated (i.e., the sludge generated from treatment of about 30 trillion L per day of wastewater). 

An estimated 196 kg of sludge solids are generated for every 1 million L of wastewater subjected 

to primary and secondary treatment (Water Pollution Control Federation, 1990). Thus, 

multiplying 30 billion L/day (25% of 122 billion L) by 196 kg/million L and 365 days/yr yields 

an annual estimate of 2 million metric tons of sludge solids that may be anaerobically digested in 

POTWs annually. 

The volume of sludge digestor gas combusted in flares annually can be estimated using 

operation parameters for a “typical” anaerobic digestor system, as described in Water Pollution 

Control Federation (1990). Multiplying the annual amount of sludge solids of 2 million metric 

tons by the following parameters and appropriate conversion factors yields an annual flared 

digestor gas volume of 467-million Nm3: 

• Fraction of total solids that are volatile solids is 75%. 

• Reduction of volatile solids during digestion is 50%. 

• Specific gas production is 0.94 m3/kg volatile solids reduced. 

• Fraction of produced gas that is flared is 66%. 

Because there are no direct measurements of CDD/CDF emissions from U.S. anaerobic 

sludge digestor flares and because of uncertainties about the activity level for biogas combustion, 

no national emissions estimate has been developed for inclusion in the national inventory. 

However, a preliminary estimate of the potential annual TEQ emissions from this source can be 

obtained by multiplying the emission factor of 0.46 ng I-TEQDF/Nm3 of digestor gas flared by the 

estimated volume of gas flared annually in the United States, 467 million Nm3

. This calculation 

yields an annual potential release in 2000 of 0.22 g. This estimate should be regarded as a 

preliminary indication of possible emissions from this source category; further testing is needed 

to confirm the true magnitude of these emissions. 

3-89
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combusto

operating in 2000 

3-90

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

I. Large MWCs (>250 tpd/unit) 

Huntsville AL Madison MB/WW 1 345 88,154.49 SD/FF/CI/SNCR Huntsville AL Madison MB/WW 2 345 88,633.51 SD/FF/CI/SNCR Long Beach SERRF CA Los Angeles MB/WW 1 360 95,572.00 SD/FF/SNCR Long Beach SERRF CA Los Angeles MB/WW 1 460 154,264.93 SD/FF/SNCR Long Beach SERRF CA Los Angeles MB/WW 2 460 164,072.92 SD/FF/SNCR Long Beach SERRF CA Los Angeles MB/WW 3 460 160,371.15 SD/FF/SNCR Stanislaus (Modesto) CA Stanislaus MB/WW 1 400 131,607.90 SD/FF/CI/SNCR Stanislaus (Modesto) CA Stanislaus MB/WW 2 400 129,798.10 SD/FF/CI/SNCR Wheelabrator Bridgeport, L.P. CT Fairfield MB/WW 1 750 238,974.40 SD/FF/CI Wheelabrator Bridgeport, L.P. CT Fairfield MB/WW 2 750 237,183.53 SD/FF/CI Wheelabrator Bridgeport, L.P. CT Fairfield MB/WW 3 750 231,472.07 SD/FF/CI Bristol RRF CT Hartford MB/WW 2 325 92,453.00 SD/FF/CI/SNCR Bristol RRF CT Hartford MB/WW 2 325 93,627.93 SD/FF/CI/SNCR Mid-Connecticut RRF CT Hartford RDF 1 675 246,570.00 SD/FF/SNCR Mid-Connecticut RRF CT Hartford RDF 2 675 251,454.04 SD/FF/SNCR Mid-Connecticut RRF CT Hartford RDF 3 675 252,415.96 SD/FF/SNCR Southeastern Connecticut RRF CT New London MB/WW 1 345 122,528.01 SD/FF/CI Southeastern Connecticut RRF CT New London MB/WW 2 345 121,053.99 SD/FF/CI Wheelabrator Lisbon Inc. CT New London MB/WW 1 250 89,556.15 SD/FF/CI/SNCR Wheelabrator Lisbon Inc. CT New London MB/WW 2 250 89,452.85 SD/FF/CI/SNCR Wheelabrator South Broward FL Ft. Lauderdale MB/WW 1 750 259,426.57 SD/FF/SNCR Wheelabrator South Broward FL Ft. Lauderdale MB/WW 2 750 244,492.13 SD/FF/SNCR Wheelabrator South Broward FL Ft. Lauderdale MB/WW 3 750 252,013.30 SD/FF/SNCR 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-91

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Wheelabrator North Broward FL Broward MB/WW 1 750 260,456.43 SD/FF/SNCR Wheelabrator North Broward FL Broward MB/WW 2 750 255,371.04 SD/FF/SNCR Wheelabrator North Broward FL Broward MB/WW 3 750 259,802.53 SD/FF/SNCR Dade Co. RRF FL Dade RDF 1 672 172,792.98 SD/FF/CI/SNCR Dade Co. RRF FL Dade RDF 2 672 171,880.39 SD/FF/CI/SNCR Dade Co. RRF FL Dade RDF 3 672 167,673.72 SD/FF/CI/SNCR Dade Co. RRF FL Dade RDF 4 672 155,352.91 SD/FF/CI/SNCR Hillsborough Co. RRF FL Hillsborough MB/WW 1 400 116,426.35 SD/FF/CI/SNCR Hillsborough Co. RRF FL Hillsborough MB/WW 2 400 120,265.74 SD/FF/CI/SNCR Hillsborough Co. RRF FL Hillsborough MB/WW 3 400 121,674.90 SD/FF/CI/SNCR McKay Bay REF FL Hillsborough MB/WW 1 250 90,232.04 SD/FF/CI/SNCR McKay Bay REF FL Hillsborough MB/WW 2 250 90,232.04 SD/FF/CI/SNCR McKay Bay REF FL Hillsborough MB/WW 3 250 90,413.44 SD/ESP/CI/SNCR McKay Bay REF FL Hillsborough MB/WW 4 250 90,050.47 SD/ESP/CI/SNCR Lake Co. RRF FL Lake MB/WW 1 264 82,586.03 SD/FF/CI/SNCR Lake Co. RRF FL Lake MB/WW 2 264 83,262.97 SD/FF/CI/SNCR Lee County Solid Waste RRF FL Lee MB/WW 1 600 197,620.93 SD/FF/CI/SNCR Lee County Solid Waste RRF FL Lee MB/WW 2 600 197,203.07 SD/FF/CI/SNCR Pasco County Solid Waste RRF FL Pasco MB/WW 1 350 98,891.97 SD/FF/CI/SNCR Pasco County Solid Waste RRF FL Pasco MB/WW 2 350 102,727.33 SD/FF/CI/SNCR Pasco County Solid Waste RRF FL Pasco MB/WW 3 350 108,759.70 SD/FF/CI/SNCR Pinellas County RRF FL Pinellas MB/WW 1 1000 299,275.82 SD/FF/CI/SNCR Pinellas County RRF FL Pinellas MB/WW 2 1000 289,763.18 SD/FF/CI/SNCR Pinellas County RRF FL Pinellas MB/WW 3 1000 301,717.00 SD/FF/CI/SNCR 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-92

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

North County RRF FL West Palm 

Beach 

RDF 1 1000 273,939.00 SD/ESP 

North County RRF FL West Palm 

Beach 

RDF 2 1000 288,988.00 SD/ESP 

Savannah RRF GA Chatham MB/WW 1 250 61,886.44 SD/FF/CI/SNCR Savannah RRF GA Chatham MB/WW 2 250 58,830.56 SD/FF/CI/SNCR Honolulu RRF HI Honolulu RDF 1 1080 258,726.76 SD/ESP Honolulu RRF HI Honolulu RDF 2 1080 256,019.24 SD/ESP Indianapolis RRF IN Marion MB/WW 1 787 209,507.33 SD/FF/CI/SNCR Indianapolis RRF IN Marion MB/WW 2 787 220,779.62 SD/FF/CI/SNCR Indianapolis RRF IN Marion MB/WW 3 787 224,095.05 SD/FF/CI/SNCR Haverhill RRF MA Essex MB/WW 1 825 281,121.60 SD/FF/CI/SNCR Haverhill RRF MA Essex MB/WW 2 825 286,516.40 SD/FF/CI/SNCR Wheelabrator North Andover MA Essex MB/WW 1 750 202,456.92 SD/FF/CI/SNCR Wheelabrator North Andover MA Essex MB/WW 2 750 180,415.08 SD/FF/CI/SNCR Wheelabrator Saugus MA Essex MB/WW 1 750 217,656.46 SD/FF/CI/SNCR Wheelabrator Saugus MA Essex MB/WW 2 750 214,057.54 SD/FF/CI/SNCR SEMASS RRF MA Plymouth RDF 1 1000 368,821.00 SD/ESP/FF/CI SEMASS RRF MA Plymouth RDF 2 1000 372,224.92 SD/ESP/FF/CI SEMASS RRF MA Plymouth RDF 3 1000 364,595.08 SD/FF/SNCR Wheelabrator Millbury MA Worcester MB/WW 1 750 230,829.25 SD/ESP/CI/SNCR Wheelabrator Millbury MA Worcester MB/WW 2 750 233,033.75 SD/ESP/CI/SNCR Wheelabrator Baltimore MD Independent 

City 

MB/WW 1 750 242,224.16 SD/ESP/CI/SNCR 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-93

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Wheelabrator Baltimore MD Independent 

City 

MB/WW 2 750 241,753.99 SD/ESP/CI/SNCR 

Wheelabrator Baltimore MD Independent 

City 

MB/WW 3 750 231,232.85 SD/ESP/CI/SNCR 

Montgomery Co. RRF MD Montgomery MB/WW 1 600 178,986.24 SD/FF/CI/SNCR Montgomery Co. RRF MD Montgomery MB/WW 2 600 170,879.15 SD/FF/CI/SNCR Montgomery Co. RRF MD Montgomery MB/WW 3 600 170,131.61 SD/FF/CI/SNCR Greater Portland Region RRF ME Cumberland MB/WW 1 250 87,323.81 SD/ESP/CI/SNCR Greater Portland Region RRF ME Cumberland MB/WW 2 250 85,676.19 SD/ESP/CI/SNCR Penobscot Energy Recovery ME Penobscot RDF 1 360 108,975.08 SD/FF Penobscot Energy Recovery ME Penobscot RDF 2 360 110,899.92 SD/FF Maine Energy Recovery ME York RDF 1 300 123,654.50 SD/FF/SNCR Maine Energy Recovery ME York RDF 2 300 123,654.50 SD/FF/SNCR Kent Co. WTE Facility MI Kent MB/WW 1 313 90,813.23 SD/FF/CI/SNCR Kent Co. WTE Facility MI Kent MB/WW 2 313 88,925.77 SD/FF/CI/SNCR Central Wayne Energy MI Wayne MB/WW 3 300 61,800.00 SD/FF/CI/SNCR Greater Detroit RRF MI Wayne RDF 1 1100 279,883.11 SD/FF Greater Detroit RRF MI Wayne RDF 2 1100 212,094.32 SD/FF Greater Detroit RRF MI Wayne RDF 3 1100 201,288.58 SD/FF Wilmarth Plant MN Blue Earth RDF 1 360 94,983.00 SD/FF/SNCR Wilmarth Plant MN Blue Earth RDF 2 360 108,339.00 SD/FF/SNCR Red Wing Plant MN Goodhue RDF 1 360 92,993.00 DSI/FF Red Wing Plant MN Goodhue RDF 2 360 89,211.00 DSI/FF Covanta Hennepin MN Hennepin MB/WW 1 600 182,903.45 SD/FF/CI/SNCR Covanta Hennepin MN Hennepin MB/WW 2 600 182,190.55 SD/FF/CI/SNCR 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-94

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Great River Energy MN Sherburne RDF 1 250 75,878.19 SD/FF Great River Energy MN Sherburne RDF 2 250 75,122.55 SD/FF Great River Energy MN Sherburne RDF 3 500 133,529.26 SD/FF New Hanover Co. WTE NC New Hanover MB/WW 3 301 126,709.00 SD/FF/CI/SNCR Wheelabrator Concord NH Merrimack MB/WW 1 250 91,065.04 SD/FF/CI/SNCR Wheelabrator Concord NH Merrimack MB/WW 2 250 92,550.96 SD/FF/CI/SNCR Camden RRF NJ Camden MB/WW 1 350 97,654.44 SD/ESP/CI/SNCR Camden RRF NJ Camden MB/WW 2 350 93,330.56 SD/ESP/CI/SNCR Camden RRF NJ Camden MB/WW 3 350 85,658.00 SD/ESP/CI/SNCR Essex Co. RRF NJ Essex MB/WW 1 900 322,862.32 SD/ESP/CI/SNCR Essex Co. RRF NJ Essex MB/WW 2 900 327,332.25 SD/ESP/CI/SNCR Essex Co. RRF NJ Essex MB/WW 3 900 335,288.43 SD/ESP/CI/SNCR Gloucester County NJ Gloucester MB/WW 1 288 90,774.74 SD/FF/CI/SNCR Gloucester County NJ Gloucester MB/WW 2 288 90,397.26 SD/FF/CI/SNCR Union Co. RRF NJ Union MB/WW 1 480 169,630.18 SD/FF/CI/SNCR Union Co. RRF NJ Union MB/WW 2 480 171,047.55 SD/FF/CI/SNCR Union Co. RRF NJ Union MB/WW 3 480 167,844.27 SD/FF/CI/SNCR Hempstead NY Nassau MB/WW 1 835 294,685.61 SD/FF/SNCR Hempstead NY Nassau MB/WW 2 835 299,122.26 SD/FF/SNCR Hempstead NY Nassau MB/WW 3 835 293,532.12 SD/FF/SNCR Niagara Falls NY Niagara MB/WW 3 1100 358,183.15 SD/FF/CI/SNCR Niagara Falls NY Niagara MB/WW 4 1100 355,338.85 SD/FF/CI/SNCR Onondaga Co. RRF NY Onondaga MB/WW 1 330 117,389.13 SD/FF/CI/SNCR Onondaga Co. RRF NY Onondaga MB/WW 2 330 99,956.76 SD/FF/CI/SNCR 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-95

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Onondaga Co. RRF NY Onondaga MB/WW 3 330 117,794.12 SD/FF/CI/SNCR Babylon RRF NY Suffolk MB/WW 1 375 112,323.03 SD/FF/CI/SNCR Babylon RRF NY Suffolk MB/WW 2 375 107,951.97 SD/FF/CI/SNCR Huntington RRF NY Suffolk MB/WW 1 250 102,526.89 SD/FF/CI/SNCR Huntington RRF NY Suffolk MB/WW 2 250 102,744.63 SD/FF/CI/SNCR Huntington RRF NY Suffolk MB/WW 3 250 110,853.48 SD/FF/CI/SNCR Adirondack RRF NY Washington MB/WW 1 250 82,458.60 SD/ESP/CI Adirondack RRF NY Washington MB/WW 2 250 79,860.40 SD/ESP/CI Wheelabrator Westchester, L.P. NY Westchester MB/WW 1 750 226,329.03 SD/FF/CI/SNCR Wheelabrator Westchester, L.P. NY Westchester MB/WW 2 750 217,185.24 SD/FF/CI/SNCR Wheelabrator Westchester, L.P. NY Westchester MB/WW 3 750 206,044.73 SD/FF/CI/SNCR Walter B. Hall RRF (Tulsa) OK Tulsa MB/WW 1 375 111,998.10 SD/FF/CI/SNCR Walter B. Hall RRF (Tulsa) OK Tulsa MB/WW 2 375 113,501.52 SD/FF/CI/SNCR Walter B. Hall RRF (Tulsa) OK Tulsa MB/WW 3 375 113,849.38 SD/FF/CI/SNCR Marion Co. WTE OR Marion MB/WW 1 275 92,091.75 SD/FF/CI/SNCR Marion Co. WTE OR Marion MB/WW 2 275 92,288.25 SD/FF/CI/SNCR Lancaster County PA Bainbridge MB/WW 1 400 125,595.02 SD/FF/CI/SNCR Lancaster County PA Bainbridge MB/WW 2 400 127,240.06 SD/FF/CI/SNCR Lancaster County PA Bainbridge MB/WW 3 400 128,049.91 SD/FF/CI/SNCR Wheelabrator Falls RRF PA Bucks MB/WW 1 750 264,839.58 SD/FF/CI/SNCR Wheelabrator Falls RRF PA Bucks MB/WW 2 750 259,729.42 SD/FF/CI/SNCR American Ref-fuel of Delaware Valley 

PA Delaware MB/RC 1 448 179,845.19 SD/FF 

American Ref-fuel of Delaware Valley 

PA Delaware MB/RC 2 448 190,795.79 SD/FF 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-96

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

American Ref-fuel of Delaware Valley 

PA Delaware MB/RC 3 448 186,088.66 SD/FF 

American Ref-fuel of Delaware Valley 

PA Delaware MB/RC 4 448 189,206.50 SD/FF 

American Ref-fuel of Delaware Valley 

PA Delaware MB/RC 5 448 176,939.66 SD/FF 

American Ref-fuel of Delaware Valley 

PA Delaware MB/RC 6 448 191,389.21 SD/FF 

Montenay Montgomery PA Montgomery MB/WW 1 600 204,290.69 SD/FF/CI/SNCR Montenay Montgomery PA Montgomery MB/WW 2 600 198,884.31 SD/FF/CI/SNCR York County PA York MB/RC 1 448 135,266.80 SD/FF/CI/SNCR York County PA York MB/RC 2 448 135,104.31 SD/FF/CI/SNCR York County PA York MB/RC 3 448 126,412.88 SD/FF/CI/SNCR Foster Wheeler Charleston RRF SC Charleston MB/WW 1 300 112,950.14 SD/ESP/CI Foster Wheeler Charleston RRF SC Charleston MB/WW 2 300 98,453.86 SD/ESP/CI Nashville Thermal Transfer Corp. TN Davidson MB/WW 1 330 67,181.87 SD/FF/CI/SNCR Nashville Thermal Transfer Corp. TN Davidson MB/WW 2 330 72,257.44 SD/FF/CI/SNCR Nashville Thermal Transfer Corp. TN Davidson MB/WW 3 390 85,979.69 SD/FF/CI/SNCR Alexandria/Arlington RRF VA Alexandria MB/WW 1 325 113,578.13 SD/FF/CI/SNCR Alexandria/Arlington RRF VA Alexandria MB/WW 2 325 109,442.21 SD/FF/CI/SNCR Alexandria/Arlington RRF VA Alexandria MB/WW 3 325 109,410.66 SD/FF/CI/SNCR I-95 Energy RRF VA Fairfax MB/WW 1 750 270,077.99 SD/FF/CI/SNCR I-95 Energy RRF VA Fairfax MB/WW 2 750 271,967.16 SD/FF/CI/SNCR I-95 Energy RRF VA Fairfax MB/WW 3 750 272,396.24 SD/FF/CI/SNCR I-95 Energy RRF VA Fairfax MB/WW 4 750 271,628.61 SD/FF/CI/SNCR 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-97

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

SPSA Waste To Energy VA Portsmouth RDF 1 500 122,153.75 SD/FF SPSA Waste To Energy VA Portsmouth RDF 2 500 122,153.75 SD/FF SPSA Waste To Energy VA Portsmouth RDF 3 500 122,153.75 SD/FF SPSA Waste To Energy VA Portsmouth RDF 4 500 122,153.75 SD/FF Spokane Regional Disposal Facility WA Spokane MB/WW 1 400 141,392.12 SD/FF/CI/SNCR 

Spokane Regional Disposal Facility WA Spokane MB/WW 2 400 143,927.88 SD/FF/CI/SNCR 

LaCrosse Co. WI LaCrosse RDF 1 288 22,727.86 DSI/FF/H2O/SNCLaCrosse Co. WI LaCrosse RDF 2 288 21,718.14 DSI/FF/H2O/SNC

Large MWC totals 2.80e+007 II. Small MWCs (<250 tpd/unit) Juneau RRF AK Juneau 

Borough 

MOD/SA 1 35 10,975.85 ESP 

Juneau RRF AK Juneau 

Borough 

MOD/SA 2 35 10,975.85 ESP 

Wallingford RRF CT New Haven MOD/EA 1 140 43,903.39 SD/FF Wallingford RRF CT New Haven MOD/EA 2 140 43,903.39 SD/FF Wallingford RRF CT New Haven MOD/EA 3 140 43,903.39 SD/FF Bay Resource Management Center FL Bay MB/RC 1 245 76,830.94 ESP Bay Resource Management Center FL Bay MB/RC 2 245 76,830.94 ESP Miami International Airport FL Dade MOD/SA 1 60 18,815.74 Unc Southernmost WTE FL Monroe MB/WW 1 75 23,519.67 ESP Southernmost WTE FL Monroe MB/WW 2 75 23,519.67 ESP Pittsfield RRF MA Berkshire MOD/EA 1 120 37,631.48 WS/ESP Pittsfield RRF MA Berkshire MOD/EA 2 120 37,631.48 WS/ESP Pittsfield RRF MA Berkshire MOD/EA 3 120 37,631.48 WS/ESP 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-98

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Springfield RRF MA Hampden MOD/EA 1 120 37,631.48 DSI/FF Springfield RRF MA Hampden MOD/EA 2 120 37,631.48 DSI/FF Springfield RRF MA Hampden MOD/EA 3 120 37,631.48 DSI/FF Harford Co. WTE Fac. MD Harford MOD/SA 1 90 28,223.61 DSI/ESP Harford Co. WTE Fac. MD Harford MOD/SA 2 90 28,223.61 DSI/ESP Harford Co. WTE Fac. MD Harford MOD/SA 3 90 28,223.61 DSI/ESP Harford Co. WTE Fac. MD Harford MOD/SA 4 90 28,223.61 DSI/ESP Mid Maine Waste Action Corp. ME Androscoggin MB 1 100 31,359.57 SD/FF Mid Maine Waste Action Corp. ME Androscoggin MB 2 100 31,359.57 SD/FF Jackson Co. RRF MI Jackson MB/WW 1 100 31,359.57 SD/FF Jackson Co. RRF MI Jackson MB/WW 2 100 31,359.57 SD/FF Central Wayne Co. MI Wayne RDF 1 249 78,085.32 ESP Central Wayne Co. MI Wayne RDF 2 249 78,085.32 ESP Pope-Douglas Waste MN Douglas MOD/EA 1 36 11,289.44 ESP Pope-Douglas Waste MN Douglas MOD/EA 2 36 11,289.44 ESP Red Wing Solid Waste Boiler Facility 

MN Goodhue MOD/EA 1 36 11,289.44 ESP 

Red Wing Solid Waste Boiler Facility 

MN Goodhue MOD/EA 2 36 11,289.44 ESP 

Olmstead WTE Facility MN Olmstead MB/WW 1 100 31,359.57 ESP Olmstead WTE Facility MN Olmstead MB/WW 2 100 31,359.57 ESP Fergus Falls MN Otter Tail MOD/SA 1 47 14,739.00 WS Fergus Falls MN Otter Tail MOD/SA 2 47 14,739.00 WS Perham Renewable RF MN Otter Tail MOD/SA 1 57 17,874.95 ESP Perham Renewable RF MN Otter Tail MOD/SA 2 57 17,874.95 ESP 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-99

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Polk Co. MN Polk MOD/SA 1 40 12,543.83 ESP Polk Co. MN Polk MOD/SA 2 40 12,543.83 ESP Pascagoula MS Jackson MOD/EA 1 75 23,519.67 ESP Pascagoula MS Jackson MOD/EA 2 75 23,519.67 ESP Livingston/Park County MT Park MOD/SA 1 36 11,289.44 Unc Livingston/Park County MT Park MOD/SA 2 36 11,289.44 Unc New Hanover Co. NC New Hanover MB/WW 1 100 31,359.57 SD/FF New Hanover Co. NC New Hanover MB/WW 2 100 31,359.57 SD/FF SES Claremont NH Sullivan MB/WW 1 100 31,359.57 DSI/FF SES Claremont NH Sullivan MB/WW 2 100 31,359.57 DSI/FF Warren Energy RF NJ Warren MB/WW 1 200 62,719.13 SD/FF/CI Warren Energy RF NJ Warren MB/WW 2 200 62,719.13 SD/FF/CI Dutchess Co. RRF NY Dutchess MB/RC 1 200 62,719.13 DSI/FF Dutchess Co. RRF NY Dutchess MB/RC 2 200 62,719.13 DSI/FF Oswego Co. WTE NY Oswego MOD/SA 3 50 15,679.78 SD/FF/CI Oswego Co. WTE NY Oswego MOD/SA 4 50 15,679.78 SD/FF/CI Oswego Co. WTE NY Oswego MOD/SA 1 50 15,679.78 SD/FF/CI Oswego Co. WTE NY Oswego MOD/SA 2 50 15,679.78 SD/FF/CI MacArthur WTE NY Suffolk MB/RC 1 243 76,203.74 DSI/FF MacArthur WTE NY Suffolk MB/RC 2 243 76,203.74 DSI/FF Miami RRF OK Ottawa MOD/SA 1 35 10,975.85 Unc Miami RRF OK Ottawa MOD/SA 2 35 10,975.85 Unc Miami RRF OK Ottawa MOD/SA 3 35 10,975.85 Unc Coos Bay Incinerator OR Coos MOD/SA 1 50 15,679.78 DSI/FF 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-100

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Coos Bay Incinerator OR Coos MOD/SA 2 50 15,679.78 DSI/FF Coos Bay Incinerator OR Coos MOD/SA 3 50 15,679.78 DSI/FF Harrisburg WTE PA Dauphin MB/WW 1 245 76,830.94 ESP Harrisburg WTE PA Dauphin MB/WW 2 245 76,830.94 ESP Sumner Co. TN Sumner MB/RC 1 100 31,359.57 ESP Sumner Co. TN Sumner MB/RC 2 100 31,359.57 ESP City of Cleburne TX Johnson MOD/SA 1 38 11,916.63 DSI/ESP City of Cleburne TX Johnson MOD/SA 2 38 11,916.63 DSI/ESP City of Cleburne TX Johnson MOD/SA 3 38 11,916.63 DSI/ESP Panola Co. WTE TX Panola MOD/SA 1 40 12,543.83 WS Center RRF TX Shelby MOD/SA 1 40 12,543.83 WS Davis/Wasatch UT Davis MB/REF 1 200 62,719.13 DSI/ESP Davis/Wasatch UT Davis MB/REF 2 200 62,719.13 DSI/ESP Arlington -Pen tagon VA Arlington MOD/SA 1 50 15,679.78 FF Arlington -Pen tagon VA Arlington MOD/SA 2 50 15,679.78 FF Galax City SW VA Grayson MB/WW/RC 1 56 17,561.36 FF NASA Refuse-fired Steam Generator 

VA Hampton City MB/WW 1 100 31,359.57 ESP 

NASA Refuse-fired Steam Generator 

VA Hampton City MB/WW 2 100 31,359.57 ESP 

Harrisonburg VA Rockingham MB/WW 1 50 15,679.78 Unc Harrisonburg VA Rockingham MB/WW 2 50 15,679.78 Unc Tacoma WA Pierce FB/RDF 1 150 47,039.35 DSI/FF Tacoma WA Pierce FB/RDF 2 150 47,039.35 DSI/FF Barron Co. WI Barron MOD/SA 1 50 15,679.78 ESP 
Table 3-5. National CDD/CDF TEQ emissions (g/yr) for large and small municipal waste combust

operating in 2000 (continued) 

3-101

Facility name State County/city MWC type 

Unit 

no. 

MSW 

capacity 

(tons/day) 

MSW 

combusted 

(tons/yr) APCD 

Barron Co. WI Barron MOD/SA 2 50 15,679.78 ESP Small MWC totals 2.60e+006 All MWCs operating in 2000 (totals) 

3.06e+007 

RRF = Resource recovery facility (steam-generating facility)MSW = Municipal solid wasteWTE = Waste-to-energy (facility)APCD (air pollution control device): MWC type:

CI = Carbon injection FB/RDF = Fluidized-bed refuse-derived fuel DS = Dry scrubber MB/REF = Mass burn refractory DSI = Dry sorbent injection MB/WW = Mass burn waterwall ESP = Electrostatic precipitator MB = Mass burn FF = Fabric filter MB/RC = Mass burn rotary kiln SD = Spray dryer MOD/EA = Modular excess air SNRL = Selective noncatalytic reduction MOD/SA = Modular starved air Unc = Uncontrolled RDF = Refuse-derived fuel WS = Wet scrubber 
4. COMBUSTION SOURCES OF CDDs/CDFs: POWER/ENERGY GENERATION 

4.1. MOTOR VEHICLE FUEL COMBUSTION 

Ballschmiter et al. (1986) reported detecting CDDs/CDFs in used motor oil, thus 

providing some of the first evidence that CDDs/CDFs might be emitted by the combustion 

processes in gasoline- and diesel-fueled engines. Incomplete combustion and the presence of a 

chlorine source in the form of additives such as dichloroethane or pentachlorophenate in the oil 

or the fuel were speculated to lead to the formation of CDDs/CDFs. The congener patterns found 

in the used oil samples were characterized by Ballschmiter et al. as being similar to the patterns 

found in fly ash and stack emissions from municipal waste incinerators. 

Since 1986, several studies have been conducted to measure or estimate CDD/CDF 

concentrations in emissions from vehicles. Although there is no standard approved protocol for 

measuring CDDs/CDFs in vehicle exhaust, some researchers have developed and implemented 

several approaches for collecting and analyzing tailpipe emissions. Other researchers have 

estimated vehicle exhaust emissions of CDDs/CDFs indirectly from studies of tunnel air. The 

results of these two types of studies are summarized in Sections 4.1.1 and 4.1.2. 

Estimates of national annual CDD/CDF TEQ emissions from on-road and off-road motor 

vehicles fueled with leaded gasoline, unleaded gasoline, and diesel fuel based on the results of 

those studies are presented in Section 4.1.3. It should be noted, however, that relatively few tests 

on emissions from diesel- and unleaded gasoline-fueled vehicles are available, considering the 

variety and number of such vehicles currently in operation and the range of operational, 

technical, and environmental conditions in which they are operated. As a result, the emission 

factors developed in this report for on-road and off-road motor vehicles are quite uncertain. 

4.1.1. Tailpipe Emission Studies 

Marklund et al. (1987) provided the first direct evidence of the presence of CDDs/CDFs 

in car exhaust by measuring emissions from tailpipes of Swedish cars. Approximately 20 to 220 

pg I-TEQDF/km driven from tetra- and penta-CDDs/CDFs were reported for four cars running on 

leaded gasoline. For this study, an unleaded gasoline was used, with tetramethyl lead (0.15 g/L 

[0.57 g/gal]) and 1,2-dichloroethane (0.1 g/L as a scavenger) added. The fuel used may not have 

accurately represented commercial fuels at that time, which typically contained a mixture of 

chlorinated and brominated scavengers (Marklund et al., 1990). Also, the lead content of the fuel 

used (0.15 g lead/L) was the normal content for Swedish fuels at the time, but it was higher than 

that of leaded gasoline in the United States during the late 1980s (lowered to 0.1 g lead/gal 

[0.026 g lead/L] effective January 1, 1986). The authors reported a striking similarity between 

the TCDF and PeCDF congener profiles in the car exhausts and those found in emissions from 

municipal waste incinerators. For two cars running on unleaded gasoline, CDD/CDF emissions 

4-1
 

were below the detection limit (DL), which corresponded to approximately 13 pg I-TEQDF/km 

driven. 

Table 4-1 presents a summary of the results of Marklund et al. (1987) and subsequent 

studies, which are discussed below. Tables 4-2 and 4-3 present the results of tailpipe emission 

studies reported for diesel-fueled cars and trucks, respectively. The results of studies of leaded 

gasoline-fueled cars are shown in Table 4-4 and those for unleaded gasoline-fueled cars inTables 

4-5 and 4-6. Figures 4-1, 4-2, and 4-3 present congener and congener group profiles for 

emissions from diesel-fueled vehicles, leaded gasoline-fueled vehicles, and unleaded gasolinefueled vehicles, respectively. 

Virtually no testing of vehicle emissions for CDDs/CDFs in the United States has been 

reported. In 1987, the California Air Resources Board (CARB) produced a draft report on the 

testing of exhausts from four gasoline-fueled cars and three diesel-fueled vehicles (one truck, one 

bus, and one car) (CARB, 1987). However, CARB indicated to EPA that the draft report should 

not be cited or quoted to support general conclusions about CDDs/CDFs in motor vehicle 

exhausts because of the small sample size of the study and because the use of low-resolution 

rather than high-resolution mass spectrometry in the study resulted in high DLs and inadequate 

selectivity in the presence of interferences (letter dated July 1993, from G. Lew, California Air 

Resources Board, to J. Schaum, U.S. EPA). 

CARB stated 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. This test was conducted under steady-state conditions (50 km/hr) for 6 hr with an 

engine with a fuel economy of 5.5 km/L. The TEQ emission factor of this one sample was 

equivalent to 7,190 pg TEQDF-WHO98/L (7,290 pg I-TEQDF/L) fuel burned. An assumed fuel 

economy of 5.5 km/L yields an emission factor of 1,307 pg TEQDF-WHO98/km (1,325 pg ITEQDF/km). Assuming that nondetect values were zero, the TEQ emission factors were 3,280 pg 

TEQDF-WHO98/L (3,720 pg I-TEQDF/L) fuel burned and 596 pg TEQDF-WHO98/km (676 pg ITEQDF/km) driven (letter dated January 11, 1996, from G. Lew, California Air Resources Board, 

to G. Schweer, Versar, Inc.). 

Haglund et al. (1988) sampled exhaust gases from three vehicles (one car fueled with 

leaded gasoline and one with unleaded gasoline and a heavy-duty diesel truck) for the presence 

of brominated dibenzo-p-dioxins (BDDs) and ethylene dibromide dibenzofurans (BDFs). The 

authors concluded that the ethylene dibromide scavenger added to the tested gasoline probably 

acted as a halogen source. Tetra-BDF emissions were measured as 23,000 pg/km in the car with 

leaded gasoline and 240 pg/km in the car with unleaded gasoline. Tetra- and penta-emissions 

were measured as 3,200 and 980 pg/km, respectively, in the car with leaded gasoline. All 

BDDs/BDFs were below DLs in the diesel truck emissions. 

4-2
 
Table 4-1. Description and results of vehicle emission testing studies for CDDs and CDFs 

4-3
 


Study Country Fuel type Scavengera 

Catalyst 

equipped 

Number 

of test 

vehicles 

TEQ emission factorb 

(pg/km driven) 

Drivi

CARB (1987); letter United States Diesel (truck) No NR 1 676–1325c [597–1307] 6-hr dynam

dated January 11, 1996, from G. Lew, California Air Resources Board, to G. Schweer, Versar, Inc. Marklund et al. (1987) Sweden Unleaded No Yes 2 Not detected (<13) A10 (2 cyc

Leaded Yes No 4 Approx. 20–220 A10 (2 cyc

Bingham et al. (1989) New Zealand Unleaded No NR 1 Not detected (<20) A10 (3 or 4 

Leaded Yes NR 4 1–39 A10 (3 or 4 

Marklund et al. (1990) Sweden Unleaded 

Leaded Unleaded Leaded Diesel (truck) 

No 

Yes 

No 

Yes 

No 

No No 

Yes 

No 

NR 

2 2 1 2 1 

0.36–0.39 2.4–6.3 0.36 1.1–2.6d 

Not detected (<18)c 

FTP-73 tesFTP-73 tesFTP-73 tesFTP-73 tesU.S. federa

Hagenmaier et al. Germany Unleaded No No 1 5.1c [6.0] Comparabl

(1990) Unleaded No Yes 1 0.7c [0.8] Comparabl

Leaded Yes No 1 108c [129] Comparabl

Diesel (car) No NR 1 2.1c [2.5] Comparabl

Oehme et al. (1991) (tunnel study) 

Norway – – – e 520f 

38f 

Cars movin

60 km/hr 

Cars movin

Avg = 280 9500f 

70 km/hr 

Car averagTrucks mo

720f 

60 km/hr 

Trucks mo

Avg = 5,100 

70 km/hr 

Truck aver
Table 4-1. Description and results of vehicle emission testing studies for CDDs and CDFs (continue

4-4
 


Study Country Fuel type Scavengera 

Catalyst

equipped 

Number 

of test 

vehicles 

TEQ emission factorb 

(pg/km driven) 

Drivi

Schwind et al. (1991) Hutzinger et al. (1992) 

Germany Leaded 

Unleaded 

Yes 

No 

No No 

1 1 

5.2–118c [7.2–142] 

9.6–17.7c [10.2–18.1] 

Various tesVarious tes

Unleaded Diesel (car) 

No No 

Yes 

No 

1 1 

1–2.6c [1–2.8] 

1–13c [1.2–14] 

Various tesVarious tes

Diesel (truck) No No 1 13–15c [14–15] Various tes

Gertler et al. (1996, 1998) (tunnel study) 

United States Diesel (truck) – – g Mean = 172 Mean of se

Gullett and Ryan (1997) 

United States Diesel (truck) No – 1 Mean = 29 Mean of fiv

aExcept in Marklund et al. (1987), dichloroethane and dibromoethane were used as scavengers. bValues are in units of I-TEQDF; values in brackets are in units of TEQDF-WHO98. cResults reported were in units of pg TEQ/L of fuel. For purposes of this table, the fuel economy factor used by Marklund et al. (1990), 

 miles/gal), was used to convert the emission rates into units of pg TEQ/km driven for the cars. For the diesel-fueled truck, the fuel econ CARB (1987) for a 1984 heavy-duty diesel truck, 5.5 km/L (13.2 miles/gal), was used. 

dTable reflects the range of summary results reported in Marklund et al. (1990); however, the congener-specific results for the single run 

 emission rate of about 7.3 pg I-TEQDF/km. 

eTests were conducted over portions of 4 days, with traffic rates of 8,000–14,000 vehicles/day. Heavy-duty vehicles (defined as vehicle

 ranged from 4 to 15% of total. 

fEmission factors are reported in units of pg Nordic TEQ/km driven; the values in units of I-TEQDF/km are expected to be about 3 to 6% gTests were conducted over 5 days, with heavy-duty vehicle rates of 1,800–8,700 vehicles per 12-hr sampling event. Heavy-duty vehicl

 to 28% of all vehicles. NR = Not reported – = No data 
Table 4-2. CDD/CDF congener emission factors (pg/L) for diesel-fueled automobiles 

4-5
 


Congener/congener group 

Automobile tailpipe emission study results Mea

63 km/hra 

Idling 

(test no. 25)b 

57 km/hr 

(test no. 24)b 

57 km/hr 

(full load) 

(test no. 28)b 

Assumi

nondetec

to zero 

2,3,7,8-TCDD 1,2,3,7,8-PeCDD 1,2,3,4,7,8-HxCDD 1,2,3,6,7,8-HxCDD 1,2,3,7,8,9-HxCDD 1,2,3,4,6,7,8-HpCDD OCDD 

7.9 9 

ND (5.1) ND (5.1) ND (5.1) 

44.1 

440 

13.1 

6.3 

21.4 36 28 

107 635 

2.4 4.1 1 1.4 2 

22.9 

525 

22 23 

7.8 

21 10 

166 560 

11.4 10.6 

7.6 

14.6 10 85 

540 

2,3,7,8-TCDF 1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

20.5 

ND (5.1) 

7.1 6.5 6.7 

ND (5.1) ND (5.1) 

40.7 

8.5 

94.4 

79 

171 

58.7 

121 

75 17.1 52 

159 

11.9 

214 

18.1 

1.8 3.4 4.1 3 0.8 

ND (0.4) 

18.9 

7.1 

101 

236 111 

85 68 55 

4.7 

31 

214 

7.8 

305 

88.4 71 38.6 49.9 34.9 

5.7 

20.8 

108.2 

8.8 

178.6 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF (nondetect set to 0) 

Total I-TEQDF (nondetect set to 1⁄2 DL) 

Total TEQDF-WHO98 (nondetect set to 0) 

Total TEQDF-WHO98 (nondetect set to 1⁄2 DL) 

501 

184.4 

20.8 22.2c 

24.8 26.2 

846.8 

958.7 

100.7 

100.7 

103.1 103.1 

558.8 

158.2 

10.4 

10.4 

11.9 

1.9 

809.8 

1117.5 

129.6 

129.6 

140.4 140.4 

679.1 

604.9 

65.5 

70.1 
Table 4-2. CDD/CDF congener emission factors (pg/L) for diesel-fueled automobiles (continued) 

4-6
 


Congener/congener group 

Automobile tailpipe emission study results Mea

63 km/hra 

Idling

(test no. 25)b 

57 km/hr 

(test no. 24)b 

57 km/hr 

(full load) 

(test no. 28)b 

Assumi

nondetec

to zero 

Total TCDD Total PeCDD Total HxCDD Total HpCDD Total OCDD Total TCDF Total PeCDF Total HxCDF Total HpCDF Total OCDF 

37.4 19.7 23.6 88.5 

440.5 

76.7 39.3 25.6 80.6 94.4 

317 214 256 187 635 436 821 556 321 214 

31 22 20 77 

525 

58 36 26 72 

101 

394 228 164 356 560 

3,093 1,205 

472 241 305 

195 121 116 177 540 916 525 270 179 179 

Total CDD/CDF (nondetect set to 0) Total CDD/CDF (nondetect set to 1⁄2 DL) 

926.3 926.3 

3,957 3,957 

968 968 

7,018 7,018 

3,218 

aSource: Hagenmaier et al. (1990). bSource: Schwind et al. (1991); Hutzinger et al. (1992). cAn I-TEQDF emission factor of 23.6 pg/L is reported in Hagenmaier et al. (1990); however, an I-TEQDF emission factor of 22.2 pg/L is 

 reported congener levels. ND = Not detected (value in parenthesis is the detection limit) 
Table 4-3. CDD/CDF congener emission factors (pg/L) for diesel-fueled trucks 

4-7
 


Congener/congener group 

Truck tailpipe study results Mean em

50 km/hr

(test no. 40)a 

90 km/hr 

(full load)

(test no. 42)a 50 km/hr b 

Assuming 

nondetect set to 

zero 

2,3,7,8-TCDD 1,2,3,7,8-PeCDD 1,2,3,4,7,8-HxCDD 1,2,3,6,7,8-HxCDD 1,2,3,7,8,9-HxCDD 1,2,3,4,6,7,8-HpCDD OCDD 

25 

5 

14 28 14 

119 

1355 

16 18 

5.7 6 6 

74 

353 

ND (560) 

ND (1,340) ND (2,160) ND (1,770) ND (2,640) 

116,000 344,400 

13.7 

7.7 6.6 

11.3 

6.7 

38,731 

115,369 

2,3,7,8-TCDF 1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

87 45 18 56 84 

4.7 

63 

375 

40 

397 

53 34 51 29 31 

5.1 

23 71 

5.4 

104 

ND (605) 

ND (4,750) ND (5,190) ND (8,210) ND (6,480) 

13,400 

ND (7,780) 

73,460 

ND (11,700) 

140,400 

46.7 26.3 23 28.3 38.3 

4,469 

28.7 

24,636 

15.1 

46,981 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF Total I-TEQDF (nondetect set to 0) 

Total I-TEQDF (nondetect set to 1⁄2 DL) Total TEQDF-WHO98 (nondetect set to 0) 

Total TEQDF-WHO98 (nondetect set to 1⁄2 DL) 

1,560 1,169.7 

81 81 

82 82 

478.7 406.5 

70 70 

79 79 

460,400 227,260 

3,720 7,290 

3,280 7,190 

154,146 

76,292.4 

1,290 

1,150 
Table 4-3. CDD/CDF congener emission factors (pg/L) for diesel-fueled trucks (continued) 

4-8
 


Congener/congener group 

Truck tailpipe study results Mean em

50 km/hr 

(test no. 40)a 

90 km/hr

(full load) 

(test no. 42)a 50 km/hr b 

Assuming

nondetect set to 

zero 

Total TCDD Total PeCDD Total HxCDD Total HpCDD Total OCDD Total TCDF Total PeCDF Total HxCDF Total HpCDF Total OCDF 

200 

32 

130 200 

1,355 

763 230 524 509 397 

208 117 

67 

155 353 694 736 268 

76 

104 

ND (3,760) ND (3,020) 

ND (45,300) 

203,300 344,000 

25,000 47,900 

169,200 150,700 140,300 

136 

49.7 65.7 

67,892 

115,252 

8,831 

16,294 56,670 50,414 46,932 

Total CDD/CDF (nondetect set to 0) Total CDD/CDF (nondetect set to 1⁄2 DL) 

4,340 4,340 

2,778 2,778 

1,080,400 1,106,440 

362,536.4 

aSource: Schwind et al. (1991); Hutzinger et al. (1992). bSource: Letter dated July 1993, from G. Lew, California Air Resources Board, to J. Schaum, U.S. EPA; letter dated January 11, 1996, 

Air Resources Board, to G. Schweer, Versar, Inc.. ND = Not detected (value in parenthesis is the detection limit) 
Table 4-4. CDD/CDF congener emission factors (pg/L) for leaded gasoline-fueled automobiles 

4-9
 


Congener/congener group 

Automotive tailpipe emission study results 

FTP cyclea 63 km/hrb 

Idling 

(test no.

12)c 

Full load 

(test no.

13)c 

64 km/hr 

(test no.

14)c 

Rated 

power 

(test no.

15)c 

FTP cy

(test n

22)c 

2,3,7,8-TCDD ND (14.4) 128 NR 60 141 NR 5 1,2,3,7,8-PeCDD ND (36) 425 43 106 468 40 73 1,2,3,4,7,8-HxCDD ND (54) 188 17 15 206 16 41 1,2,3,6,7,8-HxCDD ND (54) 207 32 35 228 30 62 1,2,3,7,8,9-HxCDD ND (54) 188 NR NR 206 NR 35 1,2,3,4,6,7,8-HpCDD ND (54) 503 119 136 554 111 518 OCDD ND (90) 498 380 513 549 1,166 1,581 

2,3,7,8-TCDF 1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

432 

21.6 43.2 

ND (54) 

ND (54) ND (54) ND (54) ND (54) ND (54) ND (90) 

1,542 1,081 

447 856 

856 

ND (76) 

273 

4,051 

ND (76) 

230 

NR 

49 26 33 

22 

NR NR 

170 

NR 

1115 

678 367 156 

70 

60 

NR 

25 

NR NR NR 

1,697 1,190 

492 942 

942 

NR 

301 

4,460 

NR 

253 

78 45 24 31 

20 

NR NR 

158 

NR 

447 

214 218 225 381 

375 

85 

1,033 2,301 

109 

1,128 

Total 2,3,7,8-CDD ND 2,137 $591 $865 2,352 $1,363 2,315 

Total 2,3,7,8-CDF 496.8 9,336 $1,415 $1,356 $10,277 $803 6,069 

Total I-TEQDF ( nondetect set to 0) 65.9 1,075 $52 $300 $1,184 $56 419 

Total I-TEQDF (nondetect set to 

1⁄2 DL) 102 1,080 $52 $300 $1,184 $56 419 

Total TEQDF-WHO98 (nondetect 

set to 0) 65.9 1,287 $72 $352 $1,417 $75 454 

Total TEQDF-WHO98 (nondetect 

set to 1⁄2 DL) 111 1,291 $72 $352 $1,417 $75 454 
Table 4-4. CDD/CDF congener emission factors (pg/L) for leaded gasoline-fueled automobiles (con

4-10
 


Congener/congener group 

Automotive tailpipe emission study results 

FTP cyclea 63 km/hrb 

Idling

(test no. 

12)c 

Full load 

(test no. 

13)c 

64 km/hr 

(test no. 

14)c 

Rated 

power 

(test no. 

15)c 

FTP cy

(test n

22)c 

Total TCDD 5,220 4,555 517 8,134 5,012 4,558 921 Total PeCDD ND (360) 3,338 658 2,161 3,675 6,389 359 Total HxCDD ND (540) 1,868 354 623 2,056 1,973 996 Total HpCDD ND (90) 1,164 194 297 1,281 2,374 988 Total OCDD ND (90) 498 380 513 549 1,166 1,581 Total TCDF 15,300 50,743 2,167 20,513 55,857 29,353 4,290 Total PeCDF 2,430 11,591 452 3,608 12,757 10,580 3,165 Total HxCDF ND (540) 6,308 192 477 6,947 12,553 3,132 Total HpCDF ND (270) 5,642 170 NR 6,210 4,767 2,920 Total OCDF ND (90) 230 1,115 NR 253 447 1,128 

Total CDD/CDF (nondetect set 

to 0) 

Total CDD/CDF (nondetect set to 

1⁄2 DL) 

22,950 

23,940 

85,937 

85,937 

6,199 

6,199 

$36,326 $36,326 

94,597 94,597 

74,160 74,160 

19,480 19,480 

aSource: Marklund et al. (1990); values in the table were calculated from the reported units of pg/km to pg/L using a fuel economy of 9 

 reported in Marklund et al. (1990). 

bSource: Hagenmaier et al. (1990). cSource: Schwind et al. (1991); Hutzinger et al. (1992). 

ND = Not detected (value in parenthesis is the reported detection limit) NR = Not reported 
Table 4-5. CDD/CDF congener emission factors (pg/L) for unleaded gasoline-fueled automobiles w

converters 

4-11
 


Congener/congener group 

Automotive tailpipe emission study results 

FTP cyclea 63 km/hrb 

FTP cycle

(test no. 

21)c 

64 km/hr

(test no. 

17)c 

64 km/hr

(test no. 

20)c 

64 km/hr

(test no. 

31/2)c 

A

n

s

2,3,7,8-TCDD 1,2,3,7,8-PeCDD 1,2,3,4,7,8-HxCDD 1,2,3,6,7,8-HxCDD 1,2,3,7,8,9-HxCDD 1,2,3,4,6,7,8-HpCDD OCDD 

ND (5) ND (3) 

ND (40) ND (40) ND (40) ND (40) ND (50) 

2.6 

19.1 16.6 17.1 17.6 40.4 

176 

24 14 24 84 15 

192 868 

44 31 26 28 29 66 

280 

7 

11 25 42 23 

121 685 

8.9 

14.1 16.3 60.1 17.1 

197.8 

2,634 

2,3,7,8-TCDF 1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

64 

ND (7) ND (7) 

ND (40) ND (40) ND (40) ND (40) 

ND (40) ND (40) ND (70) 

44 44.5 20.7 41.9 21.2 37.8 54.3 

27.9 16.6 

119 

70 40 30 68 62 47 55 

278 

ND (1) 

374 

71 72 34 68 34 61 

88 45 27 

194 

77 69 

184 

88 35 

ND (1) 

42 22 24 

288 

295.2 161.8 135.2 129.1 113.2 

36.9 

82.1 

418 

54.5 

991 

Total 2,3,7,8-CDD Total 2,3,7,8-CDF Total I-TEQDF (nondetect set to 0) 

Total I-TEQDF (nondetect set to 

1⁄2 DL) 

Total TEQDF-WHO98 (nondetect set 

to 0) 

Total TEQDF-WHO98 (nondetect set 

to 1⁄2 DL) 

ND 

64 

6.4 

26.2 

6.4 

26.9 

289.4 427.9 

50.9 50.9 

60.2 60.2 

1,221 1,024 

96.4 96.4 

102 102 

504 694 

122 122 

138 138 

914 829 

144 144 

148 148 

2,948.3 2,417 

177 177 

181 181 
Table 4-5. CDD/CDF congener emission factors (pg/L) for unleaded gasoline-fueled automobiles w

converters (continued) 

4-12
 


Automotive tailpipe emission study results 

FTP cycle 64 km/hr 64 km/hr 64 km/hr A

(test no. (test no. (test no. (test no. n

Congener/congener group FTP cyclea 63 km/hrb 21)c 17)c 20)c 31/2)c s

Total TCDD 13 435 429 706 500 304 Total PeCDD ND (3) 481 837 784 542 170 Total HxCDD ND (40) 305 484 496 563 114 Total HpCDD ND (10) 93 392 147 225 301 Total OCDD ND (5) 176 868 280 685 2,634 Total TCDF 170 569 718 923 478 6,379 Total PeCDF ND (7) 931 531 1,513 437 1,969 Total HxCDF ND (40) 378 165 615 258 1,226 Total HpCDF ND (20) 476 278 773 445 1,088 Total OCDF ND (7) 119 374 194 288 991 

Total CDD/CDF (nondetect set to 0) 183 3,963 5,076 6,431 4,421 15,176 Total CDD/CDF (nondetect set to 249 3,963 5,076 6,431 4,421 15,176 

1⁄2 DL 

aSource: Marklund et al. (1990); the pg/L values in the table were calculated from the reported units of pg/km assuming a fuel economy 

 unleaded gas. 

bSource: Hagenmaier et al. (1990). cSource: Schwind et al. (1991); Hutzinger et al. (1992). 

ND = Not detected (value in parenthesis is the reported detection limit) 
Table 4-6. CDD/CDF congener emission factors (pg/L) for unleaded gasoline-fueled automobiles w

converters 

4-13
 


Congener/congener group 

Automotive tailpipe emission study test results M

63 km/hra 

64 km/hr 

(test no. 29)b 

64 km/hr 

(test no. 30)b 

64 km/hr 

(test no. 18)b 

Ass

non

set 

2,3,7,8-TCDD 1,2,3,7,8-PeCDD 1,2,3,4,7,8-HxCDD 1,2,3,6,7,8-HxCDD 1,2,3,7,8,9-HxCDD 1,2,3,4,6,7,8-HpCDD OCDD 

1.6 1.6 2.4 3.5 3.1 

15.3 

170 

3 2.6 5.3 6 6 

27.8 

275 

ND (7.9) ND (7.9) ND (7.9) 

6.4 

ND (7.9) 

78.1 

427 

14 

4 1 2 2 

14 

197 

3

26

2,3,7,8-TCDF 1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

4.3 3.3 2.4 4.8 6.3 0.2 4.6 

16.3 

ND (0.2) 

27.9 

10.6 

8.7 7.2 

10.6 

9.1 

ND (3.8) 

18.1 54.3 

ND (3.8) 

38 

12.7 

5.1 6.2 4.5 3.9 2.1 8.2 

154.2 

7.9 

106 

35 13 

6 5 7 5 

ND (1) 

51 

1 

140 

6

7

Total 2,3,7,8-CDD Total 2,3,7,8-CDF Total I-TEQDF (nondetect set to 0) 

Total I-TEQDF (nondetect set to 1⁄2 DL) Total TEQDF-WHO98 (nondetect set to 0) 

Total TEQDF-WHO98 (nondetect set to 1⁄2 DL) 

197.5 

70.1 

7.2 7.2 

7.8 7.8 

325.7 156.6 

16 16.2 

17.1 17.3 

511.5 310.8 

10.1 16.8 

9.6 

18.3 

234 263 

26.3 26.4 

28 28.1 

320
Table 4-6. CDD/CDF congener emission factors (pg/L) for unleaded gasoline-fueled automobiles w

converters (continued) 

4-14
 


Congener/congener group 

Automotive tailpipe emission study test results M

63 km/hra 

64 km/hr 

(test no. 29)b 

64 km/hr 

(test no. 30)b 

64 km/hr 

(test no. 18)b 

Ass

non

set 

Total TCDD Total PeCDD Total HxCDD Total HpCDD Total OCDD Total TCDF Total PeCDF Total HxCDF Total HpCDF Total OCDF 

28.6 25.5 26.3 38.7 

170 

52.6 53.4 33.3 27.1 27.9 

51 51 56 50 

275 152 122 

71 62 38 

13 

ND (15) 

36 

163 427 

79 29 60 

174 106 

82 

101 

50 25 

197 332 

84 39 83 

140 

4446

2615

7587

Total CDD/CDF (nondetect set to 0) Total CDD/CDF (nondetect set to 1⁄2 DL) 

483.4 483.4 

928 928 

1,087 1,087 

1,133 1,133 

90

aSource: Hagenmaier et al. (1990). bSource: Schwind et al. (1991); Hutzinger et al. (1992). 

ND = Not detected (value in parenthesis is the reported detection limit) 
Ratio (congener group emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 0.25 0.3 0.35 

TCDD 

PeCDD 

HxCDD 

HpCDD 

OCDD 

TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF 

Cars Trucks 

Figure 4-1. Congener and congener group profiles for air emissions from 

diesel-fueled vehicles (based on profiles calculated from emission factors 

[nondetects equal one-half the detection limit] from Tables 4-2 and 4-3). 

4-15
 

Ratio (congener group emission factor/total CDD/CDF emission factor) 

0 0.1 0.2 0.3 0.4 0.5 

TCDD
 


PeCDD
 


HxCDD
 


HpCDD
 


OCDD
 


TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF 

Figure 4-2. Congener and congener group profiles for air emissions from 

leaded gas-fueled vehicles (based on profiles calculated from emission factors 

[nondetects equal one-half the detection limit] from Table 4-4). 

4-16
 


0.6 
Ratio (congener group emission factor/total CDD/CDF emission 

0 0.05 0.1 0.15 0.2 0.25 0.3 0.35 

TCDD
 


PeCDD
 


HxCDD
 


HpCDD
 


OCDD
 


TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF 

Figure 4-3. Congener and congener group profiles for air emissions from 

unleaded gas-fueled vehicles (catalytic converter equipped vehicles; based on 

data from Table 4-6). 

4-17
 

Bingham et al. (1989) analyzed the exhausts of four cars using leaded gasoline (uniform 

lead and organics content of 0.45 g/L tetramethyl lead, 0.22 g/L dichloroethane, and 0.2 g/L 

dibromoethane) and the exhaust of one car using unleaded gasoline. Analytical results and DLs 

were reported for only 5 of the 17 toxic CDD/CDF congeners. TEQ emission rates for the cars 

using leaded fuel, based on detected congeners only, ranged from 1 to 39 pg I-TEQDF/km. 

CDDs/CDFs were not detected in the exhaust from the vehicle using unleaded fuel; the total TEQ 

emission rate for this car, based on one-half the DLs for the five reported congeners, was 20 pg ITEQDF/km. 

Marklund et al. (1990) tested Swedish cars fueled with commercial fuels, measuring 

CDD/CDF emissions before and after the muffler. Both new and old vehicles were tested. The 

tests were done on three cars using unleaded gasoline and two cars using leaded gasoline (0.15 g 

Pb/L with 1,2-dichloroethane and ethylene dibromide scavengers). CDDs/CDFs were not 

detected in the fuels at a DL of 2 pg I-TEQDF/L but were detected at a level of 1,200 pg I-TEQDF/L 

in the new semisynthetic engine lube oil used in the engines. The test driving cycle used (31.7 

km/hr as a mean speed, 91.2 km/hr as a maximum speed, and 17.9% of time spent idling) yielded 

fuel economies ranging from approximately 9 to 10 km/L (22 to 24 miles/gal) in the various cars. 

The reported ranges of emission factors were 

•	 Leaded gas, measured before muffler: 2.4 to 6.3 pg I-TEQDF/km (21 to 60 pg ITEQDF/L fuel consumed) 

•	 Leaded gas, measured in tailpipe: 1.1 to 2.6 pg I-TEQDF/km (10 to 23 pg I-TEQDF/L) 

•	 Unleaded gas, catalyst-equipped, measured in tailpipe: 0.36 pg I-TEQDF/km (3.5 pg 

I-TEQDF/L) 

•	 Unleaded gas, measured before muffler: 0.36 to 0.39 pg I-TEQDF/km (3.5 pg ITEQDF/L) 

The TEQ levels in exhaust gases from older cars using leaded gasoline were up to six 

times greater when measured before the muffler than when measured after the muffler. No 

muffler-related difference was observed in new cars running on leaded gasoline or in old or new 

cars running on unleaded gasoline. 

Marklund et al. (1990) also analyzed the emissions of a heavy-duty, diesel-fueled truck for 

CDDs/CDFs. None were detected; however, the authors pointed out that the test fuel was a 

reference fuel and may not have been representative of commercial diesel fuel. Also, due to 

analytical problems, a much higher DL (about 100 pg I-TEQDF/L) was realized in this diesel fuel 

test than in the gasoline tests conducted (5 pg I-TEQDF/L). Further uncertainty was introduced 

because the diesel emission samples were collected only before the muffler. 

4-18
 
Hagenmaier et al. (1990) ran a set of tests using conditions comparable to the FTP-73 test 

cycle on gasoline- and diesel-fueled engines for light-duty vehicles in Germany. The following 

average TEQ emission rates per liter of fuel consumed were reported: 

•	 Leaded fuel: 1,287 pg TEQDF-WHO98/L (1,080 pg I-TEQDF/L) 

•	 Unleaded fuel (catalyst equipped): 7.9 pg TEQDF-WHO98/L (7.2 pg I-TEQDF/L) 

•	 Unleaded fuel (not catalyst equipped): 60.2 pg TEQDF-WHO98/L (50.9 pg I-TEQDF/L) 

•	 Diesel fuel: 24.8 pg TEQDF-WHO98/L (20.8 pg I-TEQDF/L) 

Schwind et al. (1991) published the major findings of a German study of emissions of 

halogenated dibenzo-p-dioxins and dibenzofurans from internal combustion engines running on 

commercial fuels. The full report was published in 1992 (Hutzinger et al., 1992). The study was 

conducted by the universities of Stuttgart, Tübingen, and Bayreuth for the Federal Ministry for 

Research and Technology, the Research Association for Internal Combustion Engines, and the 

German Association for the Petroleum Industry and Coal Chemistry. Tests were conducted using 

engine test benches and rolling test benches under representative operating conditions. Tests were 

performed on leaded gasoline engines, unleaded gasoline engines, diesel car engines, and diesel 

truck engines. The reported range of CDD/CDF emission rates across the test conditions in units 

of pg I-TEQ/L of fuel consumed are presented below. 

•	 Leaded fuel: 72 to 1,417 pg TEQDF-WHO98/L (52 to 1,184 pg I-TEQDF/L) 

•	 Unleaded fuel (not catalyst equipped): 102 to 181 pg TEQDF-WHO98/L (96 to 177 pg 

I-TEQDF/L) 

•	 Unleaded fuel (catalyst equipped): 9.6 to 28 pg TEQDF-WHO98/L (10 to 26 pg ITEQDF/L) 

•	 Diesel fuel (cars): 12 to 140 pg TEQDF-WHO98/L (10 to 130 pg I-TEQDF/L) 

•	 Diesel fuel (trucks): 79 to 82 pg TEQDF-WHO98/L (70 to 81 pg I-TEQDF/L 

Tables 4-2 through 4-6 show the results from tests with commercial fuels that were not 

conducted under normal operating conditions and for which congener-specific emission results 

were presented in Hutzinger et al. (1992). 

Although no specific details on the methodology used were provided, Hagenmaier (1994) 

reported that analyses of emissions of a diesel-fueled bus run on either the steady-state or the 

“Berlin cycle” showed no CDDs/CDFs present at a DL of 1 pg/L of fuel consumed for individual 

congeners. 

4-19
 
Gullett and Ryan (1997) reported the results of the first program to sample diesel engine 

emissions for CDDs/CDFs during actual highway and city driving. The exhaust emissions from a 

1991 Freightliner diesel tractor with a 10.3 L, six-cylinder Caterpillar engine—representative of 

the first generation of computerized fuel-controlled vehicles manufactured in the United 

States—were sampled during both highway and city routes. The average emission factor for the 

three highway tests conducted (15.1 pg I-TEQDF/km; range, 11.7 to 18.7 pg I-TEQDF/km; standard 

deviation, 3.5 pg I-TEQDF/km) was below the average of the two city driving tests by a factor of 3 

(49.9 pg I-TEQDF/kg; range, 3 to 96.8 pg I-TEQDF/km). DLs were considered to be zero in the 

calculation of these emission factors. The average of all five tests was 29 pg I-TEQDF/km, with a 

standard deviation of 38.3 pg I-TEQDF/km. This standard deviation reflects the 30-fold variation 

in the two city driving route tests. 

Geueke et al. (1999) analyzed dioxin emissions from heavy-duty vehicle diesel engines in 

Germany. Table 4-7 shows the results of the analysis. I-TEQ values ranged from 2 to 18 pg ITEQ/m3

, including one value so high that it could not be reproduced. Miyabara et al. (1999) 

analyzed CDDs/CDFs found in vehicle exhaust particles from a gasoline engine and a diesel 

engine in Japan. Table 4-8 presents the data from three tests conducted on the exhaust particles 

deposited on the tailpipe of the gasoline engine. TEQ values ranged from 3.44 to 5.32 pg TEQDFWHO98/g (3.46 to 5.33 pg I-TEQ/g) exhaust particles. Suspended particulate matter (PM) was 

also collected from an electrostatic precipitator (ESP) connected to a highway tunnel. The I-TEQ 

for the suspended PM was 257.5 pg TEQDF-WHO98/g (241.6 pg I-TEQ/g), two orders of 

magnitude higher than the TEQ for exhaust particles deposited on the tailpipes. Table 4-9 depicts 

the data from three tests conducted on the exhaust particles deposited on the tailpipe of the diesel 

engine. TEQ values ranged from 7.14 to 18.1 pg TEQDF-WHO98/g (7.13 to 14 pg I-TEQ/g) soot. 

4.1.2. Tunnel Emission Studies 

Several European studies and one U.S. study evaluated CDD/CDF emissions from 

vehicles by measuring the presence of CDDs/CDFs in tunnel air. This approach has the advantage 

of allowing the random sampling of exhaust from large numbers of cars with 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 bias and make interpretation of the 

findings difficult. Concerns have been raised that tunnel monitors detect resuspended particulates 

that 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, which may 

produce emission levels different from those of the transient driving cycle and cold engine starts 

that are typical of urban driving conditions. These studies are summarized below in chronological 

order. 

4-20
 
Table 4-7. Total dioxin emission concentrations from heavy-duty diesel 

engines in Germany 

Sample 

Concentration in exhaust

a 

(pg I-TEQ/m3) 

Sample volume 

(m3) 

Stationary engine 1 (S1) 6.1 32.89 

Stationary engine 2 (S2) 61b 10.35 

Stationary engine 3 (S3) 18c 10.73 

Stationary engine 4 (S4) 6.9 10.06 

Stationary engine 5 (S5) 6.6 10.06 

Truck engine 1 (V1) 9.7 10.03 

Truck engine 2 (V2) 2.1 10.07 

Truck engine 3 (V3) 2 9.99 

a

Detection limit for sampling: 4.1 pg/m3 for stationary samples, 4.5 pg/m3 for truck samples.
 
 b

Analysis could not be confirmed.
 
 c

High analytical detection limit (11 pg/m3).
 


Source: Geueke et al. (1999). 

Rappe et al. (1988) reported the CDD/CDF content of two air samples (60 m3/sample) 

collected from a tunnel in Hamburg, Germany, in January 1986 to be 0.44 and 0.59 pg TEQDFWHO98/m3

 (0.42 and 0.58 pg I-TEQDF/m3). Each sample was collected over a period of about 

60 hr. The tunnel handled 65,000 vehicles per day, of which 17% were classified as “heavy 

traffic.” The congener-specific results of the two samples are presented in Table 4-10. Ambient 

air measured in September 1986 at a nearby highway in Hamburg was reported to contain 

CDD/CDF levels two to six times lower than those measured in the tunnel. 

Larssen et al. (1990) and Oehme et al. (1991) reported the results of a tunnel study in Oslo, 

Norway, performed during April and May of 1988. Oehme et al. estimated total vehicle emissions 

by measuring CDD/CDF concentrations in tunnel inlet and outlet air of both the uphill and the 

downhill lanes. Emission rates for light-duty and heavy-duty vehicle classes were estimated by 

counting the number of light-duty and heavy-duty vehicles passing through the tunnel on 

workdays and a weekend and assuming a linear relationship between the percentage of the lightor heavy-duty traffic and the overall emission rate. Thus, the linear relationship for each emission 

rate was based on only two points (i.e., the weekday and the weekend measurements). 

4-21
 

Table 4-8. Levels of 2,3,7,8-chlorine-substituted congeners and total 

CDDs/CDFs in vehicle exhaust particles (pg/g) for gasoline engines and 

suspended particulate matter (SPM) 

Congener 

Gasoline 

Sample 1 Sample 2 Sample 3 SPM

2,3,7,8-TCDD 

Other TCDD 

<4.4 

6.21 

<2.1 

19 

<1.2 

7.41 

<5.2 

4,580 

2,3,7,8-TCDF 

Other TCDF 

3.98 

36.8 

5.17 

68.8 

3.53 

41.9 

108 

2,830 

1,2,3,7,8-PeCDD 

Other PeCDD 

<7.6 

<7.6 

<3.6 

11.5 

<2.1 

4.25 

40.8 

1,240 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

Other PeCDF 

5.58 

2.87 

24.4 

6.46 

5.24 

53.9 

3.07 

3.66 

38.3 

184 

107 

29,700 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

Other HxCDD 

4.3 

<3.8 

<3.8 

4.14 

<1.8 

2.66 

<1.8 

20.5 

0.86 

1.36 

0.63 

10.5 

42.3 

96.7 

71 

1,100 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

Other HxCDF 

6.85 

4.94 

<1.9 

<1.9 

47.2 

3.95 

4.48 

<0.9 

4.94 

23.7 

2.26 

2.35 

<0.5 

1.99 

15.2 

243 

231 

38.6 

387 

1,600 

1,2,3,4,6,7,8-HpCDD 

Other HpCDD 

<8.2 

<8.2 

11.4 

11.3 

7.64 

9 

1,700 

1,360 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

Other HpCDF 

<1.9 

<1.9 

<1.9 

12.7 

1.06 

8.36 

7.41 

0.5 

4.88 

1,330 

143 

778 

OCDD 

OCDF 

8.76 

4.78 

13.8 

5.09 

17 

3.03 

3,650 

1450 

Total CDD/CDF 160.8 294 186.7 53,010.4 

I-TEQ 3.73 5.33 3.46 241.6 

TEQDF-WHO98 3.72 5.32 3.44 257.5 

Source: Miyabara et al. (1999). 

The emission rates estimated in this study, in units of Nordic TEQ, are as follows: 

•	 Light-duty vehicles using gasoline (approximately 70 to 75% using leaded gas): 

uphill, 520 pg TEQ/km; downhill, 38 pg TEQ/km; mean, 280 pg TEQ/km 

4-22
 
Table 4-9. Levels of 2,3,7,8-chlorine-substituted congeners and total 

CDDs/CDFs in vehicle exhaust particles (pg/g) for diesel engines 

Congener 

Diesel 

Sample 1 Sample 2 Sample 3 

2,3,7,8-TCDD 

Other TCDD 

2.81 

267 

<14.4 

117 

<2 

86.9 

2,3,7,8-TCDF 

Other TCDF 

5.71 

84.2 

15.9 

335 

7.5 

313 

1,2,3,7,8-PeCDD 

Other PeCDD 

10.5 

165 

<28.8 

73.5 

8.15 

83.6 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

Other PeCDF 

3.17 

1.11 

27.3 

16.6 

<11.5 

211 

15.1 

9.52 

243 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

Other HxCDD 

3.39 

4.59 

2.14 

40.9 

<17.3 

<17.3 

<17.3 

28.1 

4.01 

4.6 

<1.5 

26.9 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

Other HxCDF 

1.29 

<1.2 

<1.2 

<1.2 

3.7 

15.9 

31.3 

<10.1 

<10.1 

182 

9.03 

8.22 

0.86 

9.58 

79 

1,2,3,4,6,7,8-HpCDD 

Other HpCDD 

8.78 

10.1 

<36 

<36 

1.24 

<1 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

Other HpCDF 

<1 

<1 

<1 

<8.6 

<8.6 

<8.6 

4.69 

<1 

6.28 

OCDD 

OCDF 

<2.8 

<4.4 

<23 

<36 

<0.5 

4.25 

Total CDD/CDF 641.7 1,026.3 925.4 

I-TEQ 10.6 7.14 14 

TEQDF-WHO98 13.0 7.14 18.1 

Source: Miyabara et al. (1999). 

•	 Heavy-duty diesel trucks: uphill, 9,500 pg TEQ/km; downhill, 720 pg TEQ/km; 

mean, 5,100 pg TEQ/km 

The mean values are the averages of the emission rates corresponding to the two operating 

modes: vehicles moving uphill on a 3.5% incline at an average speed of 37 mph and vehicles 

4-23
 

Table 4-10. CDD/CDF concentrations (pg/m3

) from European tunnel air studies 

Congener/congener group Germanya Germanya Belgiumb 

Norway 

(workdays)

c,d 

Norway 

(weekend)c,d 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND (0.01) 

0.31 

0.37 

1.19 

0.44 

1.9 

6.3 

0.06 

0.28 

ND (0.17) 

0.66 

ND (0.17) 

2 

6.4 

0.002 

0.025 

0.025 

0.042 

0.03 

0.468 

2.19 

0.02 

0.18 

0.06 

0.29 

0.25 

1.41 

0.1 

0.02 

0.04 

0.03 

0.03 

0.06 

0.16 

0.5 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.17 

0.4 

0.19 

0.26 

0.16 

ND (0.04) 

0.12 

1.2 

ND (0.16) 

ND (1.3) 

0.72 

0.36 

NR 

0.13 

0.15 

ND (0.05) 

ND (0.05) 

0.98 

ND (0.17) 

ND (1) 

0.013 

0.143 

0.039 

0.073 

0.093 

0.143 

0.004 

0.499 

0.074 

0.25 

0.58 

0.83 

0.78 

0.79 

0.62 

0.04 

0.74 

1.78 

0.22 

1.62 

0.07 

0.75 

0.58 

0.34 

0.31 

0.03 

0.13 

0.93 

0.14 

2.54 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF (nondetect set to 0) 

Total I-TEQDF (nondetect set to 1⁄2 DL) 

Total TEQDF-WHO98 (nondetect set to 0) 

Total TEQDF-WHO98 (nondetect set to 

 1⁄2 DL) 

10.51 

2.5 

0.58 

0.59 

0.73 

0.74 

9.40 

2.34 

0.42 

0.44 

0.55 

0.58 

2.78 

1.33 

0.096 

0.096 

0.106 

0.106 

2.31 

8.0 

0.91 

0.91 

1 

1 

0.84 

5.82 

0.48 

0.48 

0.49 

0.49 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

0.23 

2.5 

7.8 

3.4 

6.3 

3.5 

3.6 

2. 

1.9 

ND (1.3) 

0.22 

1.3 

2.7 

3.4 

6.4 

6.2 

4.1 

1.1 

1.2 

ND (1) 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

0.26 

1.78 

1.32 

1.31 

0.1 

13.20 

10.17 

6.42 

2.62 

1.62 

0.16 

0.41 

0.12 

0.23 

0.5 

1.7 

7.91 

2.08 

1.41 

2.54 

Total CDD/CDF (nondetect set to 0) 

Total CDD/CDF (nondetect set to 1⁄2 DL) 

31.2 

31.9 

26.6 

27.1 

NR 

NR 

38.8 

38.8 

17.06 

17.06 

a

Source: Rappe et al. (1988).
 
 b

Source: Wevers et al. (1992).
 
 c

Source: Oehme et al. (1991).
 
 d

Listed values are the differences between the concentrations at the inlet and the outlet of the northbound tunnel lanes.
 


DL = Detection limit
 


ND = Not detected (value in parenthesis is the detection limit)
 


NR = Not reported
 


4-24
 

moving downhill on a 3.5% decline at an average speed of 42 mph. Although Oehme et al. 

reported results in units of Nordic TEQ, the results in I-TEQDF should be nearly identical (only 

about 3 to 6% higher) because the only difference between the two TEQ schemes is the TEF 

assigned to 1,2,3,7,8-PeCDF (0.1 in Nordic TEQ and 0.05 in I-TEQDF), a minor component of the 

toxic CDDs/CDFs measured in the tunnel air. Table 4-10 presents the congener-specific 

differences in concentrations between the tunnel inlet and outlet concentrations. 

Wevers et al. (1992) measured the CDD/CDF content of air samples taken during the 

winter of 1991 inside a tunnel in Antwerp, Belgium. Background concentrations outside the 

tunnel were also determined. Two to four samples were collected from each location with two 

devices: a standard high-volume sampler with a glass fiber filter and a modified two-phase, highvolume sampler equipped with a glass fiber filter and a polyurethane foam (PUF) plug. The 

I-TEQDF concentration in the air sampled with the filter with the PUF plug was 74 to 78% of the 

value obtained with the standard high-volume sampler. However, the results obtained from both 

sets of devices indicated that the tunnel air had a CDD/CDF TEQ concentration about twice as 

high as that of the outside air (filter with PUF: 80.3 fg I-TEQDF/m ­ 3 for tunnel air vs. 35 fg I

TEQDF/m3 for outside air; filter only: 100 fg I-TEQDF/m3 for tunnel air vs. 58 fg I-TEQDF/m3 for 

outside air). The authors presented the congener-specific results for only one tunnel air 

measurement; these results are presented in Table 4-10. 

During October and November of 1995, Gertler et al. (1996, 1998) conducted a study at 

the Fort McHenry Tunnel in Baltimore, MD. Their stated objective was to measure CDD/CDF 

emission factors from in-use vehicles operating in the United States, with particular emphasis on 

heavy-duty trucks. The air volume entering and leaving the tunnel bore (the area that the traffic 

goes through) that is used by most of the heavy-duty trucks (i.e., approximately 25% of the 

vehicles using the bore are heavy-duty trucks) was measured, and the air was sampled for 

CDDs/CDFs during seven 12-hr sampling periods. Three of the samples were collected during 

daytime (6 a.m. to 6 p.m.) and four samples were collected during the night (6 p.m. to 6 a.m.). 

The air volume and concentration measurements were combined with information on vehicle 

counts (obtained from videotapes) and tunnel length to determine average emission factors. 

A total of 33,000 heavy-duty trucks passed through the tunnel during the seven sample 

runs (21.2 to 28.8% of all vehicles). The emission factors, calculated on the assumption that all 

CDDs/CDFs emitted in the tunnel were from heavy-duty trucks, are presented in Table 4-11. The 

average TEQ emission factor was reported to be 181.8 pg TEQDF-WHO98/km (172 pg ITEQDF/km). The major uncertainties identified by the study authors were tunnel air volume 

measurement, sampler flow volume control, and analytical measurement of CDDs/CDFs. 

4-25
 

Table 4-11. Baltimore Harbor tunnel study: estimated emission factors (pg/km) for heavy-duty die

4-26
 


Congener/congener group 

Run-specific emission factors 

Run 2 Run 3 Run 5 Run 6 Run 8 Run 9 

2,3,7,8-TCDD 24.5 61.6 0 21.2 37.8 40.1 1,2,3,7,8-PeCDD 40.2 20.6 15.4 5.6 38.4 0 1,2,3,4,7,8-HxCDD 18.2 25.2 46.5 8.3 64.5 0 1,2,3,6,7,8-HxCDD 37.5 28.2 64.3 19.6 153 71.1 1,2,3,7,8,9-HxCDD 53.6 56.5 91.6 48.4 280 126 1,2,3,4,6,7,8-HpCDD 0 401 729 111 2,438 963 OCDD 0 3,361 3,382 1,120 9,730 5,829 

2,3,7,8-TCDF 0 94.3 67.6 152.8 155.8 73.4 1,2,3,7,8-PeCDF 0 48.9 72.6 23.6 53.3 0 2,3,4,7,8-PeCDF 24.5 75.7 131 46.6 85 63.9 1,2,3,4,7,8-HxCDF 15.4 139 204 93.8 124 164 1,2,3,6,7,8-HxCDF 0.3 75.1 73.7 51 61.3 54.4 1,2,3,7,8,9-HxCDF 27.7 14.8 75.6 0 20.6 37.2 2,3,4,6,7,8-HxCDF 15.2 82.5 152 55.7 93 86.8 1,2,3,4,6,7,8-HpCDF 12.6 280 445 154 313 354 1,2,3,4,7,8,9-HpCDF 0 58.5 60.8 31.1 25 2.3 OCDF 0 239 401 175 416 534 

Total 2,3,7,8-CDD 174 3,954 4,328.8 4,328 1,334.1 7,029 Total 2,3,7,8-CDF 95.7 1,107.8 1,683.3 1,684 783.6 1,370 Total I-TEQDF 73.8 174.8 170.5 170 95.7 152.9 

Total TEQDF-WHO98 93.9 181.8 174.8 175 97.3 147.2 
Table 4-11. Baltimore Harbor tunnel study: estimated emission factors (pg/km) for heavy-duty die

(continued) 

4-27
 


Congener/congener group 

Run-specific emission factors 

Run 2 Run 3 Run 5 Run 6 Run 8 Run 9 

Total TCDD 245 0 140 165 311 109 Total PeCDD 110 21.9 83.3 35.6 174 0 Total HxCDD 677 0 753 54.5 2,009 1,666 Total HpCDD 0 802 1,498 142 5,696 1,933 Total OCDD 0 3,361 3,382 1,120 9,730 5,829 Total TCDF 0 901 1,314 656 2,416 1,007 Total PeCDF 124 119 1,152 78.4 1,055 282 Total HxCDF 136 319 852 67.6 444 719 Total HpCDF 0 223 814 144 513 354 Total OCDF 0 239 401 175 416 534 Total CDD/CDF 1,292 5,985.9 10,389.3 2,638.1 22,764 12,433 

Heavy-duty vehicles as % of total 

vehicles 

21.2 22 22.6 34 28.8 24.2 

aValues are based on the difference between the calculated chemical mass entering the tunnel and the mass exiting the tunnel. All calcu

 emission factors were set equal to zero. All CDD/CDF emissions were assumed to result from heavy-duty diesel-fueled vehicles. Source: Gertler et al. (1996, 1998). 
EPA’s Office of Transportation and Air Quality (OTAQ) reviewed the Gertler et al. (1996) 

study (Lorang, 1996) and found it to be technologically well done; no major criticisms or 

comments on the test methodology or protocol were offered, nor did OTAQ find any reason to 

doubt the validity of the emission factor determined by the study. OTAQ noted that the 

particulate emission rate for heavy-duty vehicles measured in the study (0.32 g/mile) was lower 

than the general particulate emission rate used by EPA (about 1 g/mile) and, thus, may 

underestimate CDD/CDF emissions under different driving conditions. OTAQ cautioned that the 

reported emission factor should be regarded only as a conservative estimate of the mean emission 

factor for the interstate trucking fleet under the driving conditions of the tunnel (i.e., speeds on the 

order of 50 mph, with those of the entering traffic slightly higher and those of the exiting traffic 

slightly lower). 

Figure 4-4 graphically presents the results of the studies by Rappe et al. (1988), Oehme et 

al. (1991), Wevers et al. (1992), and Gertler et al. (1996, 1998). The figure compares the 

congener profiles (i.e., congener concentrations or emission factors normalized to total 

concentration or emission factor of 2,3,7,8-substituted CDDs and CDFs) reported in the four 

studies. The dominant congeners in the Rappe et al., Wevers et al., and Gertler et al. studies are 

OCDD; 1,2,3,4,6,7,8-HpCDD; OCDF; and 1,2,3,4,6,7,8-HpCDF. With the exception of OCDD, 

these congeners are also the major congeners reported by Oehme et al. The Oehme et al. study 

also differs from the other tunnel studies in that the total of 2,3,7,8-substituted CDFs far exceeds 

the total of 2,3,7,8-substituted CDDs (by a factor of 2), whereas the other three observed just the 

opposite. 

4.1.3. National Emission Estimates 

Estimates of national CDD/CDF TEQ emissions for reference years 1987 and 1995 are 

presented in this section only for on-road vehicles using gasoline and diesel fuel. For reference 

year 2000, EPA’s Office of Air Quality Planning and Standards (OAQPS) developed national 

CDD/CDF TEQ emission estimates for on-highway gasoline and diesel vehicles, off-highway 

gasoline and diesel equipment, diesel railroad equipment, and diesel commercial marine vessels. 

4.1.3.1. Activity Information for On-Road Vehicles 

Reference year 2000 activity information for on-highway gasoline and diesel vehicles was 

estimated by OAQPS as county-level vehicle miles driven (VMD). The estimates include 

calculations by month, road type, and vehicle type. To develop the VMD, OAQPS relied on data 

supplied by the Federal Highway Administration (FHWA). 

4-28
 

Ratio (congener conc./total 2378CD

D/CDF conc.) 

0.7 

0.6 

0.5 

0.4 

0.3 

0.2 

0.1
 


0
 
 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 

Congener number
 


Norway (Row 1)

a
 
 Germany (Row 2)b Germany (Row 3)a 

Belgium (Row 4)c Baltimore, USA (Row 5)

d 

Figure 4-4. Tunnel air concentrations (congener numbers refer to the congeners 

in order as listed in Table 4-7). 

a

Source: Oehme (1991).
 
 b

Source: Rappe et al. (1988).
 
 c

Source: Wevers et al. (1992).
 
 d

Source: Gertler et al. (1996, 1998).
 


For on-highway gasoline-driven vehicles, OAQPS calculated a national activity level of 

4,071 billion km for 2000. The activity level for each vehicle type was 

Vehicle type Billion kilometers 

Light-duty vehicles 2,574.95 

Light-duty trucks 1 1,004.23 

Light-duty trucks 2 342.79 

Heavy-duty vehicles 131.97 

Motorcycles 17.70 

4-29
 
For on-highway diesel-fueled vehicles, OAQPS estimated a national activity level of 359 

billion km for 2000. The activity level for each vehicle type was 

Vehicle type Billion kilometers 

Light-duty vehicles 6.44 

Light-duty trucks 1 6.44 

2B-heavy diesel vehicles 33.80 

Light heavy-duty vehicles 25.75 

Medium heavy-duty vehicles 59.55 

Heavy heavy-duty vehicles 217.26 

Buses heavy-duty vehicles 9.66

 For reference year 1995, FHWA reported that 1,448 billion total vehicle miles (2,330 

billion km) were driven by automobiles and motorcycles in the United States. Trucks accounted 

for 1,271 billion km (790 billion miles) and buses accounted for 10 billion km (6.2 billion miles) 

(U.S. DOC, 1997). In 1992, diesel-fueled trucks accounted for 14.4% of total truck vehicle 

kilometers driven (VKD); gasoline-fueled trucks accounted for the remaining 85.6% (U.S. DOC, 

1995a). Applying this factor of 14.4% to the 1995 truck estimate of 1,271 billion km results in an 

estimate of 183 billion km driven by diesel-fueled trucks in 1995. 

All other VKD (2,947 billion km) are assumed to be by gasoline-fueled vehicles 

(nondiesel trucks, all automobiles, all buses, and all motorcycles); although a fraction of buses 

and automobiles use diesel fuel, the exact numbers are not known. It is further assumed that all of 

these kilometers were driven by unleaded gasoline-fueled vehicles because in 1992 only 1.4% of 

the gasoline supply was leaded fuel (EIA, 1993). Use of leaded fuel should have declined further 

by 1995 because its use in motor vehicles for highway use in the United States was prohibited as 

of December 31, 1995 (Federal Register, 1985a). 

For reference year 1987, an estimated 3,092 billion km were driven in the United States, of 

which trucks accounted for 887 billion km (U.S. DOC, 1995b). Diesel-fueled trucks accounted 

for 17.2% of total truck kilometers driven (U.S. DOC, 1995a). Applying this factor of 17.2% to 

the 1987 truck kilometer estimate of 887 billion results in an estimate of 153 billion km driven by 

diesel-fueled trucks. All other VKD (2,939 billion) are assumed to have been by gasoline-fueled 

vehicles. Leaded gasoline accounted for 24.1% of the gasoline supply in 1987 (EIA, 1993); thus, 

708 billion km are estimated to have been driven by leaded gasoline-fueled vehicles. The 

remaining 2,231 billion km are estimated to have been driven by unleaded gasoline-fueled 

vehicles. These mileage estimates are given a high confidence rating because they are based on 

U.S. Census Bureau transportation studies. 

4-30
 
4.1.3.2. Activity Information for Off-Road Uses 

Although on-road vehicles are the largest users of gasoline and diesel fuel, certain sectors 

of the economy account for significant amounts of farm, railroad, marine vessel, and other 

off-highway uses. Reference year 2000 activity information for off-highway gasoline and diesel 

equipment was estimated by OAQPs from NONROAD model runs prepared for the National 

Emissions Inventory. For off-highway gasoline-driven equipment, OAQPS calculated a national 

activity level of 23,091.01 million L for 2000. The activity level for each equipment type was 

Vehicle type Million liters 

Lawn and garden equipment, 4-stroke engines 8,100.78 

Pleasure craft, 2-stroke engines 3,607.50 

Commercial equipment, 4-stroke engines 3,255.45 

Recreational equipment, 2-stroke engines 2,032.77 

Recreational equipment, 4-stroke engines 1,782.93 

Pleasure craft, 4-stroke engines 1,374.10 

Lawn and garden equipment, 2-stroke engines 1,192.40 

Industrial equipment, 4-stroke engines 579.16 

Construction and mining equipment, 4-stroke engines 473.18 

Agricultural equipment, 4-stroke engines 306.62 

Industrial equipment, 4-stroke engines; other oil field equipment 124.92 

Construction and mining equipment, 2-stroke engines 102.21 

Commercial equipment, 2-stroke engines 79.49 

Logging equipment, 4-stroke engines 37.85 

Logging equipment, 2-stroke engines 26.50 

Airport ground support equipment, 4-stroke engines 7.57 

Railroad, 4-stroke engines 3.79 

Agricultural equipment, 2-stroke engines 3.40 

Industrial equipment, 2-stroke engines 1.14

 For reference year 2000, OAQPS calculated national activity levels of 40,125.37 million 

L for off-highway diesel-driven vehicles, 12,491.86 million L for diesel railroad equipment, and 

7,684.39 million L for diesel commercial marine vessels. For diesel commercial marine vessels, 

the national activity level comprises port emissions (5,905.24 million L) and underway emissions 

(1,968.41 million L). The activity level for each type of diesel railroad equipment was: 

Diesel locomotive type Million liters
 


Class I locomotives 10,561.30
 


Class II/III locomotives 700.30
 


Yard locomotives 794.94
 


Passenger trains 230.91
 


Commuter trains 215.77
 


4-31
 

The following paragraphs define each of the off-road fuel uses listed at the beginning of 

this section and present distillate fuel sales (in liters) in each sector for reference years 1987 and 

1995 (EIA, 1992, 1997a). For these sectors, the majority of “distillate fuel” sales are diesel fuels; 

a small fraction are fuel oils. The activity level information for reference years 1987 and 1995 is 

provided for informational purposes only because emission estimates for these years could not be 

calculated due to the lack of emission factors. 

Farm fuel use includes sales for use in tractors, irrigation pumps, and other agricultural 

machinery, as well as fuel used for crop drying, in smudge pots, and for space heating of 

buildings. Sales were 11,352.45 million L in 1987 and 13,158.1 million L in 1995. 

Railroad fuel use includes sales to railroads for any use, including diesel fuel for use in 

locomotives and fuel used for heating buildings operated by railroads. Sales were 10,788.42 L in 

1987 and 12,980.18 L in 1995. 

Marine vessel fuel use includes sales for the fueling of commercial or private boats such 

as pleasure craft, fishing boats, tug boats, and oceangoing vessels, including vessels operated by 

oil companies. Excluded are sales to the U.S. Armed Forces. Sales were 7,059.79 L in 1987 and 

8,854.08 L in 1995. 

Off-highway fuel use includes sales for use in construction equipment, including, e.g., 

earthmoving equipment, cranes, stationary generators, and air compressors, and sales for 

nonconstruction off-highway uses such as logging. Sales were 5,905.24 L in 1987 and 8,225.7 L 

in 1995. 

4.1.3.3 Emission Estimates 

Using the results of the studies discussed in Section 4.1.1, separate national annual 

emission estimates were developed for vehicles burning leaded gasoline, unleaded gasoline, and 

diesel fuel. 

Leaded gasoline. The literature indicates that CDD/CDF emissions occur from full 

combustion in 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 I-TEQDF/km from four cars fueled with a reference unleaded fuel to which lead 

(0.5 g/leaded gal) and a chlorinated scavenger were added. Marklund et al. (1990) reported much 

lower emissions in the exhaust of cars using a commercial leaded fuel (0.5 g/L) containing both 

1,2-dichloroethane and ethylene dibromide as scavengers (1.1 to 6.3 pg I-TEQDF/km). The 

difference in the emission measurements in the 1987 and 1990 studies was attributed to the 

different mix of scavengers used in the two studies, which may have resulted in preferential 

formation of mixed chlorinated and brominated dioxins and furans. 

4-32
 
Hagenmaier et al. (1990) reported TEQ emissions of 1,080 pg I-TEQDF/L fuel 

(approximately 129 pg TEQDF-WHO98/km [108 pg I-TEQDF/km]) from a car fueled with a 

commercial leaded fuel (lead content not reported). Bingham et al. (1989) reported emissions 

ranging from 1 to 39 pg I-TEQDF/km from four cars using gasoline with a lead content of 1.7 g/L 

in New Zealand. The German study reported by Schwind et al. (1991) and Hutzinger et al. (1992) 

measured emissions of 52 to 1,184 pg I-TEQDF/L (approximately 7.2 to 142 pg TEQDF-WHO98/km 

[5.2 to 118 pg I-TEQDF/km]) for cars under various simulated driving conditions. The tunnel 

study by Oehme et al. (1991) estimated that emissions from cars running primarily on leaded 

gasoline (70 to 75% of the cars) ranged from 38 to 520 pg Nordic TEQ/km. 

The average emission factor (see Table 4-4) was 532 pg TEQDF-WHO98/L (450 pg ITEQDF/L), as reported for the tailpipe emission studies performed using commercial leaded fuel 

(Marklund et al., 1990; Hagenmaier et al., 1990; Schwind et al., 1991), which presented analytical 

results for all 17 toxic CDD/CDF congeners. Assuming an average fuel economy of 10 km/L, this 

emission factor was approximately 53 pg TEQDF-WHO98/km (45 pg I-TEQDF/km). A low 

confidence rating is assigned to this emission factor because it is based on European fuels and 

emission control technologies, which may have differed from U.S. leaded-fuel and engine 

technologies, and because the factor is based on tests with only nine cars. 

Combining this average emission factor (53 pg TEQDF-WHO98/km [45 pg I-TEQDF/km], 

assuming nondetect values were zero) with the estimate for kilometers driven by leaded gasolinefueled vehicles in 1987 (708 billion km) suggests that 37.5 g TEQDF-WHO98 (31.9 g I-TEQDF) 

were emitted from vehicles using leaded fuels in 1987. Although some on-road vehicles used 

leaded fuel in 1995, further use of leaded fuel in motor vehicles for highway use in the United 

States was prohibited as of December 31, 1995 (Federal Register, 1985a). In 1992, the last year 

for which data are available on consumption of leaded gasoline by on-road vehicles, only 1.4% of 

the gasoline supply was leaded gasoline (EIA, 1993). A conservative assumption that 1% of the 

total VKD in 1995 (29.5 billion km of a total of 2,947 billion km) was by leaded gasoline-fueled 

vehicles, in conjunction with the emission factor of 53 pg TEQDF-WHO98/km (45 pg I-TEQDF/km), 

yields an annual emission of 1.6 g TEQDF-WHO98 (1.3 g I-TEQDF) in 1995. These emission 

estimates are assigned a low confidence rating on the basis of the low rating for the emission 

factor. 

Unleaded gasoline. The literature documenting results of European studies indicates that 

CDD/CDF emissions from vehicles burning unleaded fuels are lower than emissions from 

vehicles burning leaded gas with chlorinated scavengers. It also appears, based on the limited 

data available, that catalyst-equipped cars have lower emission factors than do noncatalystequipped cars. Marklund et al. (1987) did not detect CDDs/CDFs in emissions from two catalystequipped cars running on unleaded gasoline at a DL of 13 pg I-TEQDF/km. Marklund et al. (1990) 

4-33
 

reported emission factors of 0.36 and 0.39 pg I-TEQDF/km for two noncatalyst-equipped cars and 

an emission factor of 0.36 pg I-TEQDF/km for one catalyst-equipped car. Hagenmaier et al. (1990) 

reported an emission factor of 5.1 pg I-TEQDF/km for one noncatalyst-equipped car and 0.7 pg ITEQDF/km for one catalyst-equipped car. Schwind et al. (1991) and Hutzinger et al. (1992) 

reported emission factors of 9.6 to 17.7 pg I-TEQDF/km for several noncatalyst-equipped cars 

tested under various conditions; the reported emission factor range for catalyst-equipped cars was 

1 to 2.6 pg I-TEQDF/km. 

All automobiles running on unleaded gasoline in the United States are equipped with 

catalysts. The average emission factor reported for the tailpipe emission studies performed on 

catalyst-equipped cars (Hagenmaier et al. 1990; Schwind et al., 1991; Hutzinger et al., 1992) was 

15.6 pg TEQDF-WHO98/L (14.9 pg I-TEQDF/L) (calculated from the mean emission factor in Table 

4-6). A low confidence rating is assigned to this emission factor because the European fuels and 

emission control technology used may have differed from U.S. fuels and technology and also 

because the emission factor range is based on tests with only three catalyst-equipped cars. 

OAQPS calculated emissions for reference year 2000 for dioxins and furans from 

gasoline-fueled vehicles using the final version of the MOBILE6 model. On-road emissions were 

calculated by converting the emission factor of 15.6 pg TEQDF-WHO98/L (14.9 pg I-TEQDF/L) to a 

milligram-per-mile basis using a conversion factor of 3.78e–09 and assuming a fuel economy of 

21.5 miles/gal. The new emission factor was then multiplied by the corresponding county-level 

VMD in miles per year. The off-highway gasoline equipment emission estimates for reference 

year 2000 were developed by multiplying the mean emission factor of 15.6 pg TEQDF-WHO98/L 

(14.9 pg I-TEQDF/L) by 2000 activity estimates developed from NONROAD model runs prepared 

for the National Emissions Inventory. The activity estimates represent county-level gasoline 

consumption in gallons. The emission factor was converted from picograms per liter to 

milligrams per gallon by multiplying by a conversion factor of 3.78e–09. The use of these 

methodologies resulted in national estimates for reference year 2000 of 7 g TEQDF-WHO98 (6.7 g 

I-TEQDF) for on-highway gasoline vehicles and 0.36 g TEQDF-WHO98 (0.35 g I-TEQDF) for offhighway gasoline equipment. 

Applying the same emission factors from Gertler et al. (1996, 1998) and assuming an 

average fuel economy of 10 km/L yields an emission factor of 1.6 pg TEQDF-WHO98/km (1.5 pgITEQDF/km). Applying this emission factor to the estimate derived for VKD in 1995 by all 

gasoline-fueled vehicles (2,947 billion km) suggests that 4.7 g TEQDF-WHO98 (4.4 g I-TEQDF) 

were emitted from vehicles using unleaded fuels in 1995. Applying the same emission factors to 

the estimate derived above for VKD in 1987 by unleaded gasoline-fueled vehicles (2,231 billion 

km) suggests that 3.6 g TEQDF-WHO98 (3.3 g I-TEQDF) may have been emitted in 1987. The 

4-34
 

emission estimates for all reference years were assigned a low confidence rating on the basis of 

the low rating given to the emission factor. 

Diesel fuel. Limited data are available upon which to base an evaluation of the extent of 

CDD/CDF emissions resulting from diesel fuel combustion, and these data address only emissions 

from on-road vehicles; no emissions data are available for off-road diesel uses (construction 

vehicles, farm vehicles, and stationary equipment). Two U.S. tailpipe studies have been reported: 

CARB (1987) and Gullett and Ryan (1997). CARB reported a relatively high emission factor of 

676 pg I-TEQDF/km (nondetect values assumed to be zero) for one heavy-duty truck with a fuel 

economy of 5.5 km/L at 50 km/hr. Gullett and Ryan reported a range of emission factors for one 

diesel truck tested on six highway or city driving routes of 3 to 96.8 pg I-TEQDF/km (mean of 29 

pg I-TEQDF/km). 

The results of several tailpipe studies conducted in Europe have also been published. 

Marklund et al. (1990) reported no emissions at a DL of 100 pg I-TEQDF/L (or 18 pg I-TEQDF/km, 

assuming a fuel economy of 5.5 km/L) for one tested truck. Schwind et al. (1991) and Hutzinger 

et al. (1992) reported emission factors of 32 to 81 pg I-TEQDF/L (or 6 to 15 pg I-TEQDF/km, 

assuming a fuel economy of 5.5 km/L) for a truck engine run under various simulated driving 

conditions. Hagenmaier (1994) reported no emissions from a bus at a DL of 1 pg/L fuel 

consumed for individual congeners. For diesel-fueled cars, Hagenmaier et al. (1990) reported an 

emission factor of 24 pg I-TEQDF/L (or approximately 2.4 pg I-TEQDF/km) for one tested car. 

Schwind et al. and Hutzinger et al. reported emission factors of 5 to 13 pg I-TEQDF/km for a car 

engine run under various simulated driving conditions. 

The tunnel study by Oehme et al. (1991) generated an estimated mean emission factor of 

5,100 pg TEQ/km and a range of 720 to 9,500 pg TEQ/km (in units of Nordic TEQ) for dieselfueled trucks. Insufficient information was provided in Oehme et al. to enable an exact 

calculation of emissions in units of I-TEQDF or TEQDF-WHO98. However, based on the 

information that was provided, the mean emission factor in units of TEQ is approximately 5,250 

to 5,400 pg I-TEQDF/km. These indirectly estimated emission factors are considerably larger than 

those reported in engine studies by Marklund et al. (1990), Schwind et al. (1991), and Hutzinger 

et al. (1992); the CARB (1987) diesel truck emission factor falls at the low end of the range. 

Although aggregate samples representing several thousand heavy-duty diesel vehicles 

were collected in Oehme et al. (1991), several characteristics of the study introduce considerable 

uncertainty with regard to the use of the study’s results as a basis for estimating emissions in the 

United States: (a) heavy-duty vehicles represented only 3 to 19% of total vehicle traffic in the 

tunnel; (b) the majority of the light-duty vehicles were fueled with leaded gasoline, the 

combustion of which, as noted in Table 4-4, can release considerable amounts of CDD/CDFs; and 

4-35
 

(c) technology differences likely existed between the 1988 Norwegian and the 1987 and 1995 U.S. 

vehicle fleets. 

The tunnel study conducted in Baltimore, MD, by Gertler et al. (1996, 1998) shares the 

disadvantages of all tunnel studies relative to studies that directly measured CDDs and CDFs in 

tailpipe emissions. Specifically, tunnel studies rely on indirect measurements (rather than tailpipe 

measurements), which may introduce bias, and the emission factors calculated from these studies 

reflect driving conditions of only the vehicle fleet using the tunnel and not necessarily of the 

overall vehicle fleet under other driving conditions. 

However, the Gertler et al. study does have strengths that are lacking in the Oehme et al. 

(1991) tunnel study, and it has advantages over the two U.S. diesel truck tailpipe studies, 

including: (a) the study was conducted (fairly recently) in the United States and thus reflects 

current U.S. fuels and technology, (b) virtually no vehicle using the tunnel used leaded gasoline, 

(c) the tunnel walls and streets were cleaned 1 week prior to the start of sampling and, in addition, 

the study analyzed road dust and determined that resuspended road dust contributed only about 

4% of the estimated emission factors, (d) heavy-duty vehicles comprised, on average, a relatively 

large proportion (25.7%) of vehicles using the tunnel, and (e) a large number of heavy-duty 

vehicles—approximately 33,000—passed through the tunnel during the sampling period, which 

generates confidence that the emission factor is representative of interstate trucks. 

Considering the strengths and weaknesses of the available emission factor data from the 

tailpipe and tunnel studies, the mean TEQ emission factor reported by Gertler et al. (1996, 

1998)—182 pg TEQDF-WHO98/km (172 pg I-TEQDF/km)—is assumed to represent the best current 

estimate of the average emission factor for on-road diesel-fueled trucks. This emission factor is 

assigned a low confidence rating because it may not be representative of emission rates for the 

entire fleet of diesel-fueled trucks under the wide array of driving conditions encountered on the 

road. 

For reference year 2000, OAQPS developed national CDD/CDF TEQ emission estimates 

for on-highway diesel vehicles, off-highway diesel equipment, diesel railroad equipment, and 

diesel commercial marine vessels. For on-highway diesel vehicles, OAQPS combined the 

calculated mean emission factors from Gertler et al. (1996, 1998) with the OAQPS estimate for 

VMD. The picogram-per-kilometer emission factors were first converted to a miligram-per-mile 

basis using a conversion factor of 1.61e–09. OAQPS estimated national emissions of 65.4 g 

TEQDF-WHO98 (61.7 g I-TEQDF) from on-highway diesel-fueled vehicles for reference year 2000. 

For all years, the emissions from diesel vehicles were assigned a low confidence rating because 

the emission factors were assigned a low confidence rating. 

For off-highway diesel equipment, OTAQ developed the NONROAD emissions model to 

estimate emissions from nonroad (off-road) equipment types. However, the NONROAD model 

4-36
 

does not contain emission factors for calculating CDD/CDF emissions. To calculate emissions 

for 2000, OAQPS estimated fuel consumption, as reported by the May 2002 “Lockdown C” draft 

version of NONROAD, and multiplied this estimate by an average fuel efficiency of 7 miles/gal 

and the emission factor from Gertler et al. (1996, 1998). The NONROAD model does not contain 

activity estimates for commercial marine vessels and railroad equipment. 

OAQPS developed estimates for county-level diesel consumption, in gallons, for diesel 

commercial marine vessels and diesel railroad equipment and multiplied these estimates by an 

average fuel efficiency of 7 miles/gal and the emission factor from Gertler et al. (1996, 1998). 

The results from using these methodologies suggest that 22 g TEQDF-WHO98 (21 g I-TEQDF), 4.3 g 

TEQDF-WHO98 (4 g I-TEQDF), and 6.8 g TEQDF-WHO98 (6.4 g I-TEQDF) were emitted from offhighway diesel equipment, diesel commercial marine vessels, and diesel railroad equipment, 

respectively, in reference year 2000. 

 The use of the same emission factors from Gertler et al. (1996, 1998) and an assumption 

of an average fuel economy of 10 km/L results in an emission factor of 1.6 pg TEQDF-WHO98/km 

(1.5 pg I-TEQDF/km). Applying this factor to the estimate for VKD in 1995 in the United States 

by diesel-fueled trucks (183 billion km) suggests that 33.3 g TEQDF-WHO98 (31.5 g I-TEQDF) 

were emitted from diesel-fueled trucks in 1995. Combining the same emission factors with the 

estimate derived above for VKD in 1987 by diesel-fueled trucks (153 billion km) suggests that 

27.8 g TEQDF-WHO98 (26.3 g I-TEQDF) were emitted from diesel-fueled trucks in 1987. 

For 1987 and 1995 off-road diesel emissions, EPA used the emission factor from Gertler 

et al. (1996, 1998) and multiplied it by an average fuel efficiency of 2.98 km/L (U.S. EPA, 2003b) 

and a conversion factor of 1.61e–09 g-km/pg-mile to obtain emission factors of 0.51 ng ITEQDF/L and 0.54 ng TEQDF-WHO98/L. These emission factors are assigned a low confidence 

rating because they possibly are nonrepresentative of the source. Multiplying these emission 

factors by the 1987 activity factors for off-highway equipment (17,278.61 million L), marine 

vessels (7,068.35 million L), and railroad use (10,801.5 million L), EPA estimated the following 

emissions for 1987: 8.8 g I-TEQDF (9.4 g TEQDF-WHO98) for off-highway equipment, 3.6 g ITEQDF (3.8 g TEQDF-WHO98) for marine vessels, and 5.5 g I-TEQDF (5.8 g TEQDF-WHO98) for 

railroad use. Similarly, using the 1995 activity factors for off-highway equipment (21,409.71 

million L), marine vessels (8,864.81 million L), and railroad use (12,995.91 million L), EPA 

estimated the following emissions for 1995: 11 g I-TEQDF (12 g TEQDF-WHO98) for off-highway 

equipment, 4.5 g I-TEQDF (4.8 g TEQDF-WHO98) for marine vessels, and 6.6 g I-TEQDF (7 g 

TEQDF-WHO98) for railroad use. These emission estimates are given a low confidence rating 

because the emission factor may possibly be nonrepresentative of the source. 

4-37
 

4.2. WOOD COMBUSTION 

For reference year 1987, wood energy consumption is estimated to have been 2,437 trillion 

British thermal units (Btu), or 3.2% of the total primary energy consumed in the United States. In 

1995, wood fuel (including black liquor solids) provided about 2.6% (2,350 trillion Btu) of the 

total primary energy consumed (EIA, 1997b). Wood energy consumption in 2000 is estimated to 

have been 2,473 trillion Btu, or 2.5% of the total primary energy consumed (EIA, 2003a). The 

industrial sector is the largest consumer of wood fuel, accounting for 65% of total consumption in 

1987, 72% in 1995, and 80% in 2000; the residential sector accounted for 35% of total 

consumption in 1987, 25% in 1995, and 18% in 2000; and the commercial sector accounted for 

approximately 2% of total consumption in all three reference years (EIA, 2003a). 

These energy consumption estimates appear to include the energy value of black liquor 

solids, which are combusted in recovery boilers by wood pulp mills. In 1987, 1995, and 2000, the 

energy values of combusted black liquor solids were 950, 1,078, and 998 trillion Btu, respectively 

(American Paper Institute, 1992; American Forest and Paper Association, 1997; letter dated 

August 5, 2002, from W. Gillespie, National Council of the Paper Industry for Air and Stream 

Improvement, to C. Peck, Versar, Inc.). Subtracting the estimates of black liquor energy values 

from the 1987, 1995, and 2000 national totals for wood fuel yields 1,487, 1,272, and 1,475 trillion 

Btu, respectively. Assuming that 1 kg of oven-dried wood (2.15 kg of green wood) provides 

approximately 19,000 Btu (EIA, 1994), an estimated 78.3, 66.9, and 77.6 million metric tons of 

oven-dried wood equivalents were burned for energy purposes in 1987, 1995, and 2000, 

respectively. Of these totals, an estimated 44.8, 31.4, and 23 million metric tons were consumed 

by the residential sector and an estimated 33.2, 32.6, and 51.5 million metric tons were consumed 

by the industrial sector in 1987, 1995, and 2000, respectively. 

The following subsections discuss the results of relevant emission studies for the 

residential and industrial sectors and present annual TEQ emission estimates for reference years 

1987, 1995, and 2000. 

4.2.1. Flue Emissions from Wood Combustion (Residential) 

Several studies have provided direct measurement of CDDs/CDFs in flue gas emissions 

from wood-burning stoves and fireplaces (Schatowitz et al., 1993; Vikelsoe et al., 1993; Bremmer 

et al., 1994; Bröker et al., 1992; Launhardt and Thoma, 2000; Environment Canada, 2000). The 

findings of each of these studies are summarized below. 

4.2.1.1. Emissions Data 

Schatowitz et al. (1993) measured the CDD/CDF content of flue gas emissions from 

several types of wood burners used in Switzerland: a household stove (6 kW), automatic chip 

4-38
 

furnaces (110 to 1,800 kW), and a wood stick boiler (35 kW). Emissions were measured from the 

combustion of a variety of wood fuels (natural beech wood, natural wood chips, uncoated 

chipboard chips, and waste wood chips from building demolition). The results from the testing of 

the household stove are most relevant for assessing releases from residential combustion. The 

household stove was tested with the stove door both open and closed. The open-door stove can be 

assumed to be representative of fireplaces because both have an uncontrolled draft. Although the 

congener and congener group analytical results were not reported, the following emission factors 

(dry weight for wood, wet weight for household waste) and emission rates (corrected to 13% 

oxygen) for the household stoves and furnaces were reported. 

Stoves 

• Open-door burning of beech wood sticks: 0.77 ng I-TEQDF/kg 

(0.064 ng I-TEQDF/Nm3

) 

• Closed-door burning of beech wood sticks: 1.25 ng I-TEQDF/kg 

(0.104 ng I-TEQDF/Nm3) 

• Closed-door burning of household waste: 3,230 ng I-TEQDF/kg 

(114.4 ng I-TEQDF/Nm3) 

Furnaces 

• Natural wood chips: 0.79 to 2.57 ng I-TEQDF/kg 

• Chipboard chips (uncoated): 0.29 to 0.91 ng I-TEQDF/kg 

• Waste wood chips from building demolition: 26 to 173.3 ng I-TEQDF/kg 

Vikelsoe et al. (1993) studied emissions of CDD/CDF congener groups from residential 

wood stoves in Denmark. The wood fuels used in the experiments were seasoned birch, beech, 

and spruce, equilibrated to 18% absolute moisture. Four different types of stoves (including one 

experimental stove) were evaluated under both normal and optimal operating conditions (i.e., well 

controlled, with carbon monoxide [CO] emissions as low as possible). Total CDD/CDF 

emissions varied widely for the 24 fuel/stove type/operating condition combinations. Emissions 

from spruce were about twice as high as those from birch and beech. Surprisingly, the optimal 

operating condition led to significantly higher CDD/CDF emissions for two stove types but not 

for the other stoves. The predominant congener group for all experiments was TCDF. The 

weighted average emission factor and the flue gas concentration for wood stoves (considering 

4-39
 

wood and stove types) were reported to be 1.9 and 0.18 ng Nordic TEQ/Nm3

, respectively. 

Because Vickelsoe et al. did not measure congener levels, the reported emission factor and 

emission rate were estimated by assuming the same congener distribution in each congener group 

that had been found for municipal waste incinerators. 

Bremmer et al. (1994) reported results of testing performed with a cast-iron wood- burning 

stove with a combustion chamber lined with fire refractory clay. Measurements were conducted 

at three loads (maximum, average, and minimum) using clean wood as fuel. The emission factors 

ranged from 1 to 3.3 ng I-TEQDF/kg (average of about 2.2 ng I-TEQDF/kg). Bremmer et al. also 

reported results of testing conducted with a fireplace of a type that is common in the Netherlands. 

Measured emission factors from the burning of clean wood ranged from 13 to 28.5 ng I-TEQDF/kg 

(average of about 20 ng I-TEQDF/kg). The authors noted that the measured emission factors for 

fireplaces were considerably higher than those reported by others (see Bröker et al., 1992, below) 

and assigned “great uncertainty” to the emission factors. 

Bröker et al. (1992) reported results of a series of three tests with a wood stove and a 

fireplace. The average, minimum, and maximum emission factors measured for the wood stove 

tests ranged from 0.53 to 0.94 ng I-TEQDF/kg. The geometric mean of the two average values was 

0.71 ng I-TEQDF/kg. The average of the minimum and maximum emission factors measured for 

the fireplace tests ranged from 0.2 to 1.06 ng I-TEQDF/kg. The geometric mean of these two 

average values is 0.46 ng I-TEQDF/kg. 

Launhardt and Thoma (2000) conducted an investigation on organic pollutants from a 

domestic heating system using various solid biofuels. Tests were conducted using a multifuel 

furnace designed for domestic applications. Table 4-12 shows the average dioxin concentration in 

the flue gas for the four fuels used (spruce wood, wheat straw, hay, and triticale). The 

concentrations in the flue gas ranged from 52 to 891 pg TEQ/m3. 

Table 4-12. Average CDD/CDF concentration in flue gas while burning wood and 

crops 

Fuel Concentration (pg TEQ/m3) Number of trials 

Spruce wood 52 7 

Wheat straw 656 5 

Hay (set-aside land) 891 4 

Triticale (whole crop) 52 5 

Source: Launhardt and Thoma (2000). 

4-40 
Environment Canada (2000) conducted a study on the release of dioxins and furans into 

the atmosphere by residential wood combustors. The study analyzed two wood stoves believed to 

be representative of stoves used in Canada: a conventional wood stove that was popular in the 

early 1980s and an advanced combustion, noncatalytic, EPA-certified wood stove. Each stove 

was tested using hard maple and black spruce wood. Results from the study ranged from 0.222 to 

0.952 ng I-TEQ/kg wood (see Table 4-13). Because these tests took place in North America using 

indigenous wood, and they included the analysis of an EPA-certified wood stove, the mean value 

of the Environment Canada study (0.5 ng I-TEQ/kg wood) was used to determine the national 

emissions estimate for residential burning of clean wood in fireplaces and stoves. This emission 

factor is assigned a low confidence rating because it is judged to be nonrepresentative of all 

residential wood combustion (e.g., home fireplaces). 

Several studies have reported that combustion of treated or manufactured wood in stoves 

and fireplaces can result in significantly higher CDD/CDF emission factors. A few researchers 

(e.g., Vikelsoe et al., 1993) have reported high CDD/CDF emission rates when pentachlorophenol 

(PCP)-contaminated wood is combusted in residential wood stoves and furnaces. The European 

Inventory (Quab and Fermann, 1997) used the results of these studies to derive best estimates of 

CDD/CDF emission factors for combustion of “slightly contaminated wood (excluding PCP)” and 

“PCP-contaminated wood”: 50 and 500 ng I-TEQDF/kg, respectively. Although it is likely that 

there is some residential combustion of these types of wood in the United States, there are no 

corresponding activity level data upon which to base a national annual estimate of emissions. 

4.2.1.2. Activity Level Information 

In 1987, 22.5 million households in the United States burned wood (EIA, 1991). Wood 

was used as the primary heating fuel in 5 million of those households and as a secondary source 

for aesthetic purposes (i.e., in fireplaces) in 17.4 million (EIA, 1991, 1997b). Lower numbers 

were reported for 1995; wood was reported to be used as the primary fuel in only 3.53 million 

households (EIA, 1997b). More rural, low-income households consumed wood as a primary 

heating fuel than did other sectors of the population. The majority of these households used 

wood-burning stoves as the primary heating appliance. Although fireplaces were the most 

common type of wood-burning equipment in the residential sector, only 7% of fireplace users 

reported using fireplaces for heating an entire home (EIA, 1991, 1994). 

Residential wood consumption was 852 trillion Btu (44.8 million metric tons), or 35% of 

total U.S. wood energy consumption, in 1987 and 596 trillion Btu (31.4 million metric tons), or 

25% of total U.S. wood energy consumption, in 1995 (EIA, 1997b). An estimated 433 trillion Btu 

(23 million metric tons) of wood were consumed in residences in 2000 (EIA, 2003a). These 

production estimates are given high confidence ratings because they are based on recent 

government survey data. 

4-41
 
Table 4-13. CDD/CDF concentrations (pg TEQ/kg wood) in emissions from residential wood stoves 

U.S. EPA-certified Convention

Maple Spruce Maple 

Run 

Congener Run 1 Run 2 Run 3 Run 1 Run 2 Run 3 1 Run 2 Run 3 R

2,3,7,8-TCDD 212 214 256 82 110 91 68 75 56 1,2,3,7,8-PeCDD 108 138 117 41 66 57 34 56 47 1,2,3,4,7,8-HxCDD 21 16 17 10 18 14 8 13 9 1,2,3,6,7,8-HxCDD 21 16 17 10 18 14 8 13 9 1,2,3,7,8,9-HxCDD 21 16 17 10 18 14 8 13 9 1,2,3,4,6,7,8-HpCDD 2 2 2 1 2 2 1 1 2 OCDD 1 0 0 0 0 0 0 1 1 

2,3,7,8-TCDF 129 134 127 95 47 55 28 38 36 1,2,3,7,8-PeCDF 22 24 23 12 13 17 6 4 4 

2,3,4,7,8-PeCDF 243 371 350 186 149 302 85 78 66 

4-42
 1,2,3,4,7,8-HxCDF 22 31 23 12 13 23 27 11 10 

1,2,3,6,7,8-HxCDF 15 18 20 8 13 15 10 8 6 2,3,4,6,7,8-HxCDF 17 17 14 8 13 10 7 8 6 

1,2,3,7,8,9-HxCDF 10 10 11 8 13 10 4 8 6 1,2,3,4,6,7,8-HpCDF 2 1 2 1 2 2 2 2 1 1,2,3,4,7,8,9-HpCDF 2 1 1 1 2 2 1 1 1 OCDF 0 0 0 0 0 0 0 1 0 

TOTAL 848 1,009 997 485 497 628 297 331 269 2

Mean Emission Factor 951 537 299 

Source: Environment Canada (2000). 
OAQPS developed emission estimates for residential wood combustion from the results of 

a study by EPA’s Emission Factor and Inventory Group (U.S. EPA, 2001a). The activity data for 

residential wood combustion were based on the type of combustion unit, and the activity data for 

wood stoves and fireplaces with inserts were estimated on the basis of total amount of wood 

consumed in a year. OAQPS used 1997 national activity data to extrapolate an estimate for 1999 

by applying a growth rate factor based on wood energy consumption data from the Energy 

Information Administration (EIA). Activity data for fireplaces were estimated on the basis of 

number of homes in the U.S. with usable fireplaces, as reported by the U.S. Census Bureau. 

OAQPS assumed that the extent of wood consumption in residential combustion units is 

directly related to ambient temperature (with more wood consumption in colder climates). 

Historical climate data were used to assign each U.S. county to one of five climate zones, as 

defined by the National Climatic Data Center. Each climate zone was then assigned a percentage 

of total national wood consumption on the basis of information contained in the EIA’s Residential 

Energy Consumption database. 

The consumption in each climate zone was then allocated to individual counties in that 

zone. Each county was designated as urban or rural to reflect unit location preferences reported in 

the 1999 American Housing Survey, which estimated that 68% of fireplaces are found in urban 

areas, compared with 32% in rural areas. An estimated 69% of wood stoves are found in rural 

areas, compared with 31% in urban areas. Fireplaces with inserts were evenly split between urban 

and rural areas. In each zone, the total urban and rural county wood consumption was summed 

and an adjustment was made within the zone for each county’s consumption if the urban and rural 

totals did not match the expected percentage. These steps resulted in final cordwood consumption 

by county, which was converted to tons of wood consumed using a conversion factor of one cord 

of wood equaling 1.163 tons. 

Wood consumption estimates for stoves and fireplaces with inserts were further 

categorized to account for the different designs of units that exist in the marketplace. Different 

designs of stoves and inserts have been found to have different levels of emissions. According to 

data received from the Hearth Products Association, the three primary types of units currently in 

use are noncertified (92% of the stoves manufactured), certified noncatalytic (5.7%), and certified 

catalytic (2.3%). These proportions were applied to the national, state, and county cordwood 

consumption estimates prior to the application of emission factors. 

4-43
 

Activity levels were estimated to be as follows: 

Activity level 

Wood combustion category (million metric tons/yr) 

Fireplaces 2.79 

Fireplaces with inserts, certified catalytic 0.92 

Fireplaces with inserts, certified noncatalytic 0.47 

Fireplaces with inserts, noncertified 7.64 

Noncatalytic wood stoves 0.26 

Catalytic wood stoves 0.64 

Conventional wood stoves 10.60 

4.2.1.3. Emission Estimates 

The emission factor used to determine national emission estimates (0.5 ng I-TEQ/kg 

wood) was obtained from Environment Canada (2000) because it was the most comprehensive 

and recent study. Combining this emission factor with the mass of wood consumed in residences 

in 1987, 1995, and 2000 yields annual TEQ air emissions from this source of approximately 22, 

15.7, and 11.3 g I-TEQDF, respectively. These estimates are given a low confidence rating for all 

years because the emission factor was judged to be of low confidence. 

4.2.2. Stack Emissions from Wood Combustion (Industrial) 

4.2.2.1. Emissions Data 

Congener-specific measurements of CDDs/CDFs in stack emissions from industrial woodburning furnaces were measured by CARB at four facilities in 1988 (CARB, 1990b, c, d, e). 

Measurements of CDD/CDF congener groups and 2,3,7,8-TCDD and 2,3,7,8-TCDF were 

reported for one facility by EPA (U.S. EPA, 1987a). The National Council of the Paper Industry 

for Air and Stream Improvement (NCASI) presented congener-specific emission factors for five 

boilers tested during burns of bark and wood residue (NCASI, 1995). The average congenerspecific emission factors derived from the four CARB and five NCASI studies are presented in 

Table 4-14. Average congener and congener group profiles are presented in Figure 4-5a for the 

four CARB studies and in Figure 4-5b for the five NCASI studies. 

CARB (1990b) measured CDDs/CDFs in the emissions from a quad-cell wood-fired boiler 

used to generate electricity. The fuel consisted of coarse wood waste and sawdust from 

nonindustrial logging operations. The exhaust gases passed through a multicyclone before 

entering the stack. From this study, the average TEQ emission factor for total CDDs/CDFs was 

calculated to be 0.64 ng I-TEQDF/kg of wood burned. 

4-44
 

Table 4-14. CDD/CDF mean emission factors (ng/kg wood) for industrial wood combustors 

4-45
 


Congener/congener group 

Four facilities tested by CARB Five facilities tested by NCASI 

Nine fa

CAR

Nondetect set to 

zero 

Nondetect set to 

1⁄2 detection limit 

Nondetect set to zero 

Nondetect set to 1⁄2 

detection limit 

Nondetect set to 

zero 

2,3,7,8-TCDD 1,2,3,7,8-PeCDD 1,2,3,4,7,8-HxCDD 1,2,3,6,7,8-HxCDD 1,2,3,7,8,9-HxCDD 1,2,3,4,6,7,8-HpCDD OCDD 

0.007 0.044 0.042 0.086 0.079 0.902 6.026 

0.016 0.054 0.055 0.096 0.132 0.905 6.026 

0.066 0.11 0.179 0.191 0.522 0.635 1.317 

0.068 0.112 0.183 0.193 0.524 0.637 1.317 

0.04 0.079 0.115 0.138 0.321 0.745 3.329 

2,3,7,8-TCDF 1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

0.673 0.79 0.741 0.761 0.941 0.343 0.45 2.508 0.26 1.587 

0.673 0.79 0.741 0.768 0.941 0.35 0.491 2.749 0.344 1.59 

0.707 0.145 0.159 0.108 0.071 0.064 0.015 0.072 0.017 0.049 

0.719 0.149 0.164 0.111 0.073 0.067 0.017 0.074 0.02 0.06 

0.684 0.406 0.389 0.375 0.418 0.178 0.192 1.062 0.113 0.674 

Total TCDD 

Total PeCDD Total HxCDD Total HpCDD Total OCDD Total TCDF Total PeCDF Total HxCDF Total HpCDF Total OCDF 

0.151 

1.039 1.748 2.936 6.026 4.275 9.75 7.428 3.747 1.588 

0.154 

1.039 1.748 2.936 6.026 4.275 9.75 7.428 3.988 1.59 

1.628 1.958 1.792 1.12 1.317 4.532 1.548 0.536 0.111 0.049 

1.629 1.98 1.796 1.132 1.317 4.552 1.549 0.543 0.116 0.06 

0.969 1.521 1.663 1.821 3.329 4.353 4.93 3.316 1.58 0.674 

Total I-TEQDF 

Total TEQDF-WHO98 

0.82 0.84 

0.85 0.87 

0.4 0.46 

0.41 0.47 

0.56 0.6 

Total CDD/CDF 38.69 38.93 14.59 14.67 24.16 

CARB = California Air Resources Board NCASI = National Council for Air and Stream Improvement Sources: CARB (1990b, c, d, e); NCASI (1995). 
Ratio (mean congener group emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 0.25 

TCDD 

PeCDD 

HxCDD 

HpCDD 

OCDD 

TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF 

Figure 4-5a. Congener and congener group profiles for air emissions from 

industrial wood combustors (nondetects set equal to zero). 

Sources: CARB (1990b, c, d, e). 

4-46
 


0.3 
Ratio (mean congener group emission factor/total CDD/CDF emission factor) 

0 0.05 0.1 0.15 0.2 0.25 

TCDD
 


PeCDD
 


HxCDD
 


HpCDD
 


OCDD
 


TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF 

Figure 4-5b. Congener and congener group profiles for air emissions from bleached 

Kraft mill bark combustors (nondetects set equal to zero). 

Source: NCASI (1995). 

4-47
 


0.3 
In the second study (CARB, 1990d), CDDs/CDFs in the emissions from two spreaderstoker wood-fired boilers operated in parallel by an electric utility for generating electricity were 

measured. The exhaust gas stream from each boiler was passed through a dedicated ESP, after 

which the gas streams were combined and emitted to the atmosphere through a common stack. 

Stack tests were conducted when the facility burned fuels allowed by existing permits and when it 

burned a mixture of permitted fuel supplemented by urban wood waste at a ratio of 7:3. From this 

study, the average TEQ emission factor for total CDDs/CDFs was calculated to be 0.82 ng ITEQDF/kg of wood burned. 

In the third study (CARB, 1990e), CDDs/CDFs in the emissions from twin fluidized-bed 

combustors designed to burn wood chips for the generation of electricity were measured. The air 

pollution control device (APCD) system consisted of ammonia injection for controlling nitrogen 

oxides and a multicyclone and ESP for controlling PM. During testing, the facility burned wood 

wastes and agricultural wastes allowed by existing permits. From this study, the average TEQ 

emission factor for total CDDs/CDFs was calculated to be 1.32 ng I-TEQDF/kg of wood burned. 

In the fourth study (CARB, 1990f), CDDs/CDFs in the emissions from a quad-cell woodfired boiler were measured. During testing, the fuel consisted of wood chips and bark. The flue 

gases passed through a multicyclone and an ESP before entering the stack. From this study, the 

average TEQ emission factor for total CDDs/CDFs was calculated to be 0.5 ng I-TEQDF/kg of 

wood burned. 

NCASI (1995) presented stack emission test results for five boilers burning bark or wood 

residues. One of these facilities, which was equipped with a multicyclone, normally burned bark 

in combination with sludge and coal. Another facility, which was equipped with an ESP, 

normally fired pulverized coal. The other three facilities were spreader-stokers equipped with 

multicyclones or ESPs. Although stack gas flow rates were obtained during these tests, accurate 

measurements of the amounts of bark and wood fired were not made and had to be estimated from 

steam production rates. The average TEQ emission factor for these facilities was 0.46 ng TEQDFWHO98/kg (0.4 ng I-TEQDF/kg of feed). 

The mean of the emission factors derived from the four CARB studies and five NCASI 

studies—0.6 ng TEQDF-WHO98/kg wood (0.56 ng I-TEQDF/kg wood), assuming nondetect values 

were zero—is used in this document as the most representative of industrial wood combustion. 

This emission factor was assigned a medium confidence rating. However, these mean emission 

factors may not be appropriate for the combustion of waste wood containing elevated chlorine 

content. NCASI (1995) concluded that CDD/CDF emissions from facilities burning salt-laden 

wood residue may be considerably higher than those from facilities burning salt-free wood. 

Umweltbundesamt (1996) reported the results of stack gas testing at approximately 30 

facilities of varying design types burning various types of wood fuel. The author noted that 

4-48
 

CDD/CDF emissions were elevated when the combustion conditions were poor, as evidenced by 

elevated CO emissions, or when the fuel contained elevated chlorine levels. Umweltbundesamt 

attributed the correlation between elevated CDD/CDF emissions and elevated chlorine content of 

the fuel to the fire-retardant effects of chlorine, which may have inhibited complete combustion. 

The chlorine content of untreated wood and bark were reported as 0.001 to 0.01% by weight and 

0.01 to 0.02% by weight, respectively. Chipboard can contain up to 0.2% chlorine by weight 

because of the binding agents used to manufacture the chipboard. Preservative-treated wood and 

PVC-coated wood were reported to contain chlorine contents as high as 1.2 and 0.3% by weight, 

respectively. 

The facility tested by EPA in 1987 (U.S. EPA, 1987a) was located at a lumber products 

plant that manufactured overlay panels and other lumber wood products. 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 wood-fired boiler tested was a three-cell 

dutch oven equipped with a waste heat boiler. The feed wood was a mixture of bark, hogged 

wood, and green and dry planer shavings. The exhaust gases from the boiler passed through a 

cyclone and fabric filter (FF) prior to discharge from the stack. From this study, an average 

emission factor for total CDDs/CDFs of 1,020 ng/kg wood burned (range, 552 to 1,410 ng/kg) 

was reported for the three collected samples. An average TEQ emission factor of 17.1 ng 

I-TEQDF/kg wood burned (range, 7.34 to 22.8 ng/kg) was estimated by EPA using measured 

congener group concentrations and concentrations of 2,3,7,8-TCDD and 2,3,7,8-TCDF. Similar 

emission factors were reported by Luthe et al. (1998) from testing conducted during the 1990s at 

four Canadian coastal, salt-laden wood-fueled boilers—1.4, 2.6, 17.4, and 27.6 ng I-TEQDF/kg 

wood combusted. 

The overall average of the five tested facilities in Canada and the United States was 13.2 

ng I-TEQDF/kg of wood combusted. The confidence rating assigned to this emission factor is low 

because it is based on reporting of limited congener data at one U.S. facility and testing at four 

non-U.S. sources and because the fraction of salt-laden wood combusted across facilities is likely 

to have been highly variable. 

For reference year 2000, NCASI provided congener-specific estimates of CDD/CDF 

releases from the pulp and paper industry, including emissions from wood residue-fired boilers 

(letter dated August 5, 2002, from W. Gillespie, National Council of the Paper Industry for Air 

and Stream Improvement, to C. Peck, Versar, Inc.). The emission factors were taken from 

“NCASI Handbook of Chemical Specific Information for SARA (Superfund Amendments and 

Reauthorization Act) Section 313 Form R Reporting.” The factors provided in the handbook were 

compiled from valid test data supplied to NCASI by a variety of sources, including NCASI 

member companies that had performed the tests in response to a regulatory program. Data from 

4-49
 

11 bark and wood residue-fired boilers used by the forest products industry were used to calculate 

an emissions estimate. Concentrations of emissions from the wood residue-fired boilers were 

0.017 :g TEQDF-WHO98/ton wood (see Table 4-15). 

Table 4-15. NCASI CDD/CDF TEQ concentrations and emissions for wood 

residue-fired boilers 

Congener 

Wood-fired boiler emissions 

Wood-fired boiler ash not landfilled 

(72% of total ash landfilled) 

TEQDF-WHO98 

concentrations 

(median

μg/ton) 

Emissions 

(ng/yr) 

TEQ 

concentrations 

(ng/kg) 

Emissions 

(ng/yr) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

1,2,3,4,6,7,8,9-OCDD 

0 

0 

0 

4.00e!04 

5.00e!04 

1.05e!03 

5.69e!04 

0 

0 

0 

1.68e+07 

2.10e+07 

4.41e+07 

2.39e+07 

1.84e+00 

1.73e+00 

3.25e!01 

4.28e!01 

2.60e!01 

4.01e!01 

1.90e!02 

3.06e+08 

2.88e+08 

5.41e+07 

7.12e+07 

4.33e+07 

6.71e+07 

3.96e+06 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

1,2,3,4,6,7,8,9-OCDF 

4.40e!03 

7.50e!04 

5.00e!03 

9.00e!04 

7.00e!04 

2.10e!03 

9.00e!04 

2.80e!04 

1.10e!04 

2.10e!05 

1.85e+08 

3.15e+07 

2.10e+08 

3.78e+07 

2.94e+07 

8.82e+07 

3.78e+07 

1.18e+07 

4.62e+06 

8.82e+05 

4.20e+00 

3.35e!01 

3.23e+00 

2.21e!01 

1.60e!01 

5.40e!02 

3.80e!02 

4.10e!03 

1.30e!03 

5.40e!04 

7.06e+08 

5.56e+07 

5.37e+08 

3.68e+07 

2.66e+07 

8.98e+06 

6.32e+06 

6.82e+05 

2.16e+05 

1.64e+05 

TOTAL 1.72e!02 7.42e+08 1.32e+01 6.19e+08 

(ash not landfilled) 

2.21e+09 

(ash landfilled) 

NCASI = National Council for Air and Stream Improvement 

Source: Letter dated August 5, 2002, from W. Gillespie, National Council of the Paper Industry for Air and Stream 

Improvement, to C. Peck, Versar, Inc. 

4.2.2.2. Activity Level Information 

In 1987, 33.2 million metric tons of wood were burned for fuel in industrial furnaces. In 

1995, industrial wood consumption totaled 32.6 million metric tons. EIA (2003b) estimated that 

industrial wood consumption totaled 1988 trillion Btu (104.6 million metric tons) in 2000. This 

4-50
 

total becomes 51.5 million metric tons with the removal of kraft black liquor combustion. The 

majority of wood fuel consumed in the industrial sector consists of wood waste (chips, bark, 

sawdust, and hogged fuel). Consumption in the industrial sector is dominated by two industries: 

paper and allied products and lumber and wood products (EIA, 1994). These activity level 

estimates are assigned a high confidence rating because they are based on recent government 

survey data. 

Activity level data on combustion of salt-laden wood are not normally collected, even 

though the associated emission factor is greater than the factor associated with nonsalt-laden 

wood. Nonetheless, attempts have been made to estimate this activity level. NCASI combined 

the results from a 1995 survey of combustion units in the pulp and paper industry with those from 

an ad hoc telephone survey of mills in the Pacific Northwest (Oregon and Washington) to produce 

a conservative estimate of the amount of salt-laden wood burned at U.S. pulp and paper mills in 

1995: 254,000 metric tons (0.8% of the estimated 32.6 million metric tons of industrial wood 

consumed that year). NCASI suspected that a similar fraction of industrial wood combusted in 

1987 by pulp and paper mills was salt laden (letter dated October 8, 1998, from W. Gillespie, 

National Council of the Paper Industry for Air and Stream Improvement, to G. Schweer, Versar, 

Inc.). 

For purposes of the NCASI survey, salt-laden wood was defined as wood that had been 

transported, stored, or otherwise exposed to saltwater prior to being processed as fuel. None of 

the three responding mills in Oregon reported the use of salt-laden wood. Eight of the 13 

responding mills in Washington reported some combustion of salt-laden wood. Of the total wood 

consumed in the Washington mills, 17% was estimated to be salt-laden wood. 

As noted above, the majority of industrial wood combustion (97%) occurs in two 

industries: the paper and allied products industry and the lumber and wood products industry. 

The relative amount of wood combusted by each of these two industries was the same in 1990 and 

1992, the only years for which these statistics are readily available (EIA, 1991, 1994). It can be 

assumed that the percentage of total wood combusted nationally by the lumber and wood products 

industry that is salt laden is the same percentage as for the paper and allied products industry, 

0.8%; therefore, the total percentage of wood combusted by industry that is salt laden is 1.6%. 

For reference years 1987, 1995, and 2000, this equates to 0.5, 0.5, and 0.8 million metric tons, 

respectively. These activity level estimates are assigned a low confidence rating because they are 

possibly nonrepresentative of the activity levels for the source category combusting salt-laden 

wood. 

4.2.2.3. Emission Estimates 

 Applying the average TEQ emission factor from the four CARB and five NCASI studies 

(0.6 ng TEQDF-WHO98/kg wood [0.56 ng I-TEQDF/kg wood]) to the estimated quantities of 

nonsalt-laden wood burned by industrial facilities in 1987 (33.2 million metric tons), 1995 (32.6 

4-51
 

million metric tons), and 2000 (51.5 million metric tons) yields estimated TEQ emissions to air of 

19.9 g TEQDF-WHO98 (18.6 g I-TEQDF) in 1987, 19.6 g TEQDF-WHO98 (18.3 g I-TEQDF) in 1995, 

and 30.9 g TEQDF-WHO98 (28.8 g I-TEQ) in 2000. 

Applying the average TEQ emission factor from the five studies on boilers combusting 

salt-laden wood (13.2 ng I-TEQDF/kg wood) to the estimated quantities of salt-laden wood burned 

by industrial facilities in 1987 (0.5 million metric tons), 1995 (0.5 million metric tons), and 2000 

(0.8 million metric tons) yields estimated TEQ emissions to air of 6.6 g I-TEQDF in both 1987 and 

1995 and 10.6 g I-TEQDF in 2000. 

Total emissions for 1987, 1995, and 2000 are estimated to have been 26.5, 26.2, and 41.5 

g TEQDF-WHO98 (25.2, 24.9, and 39.4 g I-TEQDF), respectively. Of the 2000 estimate, NCASI 

estimated that 0.74 g TEQDF-WHO98/yr of dioxins were emitted from pulp and paper wood-fired 

boilers (letter dated August 5, 2002, from W. Gillespie, National Council of the Paper Industry for 

Air and Stream Improvement, to C. Peck, Versar, Inc.). As noted above, the total emissions are 

based on tests conducted at nine facilities in two industries that account for 97% of total industrial 

wood fuel combustion. The remaining 3% of industrial combustion and the combustion of wood 

by the commercial sector (for which no reliable activity level estimates are available) may not be 

well represented by the emission factors used above, particularly if poorly controlled combustors 

or treated wood (e.g., treated with PCP or plastics) are burned. The emission estimates for 1987, 

1995, and 2000 are given a low confidence level because the activity level estimates were 

assigned a low confidence rating. 

4.2.3. Solid Waste from Wood Combustion (Residential and Industrial) 

The measurement of CDDs/CDFs in chimney soot and bottom ash from wood-burning 

stoves and fireplaces has been reported by several researchers (Bumb et al., 1980; Nestrick and 

Lamparski, 1982, 1983; Clement et al., 1985; Bacher et al., 1992; Van Oostam and Ward, 1995; 

and Dumler-Gradl et al., 1995). 

Bumb et al. (1980) detected TCDDs (nondetect to 0.4 :g/kg), HxCDDs (0.2 to 3 :g/kg), 

HpCDDs (0.7 to 16 :g/kg), and OCDD (0.9 to 25 :g/kg) in residues from the wall of a home 

fireplace and from the firebrick of another home fireplace; for lack of a suitable analytical method, 

analysis was not performed for PeCDDs. Neither of the fireplaces sampled by Bumb et al. had 

burned preservative-treated wood. 

Nestrick and Lamparski (1982, 1983) expanded the research of Bumb et al. by conducting 

a survey of CDD concentrations in chimney soot from residential wood-burning units in three 

rural areas of the United States. Samples were collected from the base of six chimneys in each of 

the three study areas. Samples were not collected from units where any type of treated or 

manufactured wood had been burned. For lack of a suitable analytical method, analysis was not 

performed for PeCDDs. The results of this survey are summarized in Table 4-16. There was 

4-52
 

Table 4-16. CDD/CDF concentrations (ng/kg) in residential chimney soot from 

wood stoves and fireplaces 

U.S. Canadian Canadian 

Congener/ 

congener group 

U.S. east 

regiona 

U.S. west 

regiona 

central 

regiona 

German 

farmhouseb 

wood 

stovec 

Canadian 

fireplacec 

wood 

stoved 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

66 

NR 

250e 

250e 

208 

1,143 

2,033 

13.3 

NR 

522e

522e 

282 

1,653 

2,227 

66 

NR 

1,831e 

1,831e 

1,450 

6,160 

13,761 

150 

70 

35 

60 

30 

90 

90 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

ND (12) 

70 

ND (10) 

625 

281 

948 

530 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

930 

560 

590 

330 

400 

70 

200 

490 

40 

70 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

235 

58 

68 

51 

57 

8 

24 

97 

20 

41 

Total 2,3,7,8-CDD 3,950 5,219 21,437 525 NR NR 2,454 

Total 2,3,7,8-CDF NR NR NR 3,680 NR NR 659 

Total I-TEQDf $150 $165 $286 720 NR NR 211 

Total TEQDF-WHO98 $98 $163 $81 355 NR NR 246 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

1,987 

NR 

2,183 

2,104 

2,033 

NR 

NR 

NR 

NR 

NR 

269 

NR 

4,273 

3,243 

2,227 

NR 

NR 

NR 

NR 

NR 

1,511 

NR 

14,243 

12,603 

13,761 

NR 

NR 

NR 

NR 

NR 

3,900 

880 

600 

200 

90 

13,400 

6,100 

3,200 

720 

70 

ND (10) 

ND (10) 

ND (50) 

100 

200 

ND (10) 

ND (10) 

ND (50) 

ND (50) 

ND (50) 

ND (10) 

500 

1,700 

500 

400 

300 

1,400 

1,700 

400 

100 

11 

608 

3,450 

1,550 

530 

1,010 

948 

482 

154 

41 

Total CDD/CDF 8,307 10,012 42,118 29,160 300 7,000 8,784 

a

Source: Nestrick and Lamparski (1982, 1983); mean values listed, six samples collected in each region. b

Source: Bacher et al. (1992). c

Source: Clement et al. (1985). d

Source: Van Oostdam and Ward (1995); mean of two samples, nondetect values assumed to be zero. e

Analytical method could not distinguish between congeners; listed value is the sum of both congeners. 

ND = Not detected (value in parenthesis is the reported detection limit) 

NR = Not reported 

4-53
 

wide variation in the results across soot samples, with standard deviations for congeners and 

congener groups often equal to or exceeding the mean value; however, CDDs in each congener group 

were detected in the soot from almost all sampled units. The authors concluded that the results did 

not appear to present any easily discernible patterns with respect to geographic region, furnace 

operational parameters, or wood fuel type. They attributed the wide variability observed to 

differences in design of the units, which affected the sampling point or the conditions at the sampling 

point, and possible contamination of the fuel wood. 

Clement et al. (1985) analyzed chimney soot and bottom ash from residential wood stoves 

and fireplaces in Canada. The CDD/CDF congener concentrations are presented in Table 4-16 (soot) 

and Table 4-17 (bottom ash). CDD/CDF congeners were detected in all samples analyzed, although 

the relative amounts of the different congener groups varied considerably and inconsistently between 

wood-burning unit types and between ash and soot samples from the same unit. 

Table 4-17. CDD/CDF concentrations (ng/kg) in bottom ash from residential 

wood stoves and fireplaces 

Congener/congener 

group 

Canadian wood 

stove ash 

Canadian wood 

stove ash 

Canadian 

wood stove ash 

Canadian 

fireplace ash 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF

 ND (10)

 ND (10)

 ND (50) 

300 

2,600 

9,100 

2,200 

1,000 

700

 ND (50) 

100 

3,000 

10,000 

1,200 

900 

400 

4,600 

9,300 

1,000 

100 

100 

200 

700 

500 

100 

100 

200 

500 

300

 ND (50) 

ND (10) 

ND (10) 

300 

2,000 

3,100

 ND (10)

 ND (10) 

100 

400 

100 

Total CDD/CDF 15,900 30,600 2,700 6,000 a

No values were reported for individual congeners or for total 2,3,7,8-CDD, 2,3,7,8-CDF, or total TEQ. 

ND = Not detected (value in parenthesis is the reported detection limit) 

Source: Clement et al. (1985). 

Clement et al. also presented total CDD/CDF concentration data for bottom ash from 

open-air burning of wood. No analyses were reported for individual congeners. The results for 

the congener groups are shown below. The quantity of ash produced by the open-air burning test 

was not provided; hence, it is not possible to readily determine the quantities of CDDs/CDFs 

disposed of. 

4-54
 
Congener group Concentration (μg/kg) 

TCDDs 0.8 

PeCDDs 4.2 

HxCDDs 7.2 

HpCDDs 11 

OCDD 10 

TCDFs 2.2 

PeCDFs 7.6 

HxCDFs 8.2 

HpCDFs 11 

OCDF 1.7 

Bacher et al. (1992) characterized the full spectrum (mono through octa substitution) of 

CDD/CDF and BDD/BDF congeners in the soot from an old farmhouse in southern Germany. 

The chimney carried smoke from an oven that had used untreated wood at the rate of about 

5 m3/yr for more than 10 yr. The sample was taken during the annual cleaning by a chimney 

sweep. The only BDF detected was mono-BDF (230 ng/kg). No BDDs, BCDDs, or BCDFs were 

detected at a DL of 20 ng/kg. The results for the tetra- through octa-CDDs/CDFs are presented in 

Table 4-16. The results indicate that CDFs exceeded the CDDs in each congener group except 

octa. Also, the lower-chlorinated congener groups exceeded the higher-chlorinated congener 

groups for both the CDDs and the CDFs. The TEQ content of the chimney soot was 755 ng 

TEQDF-WHO98/kg (720 ng I-TEQDF/kg), of which less than 30% was due to CDDs. 

Van Oostdam and Ward (1995) analyzed soot from two wood stoves in British Columbia, 

Canada. The average TEQ concentration was 246 ng TEQDF-WHO98/kg (211 ng I-TEQDF/kg). 

The congener-specific results are presented in Table 4-16. The soot from a wood stove burning 

salt-laden wood in a coastal area was found to have an I-TEQDF content of 7,706 ng I-TEQDF/kg, 

or 20 to 90 times more than the concentrations found in the soot from the other two tested stoves. 

Dumler-Gradl et al. (1995) analyzed chimney soot samples collected by chimney sweeps 

from 188 residences in Bavaria, Germany. The summary results of the survey, the largest 

published survey of its kind to date, are presented in Table 4-18. As in Nestrick and Lamparski 

(1982, 1983) and Clement et al. (1985), CDDs/CDFs were detected in all samples; however, there 

was wide variability in total TEQ concentrations within and across unit type/fuel type 

combinations. 

The Washington State Department of Ecology (1998) reported CDD/CDF congener data 

for ash from hog fuel boilers at three paper mills. The data were compiled and evaluated to 

determine total I-TEQ concentrations and loading. Nondetect values were included as zero, onehalf the DL, or at the DL. The results, assuming nondetect values are at zero, are shown below. 

4-55
 

Table 4-18. CDD/CDF concentrations in chimney soot (Bavaria, Germany) 

Unit type Fuel type 

Number of 

samples 

Concentration 

(ng I-TEQDF/kg) 

Minimum Mean Maximum 

Oven Wood 33 10.4 2,015 15,849 

Tiled stove Wood 39 4 3,453 42,048 

Heating system Wood 9 16.9 1,438 20,450 

Oven Wood/coal 27 77.3 2,772 10,065 

Tiled stove Wood/coal 5 53.1 549 4,911 

Oven Wood, wood/coal, 

waste 

5 116.3 6,587 10,652 

Source: Dumler-Gradl et al. (1995). 

I-TEQDF I-TEQDF 

Location Type of residual (ng/kg) (mg/day) 

Daishowa America, 

Port Angeles Mixed ash 0.31 0.012 

Ft. James Fly ash 35.4 0.544 

Rayonier Filter ash 12,640 68.9 

Vacuum filter and grate 1,150 6.27 

Filter ash 2,299 12.5 

Fly ash 225 1.23 

Pohlandt and Marutzky (1994) presented CDD/CDF concentration data for various types 

of ash (bottom, furnace, boiler, and fly) from 12 wood-burning boilers. The fly ash samples from 

two wood-working industry boilers appeared to have the greatest concentrations of CDDs/CDFs. 

Table 4-19 lists the average congener concentration for the two boilers. Three boiler bottom ash 

samples contained detectable amounts of only total HpCDDs/HpCDFs and OCDD/OCDF. All 

the other boiler samples were from boilers that burned copper/chrome/boron-impregnated woods. 

These samples had total TEQs (assumed to be I-TEQs) ranging from 0.07 to 89 ppt, the highest 

being for the fly ash samples (52 and 89 ppt). The quantities produced by the boilers that were 

tested were not reported; hence, it is not possible to readily determine the quantities of 

CDDs/CDFs disposed of. 

The results of analyses of two ash samples from wood-burning facilities in New 

Hampshire were reported in a facsimile dated January 23, 2001, from Andrew Carpenter, 

Resource Management, Inc., to Stephen Schwartz, Versar, Inc. Both samples were from the 

4-56
 

Table 4-19. CDD/CDF concentrations (ng/kg) in fly ash from wood-working 

industry 

Congener/congener group 

Average 

concentration I-TEQDF TEQDF-WHO98 

2,3,7,8-TCDD 

Total TCDD 

<15 

1,730 

<15 

– 

<15 

– 

1,2,3,7,8-PeCDD 

Total PeCDD 

100 

1,250 

50 

– 

100 

– 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

Total HxCDD 

130 

150 

140 

750 

13 

15 

14 

– 

13 

15 

14 

– 

1,2,3,4,6,7,8-HpCDD 

Total HpCDD 

280 

470 

3 

– 

3 

– 

Total OCCD 300 0.3 0.03 

2,3,7,8-TCDF 

Total TCDF 

130 

1,300 

13 

– 

13 

– 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

Total PeCDF 

100 

120 

790 

5 

60 

– 

5 

60 

– 

1,2,3,4,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

1,2,3,6,7,8-HxCDF 

Total HxCDF 

40 

40 

<10 

150 

4 

4 

<1 

– 

4 

4 

<1 

– 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

Total HpCDF 

320 

<10 

570 

3 

<0.1 

– 

3 

<0.1 

– 

Total OCDF 60 0.06 0.006 

Estimated TEQ 89–90 89–90 

Source: Pohlandt and Marutzky (1994). 

burning of clean (i.e., untreated) wood chips, sawdust, and bark. The first sample was a 

combination of fly ash and bottom ash. The second sample was fly ash only, but it was a 

combination of fly ash from two wood-burning boilers. For the first sample, none of the 2,3,7,8­

substituted congeners were detected at DLs that ranged from 0.98 ng/kg for 2,3,7,8-TCDD and 

2,3,7,8-TCDF to 9.8 ng/kg for OCDD and OCDF. (All other congeners had a DL of 4.9 ng/kg.) 

For the second sample, all but two congeners were below DLs (which ranged from 0.379 to 0.831 

ng/kg). The two congeners that exceeded DLs were OCDD, at 1.261 ng/kg, and 1,2,3,4,6,7,8­

HpCDF, at 1.022 ng/kg. For this sample, assuming that the nondetected congeners were not 

4-57
 

present, the I-TEQDF concentration was 0.011 ng/kg. The quantities of the ash produced were not 

reported. 

In a CARB report of emissions from a wood waste-fired incinerator (CARB, 1990b), data 

are given for CDDs and CDFs for four ash samples. The concentrations of 2,3,7,8-substituted 

CDD/CDF congeners for each of those four tests were all below the DLs except for OCDD, which 

was detected in three samples at concentrations of 14, 18, and 32 ng/kg, and 2,3,7,8-TCDF, which 

was detected in one sample at a concentration of 2.2 ng/kg. The DLs for each CDD and CDF 

congener ranged from 0.63 ppt (for 2,3,7,8-TCDD) to 9.5 ppt (for HpCDF congeners). Total 

CDD and CDF values were given for each of the four samples. However, those values assumed 

that nondetected congeners were at the DL level. Consequently, the total CDD and total CDF 

values were biased high. The average of the four total CDD values was 28.8 ng/kg (range, 20.3 to 

44 ng/kg). The average of the four total CDF values was 21.9 ng/kg (range, 16 to 26.9 ng/kg). 

In CARB (1990d), data are presented for CDDs/CDFs for several samples of ESP waste 

ash from a wood-fired boiler. The report provides sample results for 2 weeks of sampling 

conducted at the facility. During the first week, the boiler burned fuels that were allowed by the 

facility permit; during the second week, the boiler burned a mixture containing 70% permitted 

fuel and 30% urban wood wastes. For the six samples collected over the 3 days of the first week, 

many of the concentrations of CDD/CDF congeners in the ESP ash were below the DLs. The 

reported CDD concentrations in ESP waste ash ranged from 24 to 264 ng/kg, and the CDF 

concentrations ranged from 12 to 151 ng/kg. However, those values assumed that nondetected 

congeners were present at the detection level. One sample did not have any nondetect values for 

CDDs. The total CDD concentration for this sample was 264 ng/kg, or about 11.4 ng/kg TEQDFWHO98 (8.3 ng/kg I-TEQDF). The TEQDF-WHO98 and I-TEQDF CDF concentrations for this 

sample were both less than 1.5 ng/kg. These values were less than 1 ng/kg for the other five 

samples. All of the samples had some nondetects for the CDF analysis. 

Six samples were also collected over 3 days during the second week of sampling, when the 

70/30 permitted/urban wood waste mix was burned. For the samples from the second week, the 

CDD concentrations in ESP waste ash ranged from 1,365 to 3,190 ng/kg, and the CDF 

concentrations ranged from 2,866 to 11,282 ng/kg. The study authors assumed that nondetected 

congeners were present at the detection level; however, this is a reasonable estimate for this data 

set because there was only one nondetect value. Table 4-20 presents the average congener 

concentrations for these samples. The report did not present quantities of ESP ash produced by 

the boiler; therefore, it is not possible to readily determine the quantities of CDDs/CDFs disposed 

of. 

4-58
 

Table 4-20. CDD/CDF concentrations (ng/kg) in electrostatic precipitator waste ash 

from wood-fired industrial boiler 

Congener/congener group 

Average 

concentration I-TEQDF TEQDF-WHO98 

2,3,7,8-TCDD 

Total TCDD 

17.85 

239 

17.85 

– 

17.85 

– 

1,2,3,7,8-PeCDD 

Total PeCDD 

30.67 

226.83 

15.33 

– 

30.67 

– 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

Total HxCDD 

20.33 

26.33 

23.33 

300 

2.03 

2.63 

2.33 

– 

2.03 

2.63 

2.33 

– 

1,2,3,4,6,7,8-HpCDD 

Total HpCDD 

325 

706.67 

3.25 

– 

3.25 

– 

Total OCDD 786.67 0.79 0.08 

Total CDD 2,439.17 44.22 58.85 

2,3,7,8-TCDF 

Total TCDF 

285 

2,713.33 

28.5 

– 

28.5 

– 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

Total PeCDF 

154.5 

641.67 

2,666.67 

7.73 

320.83 

– 

7.73 

320.83 

– 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

Total HxCDF 

244.83 

179.67 

296.67 

7.28 

1,520 

24.48 

17.97 

29.67 

0.73 

– 

24.48 

17.97 

29.67 

0.73 

– 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

Total HpCDF 

Total OCDF 

147.67 

21.33 

248.33 

48.33 

1.48 

0.21 

– 

0.05 

1.48 

0.21 

– 

0 

Total CDF 7,196.67 431.64 431.6 

Estimated TEQ 475.64 490.44 

Source: CARB (1990d). 

Appendix II in Luthe et al. (1998) shows TEQ concentrations (assumed to be I-TEQDF) in 

ashes collected from APCDs from “salt-laden” wood steam boilers. The I-TEQDF content of ashes 

from three of the primary multiclone hoppers varied significantly: 0.0978, 0.186, and 9.375 

4-59
 

μg/kg. Two samples of ash were taken from the secondary multiclone hoppers. The secondary 

multiclone removes dust from the primary multiclone emissions; therefore, the ash is finer than 

primary dust. The I-TEQDF values for the ash samples were 1.073 and 20.879 μg/kg. The ITEQDF values for two samples taken from the ESP that collected dust from the secondary 

multiclone emissions, which therefore was finer than multiclone dust, were 3.926 and 8.044 

μg/kg. No data were given for individual congeners. In fact, because the reference discusses only 

“dioxins,” it is unclear whether the TEQ data are for CDDs or for CDDs plus CDFs. Quantities of 

collected ash were not given. 

Table II in a report by Luthe et al. (1996) presents data for the “TEQs” (assumed to be ITEQs) on particulates from a secondary collection device for boilers at four paper mills burning 

salt-laden wood. Eight data points were given (two for each mill), the average of which was 3.6 

μg/kg. The range of values was 1.3 to 8 μg/kg. As in Luthe et al. (1998), no data were given for 

individual congeners. It is also unclear whether the TEQ data were for CDDs or for CDDs plus 

CDFs. Quantities of collected ash were not given. 

Table 5-16 in the National Dioxin Survey (U.S. EPA, 1987a) contains data indicating that 

the bottom ash from wood combustion from one source (it is not indicated whether it was a boiler) 

contained 140 ng/kg of 2,3,7,8-TCDD, 138,200 ng/kg of CDDs, and 7,400 ng/kg of CDFs. For a 

second wood combustion source, the ash contained no detectable 2,3,7,8-TCDD, but it did contain 

about 125 ng/kg of CDDs and nondetectable levels of CDFs. The FF dust from the second source 

contained 100 ng/kg of 2,3,7,8-TCDD, 1,143,600 ng/kg of CDDs, and 315,600 ng/kg of CDFs. 

Specific data for congeners and for ash/dust quantities were not given. 

NCASI also provided information on emissions from wood residue boiler ash for reference 

year 2000 (letter dated August 5, 2002, from W. Gillespie, National Council of the Paper Industry 

for Air and Stream Improvement, to C. Peck, Versar, Inc.). As with the boiler emissions, 

emission factors for the boiler ash were taken from “NCASI Handbook of Chemical Specific 

Information for SARA Section 313 Form R Reporting.” Total TEQ concentrations were 

estimated to be 13.2 ng/kg. Because 72% of the total ash produced is landfilled, emission 

estimates were 2.21 g TEQDF-WHO98/yr for ash landfilled and 0.62 g/yr for ash not landfilled (see 

Table 4-15). It is not known at this time whether the amount of dioxin in nonlandfilled ash results 

in an environmental release. Therefore, this value was not included in the inventory. 

4.3. OIL COMBUSTION 

The two major categories of fuel oils that are burned by combustion sources are distillate 

oils and residual oils. These oils are further distinguished by grade: numbers 1 and 2 are distillate 

4-60
 

oils, 5 and 6 are residual oils, and 4 is either distillate oil or a mixture of distillate and residual 

oils. Number 6 fuel oil is sometimes referred to as Bunker C. Distillate oils are more volatile and 

less viscous than residual oils. They have negligible nitrogen and ash content and usually contain 

less than 0.3% sulfur (by weight). Distillate oils are used mainly in domestic and small 

commercial applications. The heavier residual oils (5 and 6), being more viscous and less volatile 

than distillate oils, must be heated for ease of handling and to facilitate proper atomization. 

Because residual oils are produced from the residue after the lighter fractions (gasoline, kerosene, 

and distillate oils) are removed from the crude oil, they may contain significant quantities of ash, 

nitrogen, and sulfur. Residual oils are used mainly in utility, industrial, and large commercial 

applications (U.S. EPA, 1995a). 

4.3.1. Institutional/Commercial and Residential Oil Combustion 

No testing information on the CDD/CDF content of air emissions from institutional/ 

commercial or residential oil-fired combustion units in the United States could be located. 

However, EPA (U.S. EPA, 1997a) estimated CDD/CDF congener group and TEQ emission 

factors using average CDD/CDF concentrations reported for soot samples from 21 distillate fuel 

oil-fired furnaces used for central heating in Canada and a particulate emission factor for distillate 

fuel oil combustors (300 mg/L oil) obtained from AP-42 (U.S. EPA, 1995a). The TEQ emission 

factor estimate in U.S. EPA (1997a) was derived using the calculated emission factors for 2,3,7,8­

TCDD, 2,3,7,8-TCDF, and the 10 congener groups. These emission factors are presented in Table 

4-21 and the congener group profile is presented in Figure 4-6. 

For reference year 1987, assuming a barrel of oil contains 42 gallons, distillate fuel oil 

sales to the residential sector and the commercial sector totaled 28.1 billion L (177 million 

barrels) and 16.2 billion L (102 million barrels), respectively (EIA, 1999). Residual oil sales to 

the commercial sector in 1987 totaled 6.7 billion L (42 million barrels) (EIA, 1999). Using the 

emission factor presented in Table 4-21 (150 pg I-TEQDF/L oil combusted [190 pg TEQDFWHO98/L oil combusted]), EPA estimated that 4.22 g I-TEQDF (5.35 g TEQDF-WHO98) were 

emitted in 1987 for the residential sector. For the institutional/commercial sector, EPA estimated 

TEQ emissions of 1.34 g I-TEQDF (1.54 g TEQDF-WHO98) for residual oil and 3.24 g I-TEQDF 

(3.73 g TEQDF-WHO98) for distillate oil for 1987, using an emission factor of 200 pg I-TEQDF/L 

oil combusted (230 pg TEQDF-WHO98/L oil combusted) (see Section 4.3.2). Because the 

representativeness of the emission factor to 1987 emissions is uncertain and may not be 

representative, this estimate is assigned a low confidence rating. 

4-61
 
Table 4-21. Estimated CDD/CDF emission factors for oil-fired residential 

furnaces 

Congener/ 

congener group 

Mean 

facility 

emission 

factor 

(pg/L oil) WHO-TEF 

Emission 

factor (pg 

TEQDF ­

WHO98/L 

oil) I-TEF 

Emission 

factor (pg 

I-TEQDF/L 

oil) 

2,3,7,8-TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

OCDD 

56 

82 

66 

63 

66 

1 

1 

0.1 

0.01 

0.0001 

56 

82 

7 

1 

0 

1 

0.5 

0.1 

0.01 

0.001 

56 

41 

7 

1 

0 

2,3,7,8-TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

OCDF 

53 

420 

170 

73 

30 

0.1 

0.05 

0.1 

0.01 

0.0001 

5 

21 

17 

1 

0 

0.1 

0.05 

0.1 

0.01 

0.001 

5 

21 

17 

1 

0 

TOTAL 190 149 

Source: U.S. EPA (1997a). 

For reference year 1995, a low confidence estimate of potential national TEQ emissions 

from this source category was made using the same emission factors as for the 1987 estimates. 

Distillate fuel oil sales to the residential and commercial sector totaled 26.2 and 13.5 billion L, 

respectively, in 1995 (EIA, 1997a). Applying the respective emission factors to these fuel oil 

sales estimates results in estimated emissions of 3.93 g I-TEQDF (4.98 g TEQDF-WHO98) for the 

residential sector and 2.7 g I-TEQDF (3.11 g TEQDF-WHO98) for the institutional/commercial 

sector in 1995. Residual oil sales to the commercial sector in 1995 totaled 3.7 billion L (23 

million barrels) (EIA, 1999). Applying the emission factor of 200 pg I-TEQDF/L oil combusted 

(230 pg TEQDF-WHO98/L oil combusted) (see Section 4.3.2) yields TEQ emissions of 0.73 g I

TEQDF (0.84 g TEQDF-WHO98) for residual oil in 1995. 

For reference year 2000, EPA/OAQPS developed national emission estimates for residual 

oil and distillate oil consumed in institutional/commercial heating and distillate oil consumed in 

residential heating. EPA used state-level 2000 activity data (EIA, 1999), which were allocated to 

­

4-62
 

Figure 4-6. Congener group profile for air emissions from residential 

oil-fueled furnaces. 

Source: U.S. EPA (1995b). 

counties by the 1999 year county-to-state proportion of employment for numerous SIC codes, as 

identified in the U.S. Census Bureau’s 2000 census. EPA estimated that 2.82 billion L of residual 

oil and 12.7 billion L of distillate oil were consumed in institutional/commercial heating in 2000. 

Applying the emission factor of 200 pg I-TEQDF/L oil combusted (230 pg TEQDF-WHO98/L oil 

combusted) (see Section 4.3.2) to these activity levels yields TEQ emissions of 0.56 g I-TEQDF 

(0.65 g TEQDF-WHO98) for residual oil and 2.53 g I-TEQDF (2.92 g TEQDF-WHO98) for distillate 

oil for 2000. EPA/OAQPS estimated that 23.9 billion L of distillate oil were consumed for 

residential heating in 2000. Using the emission factors discussed above from U.S. EPA (1997a), 

EPA/OAQPS estimated emissions of 3.59 g I-TEQDF (4.54 g TEQDF-WHO98) from distillate oil 

used for residential heating in 2000. 

4-63
 
4.3.2. Utility Sector and Industrial Oil Combustion 

Preliminary CDD/CDF emission factors were reported for oil-fired utility boilers using the 

results of boiler tests conducted over several years (U.S. EPA, 1997a). The data are for a 

composite of various furnace configurations and APCD systems. Table 4-22 lists the median 

emission factors presented by EPA. The congener and congener group profiles based on these 

data are presented in Figure 4-7. The median I-TEQDF emission factor was reported to be 366 pg 

TEQDF-WHO98/L (314 pg I-TEQDF/L) oil burned. 

In 1993, the Electric Power Research Institute (EPRI) sponsored a project to gather 

information of consistent quality on power plant emissions. The Field Chemical Emissions 

Measurement (FCEM) project included testing of two cold-sided, ESP-equipped, oil-fired power 

plants for CDD/CDF emissions (EPRI, 1994). The averages of the congener and congener group 

emission factors reported for these two facilities are presented in Table 4-22. The average TEQ 

emission factor was 93.6 pg TEQDF-WHO98/L (83.1 pg I-TEQDF/L) oil burned when nondetect 

values were treated as zero. 

The TEQ emission factors reported by EPRI (1994) were less than the median TEQ 

emission factor reported by EPA by a factor of 3 to 4 (U.S. EPA, 1997a). For purposes of this 

assessment, the EPA median and EPRI mean emission factors were averaged, for an emission 

factor of 230 pg TEQDF-WHO98/L (200 pg I-TEQDF/L). Although the estimated emission factors 

are assumed to be the current best estimates for utility/industrial oil burning, they are assigned a 

low confidence rating. 

Residual fuel oil sales totaled 77.3 billion L in 1987 and 46.6 billion L in 1995 (EIA, 

1992, 1997a). Vessel bunkering was the largest consumer (48% of sales), followed by electric 

utilities and the industrial sector. A high confidence rating is assigned to these production 

estimates. Application of the TEQ emission factor of 230 pg TEQDF-WHO98/L (200 pg ITEQDF/L) to these residual fuel oil sales results in estimated TEQ emissions of 17.8 g TEQDFWHO98 (15.5 g I-TEQDF) for 1987 and 10.7 g TEQDF-WHO98 (9.3 g I-TEQDF) for 1995. 

For reference year 2000, OAQPS developed national emission estimates for residual and 

distillate oil consumption for the industrial sector. OAQPS used state-level 2000 activity data 

(EIA, 2003a), which were allocated to counties by the 1999 county-to-state proportion of 

employment for numerous SIC codes, as identified by the 2000 census. OAQPS estimated that 

7.33 billion L of residual oil and 31.5 billion L of distillate oil were consumed in the industrial 

sector in 2000. OAQPS combined these national activity levels with the emission factor of 230 

pg TEQDF-WHO98/L (200 pg I-TEQDF/L) to estimate 2000 TEQ emissions of 1.69 g TEQDFWHO98 (1.47 g I-TEQDF) from residual oil consumption and 7.25 g TEQDF-WHO98 

4-64
 
Table 4-22. CDD/CDF emission factors (pg/L oil) for oil-fired utility/industrial 

boilers 

Congener/congener group 

U.S. EPA (1997a) 

median 

emission factora,b 

EPRI (1994) mean emission factora,c 

Nondetect set to 

zero 

Nondetect set to 1⁄2 

detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

117 

104 

215 

97 

149 

359 

413 

0 

24.7 

63.3 

65.8 

79.7 

477 

2,055 

26.6 

43.1 

108 

79.3 

102 

546 

2,141 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

83 

77 

86 

109 

68 

104 

86 

169 

179 

179 

0 

64.1 

49.3 

76.5 

35.4 

0 

23.8 

164 

0 

0 

35.7 

73.9 

59.6 

94.9 

45.2 

37.7 

42.2 

218 

137 

139 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

1,454 

1,140 

314.6 

366.1 

2,765.5 

413.1 

83.1 

93.6 

3,046 

883.2 

147.4 

166.9 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

102 

104 

145 

359 

413 

90 

131 

172 

27 

179 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

Total CDD/CDF 1,722 NR NR a

Assumes a density for residual fuel oil of 0.87 kg/L.
 
 b

Number of facilities not reported.
 
 c

Based on two cold-sided power plants equipped with electrostatic precipitators.
 


NR = Not reported 

4-65
 
Figure 4-7. Congener and congener group profiles for air emissions from 

industrial oil-fueled boilers. 

Source: U.S. EPA (1995b; 1997a). 

4-66 
(6.3 g I-TEQDF) from distillate oil combustion. Emission estimates for all reference years are 

assigned a low confidence rating on the basis of the low rating for the emission factor. 

4.3.3. Used Oil Combustion 

The emission factors derived by EPA (U.S. EPA, 1997a) and EPRI (1994) were based on 

combustion of virgin oil by utility boilers. Significantly greater emission factors have been 

reported by Bremmer et al. (1994) for combustion of used oil by smaller combustion units in the 

Netherlands. Flue gases from a garage stove consisting of an atomizer fueled by spent lubricating 

oil from diesel engines (35 mg Cl!/kg) were reported to contain 0.1 ng I-TEQDF/Nm3 (2,000 pg ITEQDF/kg) oil burned. The flue gases from a hot water boiler consisting of a rotary cup burner 

fueled with the organic phase of rinse water from oil tanks (340 mg Cl!/kg) contained 0.2 ng 

I-TEQDF/Nm3 (4,800 pg I-TEQDF/kg) oil burned. The flue gases from a steam boiler consisting of 

a rotary cup burner fueled by processed spent oil (240 mg Cl!/kg) contained 0.3 ng I-TEQDF/Nm3 

(6,000 pg I-TEQDF/kg) oil burned. The emission factor for a ferry burning heavy fuel oil 

containing 11 ng/kg organic chlorine was 3,200 to 6,500 pg I-TEQDF/kg oil burned. From these 

data, Bremmer et al. derived an average emission factor for combustion of used oil of 4,000 pg ITEQDF/kg oil burned. 

Bremmer et al. (1994) also reported measuring CDD/CDF emissions from a river barge 

and a container ship fueled with gas oil (less than 2 ng/kg organic chlorine). The exhaust gases 

contained from 0.002 to 0.2 ng I-TEQDF/Nm3. From these data, Bremmer et al. derived an average 

emission factor for inland oil-fueled vessels of 1,000 pg I-TEQDF/kg oil burned. 

The applicability of these emission factors to used oil combustors in the United States is 

uncertain. Therefore, estimates of potential emissions from used oil combustion in the United 

States are not being developed at this time. 

4.4. COAL COMBUSTION 

During 2000, coal consumption accounted for approximately 18.9% of the energy 

consumed in the United States from all sources (EIA, 1999). Of the 980 million metric tons of 

coal consumed in 2000, 891 million metric tons (90.9%) were consumed by electric utilities, 

including independent power producers; 85.4 million metric tons (8.7%) were consumed by the 

industrial sector, including 26.2 million metric tons consumed by coke plants; and 3.7 million 

metric tons (0.4%) were consumed by residential and commercial sources (EIA, 1999). 

In 1995, coal consumption (872 million metric tons) accounted for approximately 22% of 

the energy consumed from all sources in the United States (U.S. DOC, 1997). Of this total, 

4-67
 

88.4% (771 million metric tons) were consumed by electric utilities; 11% (96 million metric 

tons) were consumed by the industrial sector, including consumption of 30 million metric tons by 

coke plants; and 0.6% (5.3 million metric tons) were consumed by residential and commercial 

sources (EIA, 1997b). 

In 1987 a total of 759 million metric tons of coal were consumed by electric utilities (651 

million metric tons), coke plants (33.5 million metric tons), other industries (68.2 million metric 

tons), and the residential and commercial sectors (6.3 million metric tons) (EIA, 1995). 

All of the above estimates are assigned a high confidence rating because they are based on 

detailed studies specific to the United States. 

The following two sections discuss the results of relevant emission studies for the 

utility/industrial and residential sectors and present annual TEQ emission estimates for reference 

years 1987, 1995, and 2000. 

4.4.1. Utilities and Industrial Boilers 

Few studies have been performed to measure CDD/CDF concentrations in emissions from 

coal-fired plants. Those studies did not have the congener specificity or DLs necessary to fully 

characterize this potential source (U.S. EPA, 1987a; NATO, 1988; Wienecke et al., 1992). The 

results of more recent testing of coal-fired utility and industrial boilers in the Netherlands 

(Bremmer et al., 1994), the United Kingdom (Cains and Dyke, 1994; CRE, 1994), Germany 

(Umweltbundesamt, 1996), and the United States (Riggs et al., 1995; EPRI, 1994) have achieved 

lower DLs. 

Bremmer et al. (1994) reported the results of emission measurements at two coal-fired 

facilities in the Netherlands. The emission factor reported for a pulverized coal electric power 

plant equipped with an ESP and a wet scrubber for sulfur removal was 0.35 ng I-TEQDF/kg coal 

combusted (0.02 ng I-TEQDF/Nm3

 at 11% oxygen). The emission factor reported for a grassdrying chain grate stoker equipped with a cyclone APCD was 1.6 ng I-TEQDF/kg coal fired (0.16 

ng I-TEQDF/Nm3 at 11% oxygen). 

Cains and Dyke (1994) reported an emission factor of 102 to 109 ng I-TEQDF/kg of coal at 

a small-scale facility in the United Kingdom that was equipped with an APCD consisting of only 

a grit arrester. CRE (1994) reported results of testing at 13 commercial and industrial coal-fired 

boilers in the United Kingdom, with TEQ emission factors ranging from 0.04 to 4.8 ng I

TEQDF/kg coal combusted (mean value, 0.6 ng I-TEQDF/kg). CRE also reported testing results for 

one coal-fired power plant: 0.06 ng I-TEQDF/kg of coal combusted. Umweltbundesamt (1996) 

reported that the I-TEQDF content of stack gases from 16 coal-burning facilities in Germany 

4-68
 

­
ranged from 0.0001 to 0.04 ng I-TEQDF/m3

; however, the data provided in that report did not 

enable emission factors to be calculated. 

In 1993, the U.S. Department of Energy (DOE) sponsored a project to assess emissions of 

hazardous air pollutants at coal-fired power plants. As part of the project, CDD/CDF stack 

emissions were measured at seven U.S. coal-fired power plants. The preliminary results of the 

project, concentrations in stack emissions, were reported by Riggs et al. (1995) and are 

summarized in Table 4-23. The levels reported for individual 2,3,7,8-substituted congeners were 

typically very low (less than or equal to 0.033 ng/Nm3) or not detected. In general, CDF levels 

were higher than CDD levels. OCDF and 2,3,7,8-TCDF were the most frequently detected 

congeners (at four of the seven plants). Table 4-24 presents characteristics of the fuel used and 

the APCD employed at each plant. Riggs et al. could not attribute variations in emissions 

between plants to any specific fuel or operational characteristic. 

As mentioned in Section 4.3.2, EPRI sponsored the FCEM project to gather information 

of consistent quality on power plant emissions. Testing for CDD/CDF emissions was performed 

on four coal-fired power plants equipped with cold-sided ESPs. Two plants burned bituminous 

coal and two burned sub-bituminous coal. The results of the testing were integrated into the final 

results of the DOE project discussed above (Riggs et al., 1995) and published in 1994 (EPRI, 

1994). The average congener and congener group emission factors derived from this 11-facility 

data set, as reported in EPRI (1994), are presented in Table 4-25. Congener and congener group 

profiles for the data set are presented in Figure 4-8. Assuming nondetect values were zero, the 

average emission factor was 0.078 ng TEQDF-WHO98/kg (0.079 ng I-TEQDF/kg) coal combusted. 

A medium confidence rating is assigned to the emission factors derived from the DOE and EPRI 

studies because they were based on recent testing at U.S. power plants. 

Because the EPRI and DOE data characterized emissions from units with only cold-sided 

ESPs, there has been uncertainty regarding the applicability of the emission factors derived from 

these data to units with hot-sided ESPs. In July 1999, EPA conducted testing of stack emissions 

at a coal-fired utility equipped with a hot-sided ESP. The preliminary results of this testing 

indicated that the TEQ emission factor for hot-sided ESPs is of the same order of magnitude as 

the average TEQ emission factors derived above. 

4-69
 

Table 4-23. CDD/CDF concentrations (pg/Nm3) in stack emissions from U.S. coal-fired pow

er plan

4-70

Congener/congener group 

Plant 1 

Plant 2 

Plant 3 

Plant 4 

Plant 5 

Pl

2,3,7,8-

T

C

DD 

ND (3.5) ND (3.5) 1 ND (2) ND (3.3) ND (

1,2,3,7,8-PeC

DD 

ND (0.56) ND (4.8) ND (1.8) ND (10) ND (4.7) ND (

1,2,3,4,7,8-

H

x

C

DD 

ND (0.56) ND (5.7) ND (3.6) ND (10)

 N

D (15.4)

 ND (

1,2,3,6,7,8-

H

x

C

DD 

ND (0.44)

5 

ND (1.8) ND (10) ND (9.9) ND (4

1,2,3,7,8,9-

H

x

C

DD 

ND (0.56)

4.9 

ND (1.8) ND (10)

 N

D (12.1)

1,2,3,4,6,7,8-

H

p

C

DD ND (4

ND (1.7)

29 

ND (1.8) ND (10) ND (26.4) 

OCDD ND (12) 32 ND (14) ND (20) N

D (131) 

2,3,7,8-

T

C

DF 

ND (1.7)

8.1 7.8 

ND (2) ND (3.3) 13 

1,2,3,7,8-PeC

DF 

ND (1) ND (5.7)

7.2 

ND (10) ND (3.2)

ND 

2,3,4,7,8-PeC

DF 2.4 ND (19) 6.6 

ND (10) ND (3.2)

ND 

1,2,3,4,7,8-

H

x

C

DF 3.3 16 8.4 

ND (10)

 N

D (16.4) ND 

1,2,3,6,7,8-

H

x

C

DF 1.1 

 ND (5) 2.9 ND (10) ND (5.8) ND 

1,2,3,7,8,9-

H

x

C

DF 

ND (0.44)

11

 ND (1.8) ND (10) ND (8.8)

ND 

2,3,4,6,7,8-

H

x

C

DF 

ND (2) ND (4.2) 

3 

ND (10)

 N

D (16.4) ND 

1,2,3,4,6,7,8-

H

p

C

DF 2 29 6 

 ND (10) ND (23) ND (

1,2,3,4,7,8,9-

H

p

C

DF 

ND (0.63) ND (6.1) ND (3.6) ND (10)

 N

D (15.4) ND 

OCDF 5.6 33 2.4 

 ND (20)

 N

D (131) ND (

Total 2,3,7,8-

C

DD 0 70.9 1 0 0 

Total 2,3,7,8-

C

DF 14.4 97.1 44.3 0 0 

Total T

CDD 1.8 12 12 

NR

6.7

 ND (

Total PeCDD 

ND (1) 

4.4 6 

ND (10) ND (4.7) ND (

Total HxCDD 1.3 18 2.7

 ND (10)

 N

D (26.3)

 ND (4

Total HpCDD 3.4 45 

ND (2.4) ND (10)

 N

D (26.4) ND (

Total OCDD ND (12) 32 

ND (14) ND (20) N

D (131) 

Total T

CDF 

ND (5.2) 

29 78 

ND (2) ND (3.3)

To 

tal PeCDF 5.4 33 61 

ND (10) ND (6.6)

Total HxCDF 7.6 39 29 

ND (10)

 N

D (16.4)

 ND (

Total HpCDF 4.3 34 9 

ND (10)

 N

D (29.5) ND (

Total OCDF 5.6 33 2.4

 ND (20) 

 N

D (131) ND (

Total C

D

D/C

DF 29.4 279.4 200.1 0 6.7 

ND = Not detected; value in parenthesis is the detectio

n limit 

NR = Not reported; s

uspected contamination problem 

Source: Riggs et al. (1995). 
Table 4-24. Characteristics of U.S. coal-fired power plants tested by the U.S. 

Department of Energy 

Plant no. Coal type 

Coal chlorine 

content (mg/kg) 

Temperature (°C)

a 

ESP FF FGD Stack 

1 Bituminous 800 160 – – 160 

2 Bituminous 1,400 130 – – 130 

3 Sub-bituminous 300 – 150 – 150 

4 Sub-bituminous 390 – 70 130 75 

5 Bituminous 1,400 130 – 120 40 

6 Lignite 400 170 – 170 110 

7 Bituminous 1,000 150 – – 150 a

Temperature at pollution control device and stack. 

ESP = Electrostatic precipitator 

FF = Fabric filter 

FGD = Flue gas desulfurization system 

Source: Riggs et al. (1995). 

Applying the TEQ emission factor of 0.078 ng TEQDF-WHO98/kg (0.079 ng I-TEQDF/kg) 

coal combusted to the consumption totals of 651, 771, and 891 million metric tons of coal 

consumed by U.S. utility sectors in 1987, 1995, and 2000, respectively, yields estimated annual 

emissions by the utility sector of 50.89 g TEQDF-WHO98 (51.4 g I-TEQDF) in 1987, 60.1 g 

TEQDFWHO98 (60.9 g I-TEQDF) in 1995, and 69.5 g TEQDF-WHO98 (70.4 g I-TEQDF) in 2000. 

These emission estimates are assigned a medium confidence rating because the emission factor 

rating for this category was judged to be medium. 

No testing results could be located for CDD/CDF content in air emissions from 

commercial and industrial coal-fired combustion units in the United States. It is uncertain 

whether the data collected in the European studies (Bremmer et al., 1994; CRE, 1994) accurately 

represent U.S. sources, but the data suggest that emission factors for commercial/industrial sources 

might be higher than those reported for U.S. coal-fired utilities. Therefore, no national emission 

estimate has been derived for this category. However, preliminary estimates of potential national 

TEQ emissions from this source category can be derived for 1987, 1995, and 2000 using the total 

coal consumption for each of those reference years, excluding consumption by coke plants, and 

the average emission factor, 0.6 ng I-TEQDF/kg coal combusted. 

4-71
 
Table 4-25. CDD/CDF emission factors (ng/kg coal) for coal-fired utility/industrial 

power plantsa 

Congener/congener group Nondetect set to zero 

Nondetect set to 1⁄2 

detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.005 

0 

0 

0.004 

0.004 

0.216 

0.513 

0.018 

0.016 

0.034 

0.028 

0.035 

0.241 

0.644 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.109 

0.007 

0.074 

0.098 

0.014 

0.013 

0.043 

0.354 

0.087 

0.158 

0.117 

0.021 

0.084 

0.12 

0.03 

0.038 

0.06 

0.385 

0.112 

0.281 

Total I-TEQDF 

Total TEQDF-WHO98 

0.078 

0.078 

0.124 

0.131 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

0.051 

0.014 

0.03 

0.063 

0.513 

0.154 

0.18 

0.104 

0.064 

0.158 

0.052 

0.015 

0.03 

0.074 

0.644 

0.158 

0.18 

0.104 

0.064 

0.281 

Total CDD/CDF 1.331 1.602 a

Eleven-facility data set. 

Source: EPRI (1994). 

4-72
 

Figure 4-8. Congener and congener group profiles for air emissions from 

industrial/utility coal-fueled combustors (nondetects set equal to zero). 

Source: EPRI (1994). 

4-73 
Applying the emission factor to the estimated combustion for 1987, 1995, and 2000 (68.2, 

66, and 59.14 million metric tons, respectively) yields 40.9 g I-TEQDF/kg for 1987, 39.6 g ITEQDF/kg for 1995, and 35.4 g I-TEQDF/kg coal combusted for 2000. These estimates should be 

regarded as preliminary indications of possible emissions from commercial/industrial coal-fired 

boilers; further testing is needed to confirm the true magnitude of these emissions. These 

emission estimates are assigned a Category D uncertainty rating because the emission factor has 

been judged to be clearly nonrepresentative of commercial and industrial coal-fired combustion 

units. 

4.4.2. Residential Coal Combustion 

In the residential sector, coal is usually combusted in underfed or hand-stoked furnaces. 

Other coal-fired heating units include hand-fed room heaters, metal stoves, and metal and 

masonry fireplaces. Stoker-fed units are the most common design for warm-air furnaces and for 

boilers used for steam or hot water production. Most coal combusted in these units is either 

bituminous or anthracite. These units operate at relatively low temperatures and do not efficiently 

combust the coal. Coal generally contains small quantities of chlorine; therefore, the potential for 

CDD/CDF formation exists. Typically, coal-fired residential furnaces are not equipped with PM 

or gaseous pollutant control devices that may limit emissions of any CDDs/CDFs formed (U.S. 

EPA, 1997b). No testing results for CDD/CDF content in air emissions from 

residential/commercial coal-fired combustion units in the United States could be located; 

however, several relevant studies have been performed in European countries. 

Thub et al. (1995) measured flue gas concentrations of CDDs/CDFs from a household 

heating system in Germany that was fired with either salt lignite coal (total chlorine content of 

2,000 ppm) or normal lignite coal (total chlorine content of 300 ppm). CDDs/CDFs were detected 

in the flue gases generated by combustion of both fuel types (see Table 4-26). The congener 

profiles and patterns were similar for both fuel types, with OCDD the dominant congener and 

TCDF the dominant congener group. However, the emissions were higher by a factor of 8 for the 

“salt” coal (0.109 ng I-TEQDF/m3

 [2.74 ng I-TEQDF/kg]) than for the “normal” coal (0.015 ng ITEQDF/m3

 [0.34 ng I-TEQDF/kg]). 

Using the results of testing performed by the Coal Research Establishment in the United 

Kingdom, Eduljee and Dyke (1996) estimated emission factors for residential coal combustion 

units of 2.1 ng I-TEQDF/kg for anthracite coal and 5.7 to 9.3 ng I-TEQDF/kg (midpoint, 7.5 ng 

I-TEQDF/kg) for bituminous coal. 

4-74
 

Table 4-26. CDD/CDF emission factors (ng/kg coal) for residential coal 

combustors 

Congener 

“Salt” 

lignitea 

“Normal” 

lignitea Anthraciteb Bituminousb 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.58 

0.73 

0.63 

0.6 

0.4 

3.24 

16.19 

0.06 

0.08 

0.06 

0.09 

0.06 

0.59 

2.42 

1.6 

NR 

NR 

NR 

NR 

NR 

77 

2.4 

NR 

NR 

NR 

NR 

NR 

120 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

2.49 

2.24 

2.09 

0.38 

1.86 

0.07 

1.01 

2.59 

0.25 

0.63 

0.5 

0.43 

0.31 

0.13 

0.36 

0.02 

0.12 

0.95 

0.06 

0.3 

42 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

4.2 

63 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

6.3 

Total 2,3,7,8-CDDc 

Total 2,3,7,8-CDFc 

Total I-TEQDF 

c 

22.37 

13.6 

2.74 

3.38 

3.2 

0.34 

NR 

NR 

60 

NR 

NR 

98.5 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

14.23 

14.15 

11.14 

7.06 

16.19 

80.34 

29.21 

12.72 

3.87 

0.63 

9 

2.22 

1.81 

0.82 

2.42 

20.33 

8.98 

3.78 

1.27 

0.3 

61.6 

31 

60 

57 

77 

412 

340 

130 

32 

4.2 

92.4 

46 

90 

86 

120 

613 

550 

190 

47 

6.3 

Total CDD/CDF 189.5 50.9 1,205 1,841 

a

Source: Thub et al. (1995); listed results represent means of three flue gas samples. b

Source: U.S. EPA (1997a); based on average particulate CDD/CDF concentrations from chimney soot samples

 collected from seven coal ovens and particulate emission factors for anthracite and bituminous coal combustion. c

Values as reported in sources. 

NR = Not reported 

4-75
 
CDD/CDF emission factors were estimated (U.S. EPA, 1997a) for coal-fired residential 

furnaces using average particulate CDD/CDF concentrations from chimney soot samples collected 

from seven coal ovens and PM emission factors obtained from AP-42 that are specific to 

anthracite and bituminous coal combustion (U.S. EPA, 1995a). The TEQ emission factors 

estimated (60 and 98.5 ng I-TEQDF/kg anthracite and bituminous coal, respectively) were derived 

using the calculated emission factors for 2,3,7,8-TCDD, 2,3,7,8-TCDF, and the 10 congener 

groups (U.S. EPA, 1997a). EPA stated that the estimated factors should be considered 

representative of maximum emission factors because soot may not be representative of the PM 

actually emitted to the atmosphere. These emission factors are presented in Table 4-26; congener 

group profiles are presented in Figure 4-9. 

Although the congener group profiles of the measurements by Thub et al. (1995) and the 

estimates by EPA (U.S. EPA, 1997a) are similar, the TEQ emission factors of the two studies 

differ by factors of 175 to 289. The emission factors used by Eduljee and Dyke (1996) to estimate 

national annual TEQ emissions from residential coal combustion in the United Kingdom fall in 

between the other two sets of estimates but are still about one to two orders of magnitude greater 

than the estimated emission factor for industrial and utility coal combustors (see Section 4.4.1). 

For 1987 and 1995, preliminary estimates of potential national TEQs were derived using 

the emission factors from Eduljee and Dyke (1996). U.S. EPA (1997a) reported that 72.5% of the 

coal consumed by the residential sector in 1990 was bituminous and 27.5% was anthracite. 

Assuming that these relative proportions reflect the actual usage in 1987 and 1995, then 

application of the emission factors from Eduljee and Dyke (2.1 ng I-TEQDF/kg anthracite coal and 

7.5 ng I-TEQDF bituminous coal) to the consumption values of 6.3 and 5.3 million metric tons in 

1987 and 1995, respectively (U.S. DOC, 1997), results in estimated TEQ emissions of 37.9 and 

32.0 g I-TEQDF in 1987 and 1995, respectively. These estimates should be regarded as 

preliminary indications of possible emissions from this source category because the emission 

factor is judged to be clearly nonrepresentative of the sources. Further testing is needed to 

confirm the true magnitude of these emissions. 

For 2000, OAQPS developed national emission estimates for coal combustion for 

residential heating. The activity level for residential coal combustion was taken from state-level 

2000 activity data (EIA, 2003a). Because EIA no longer disaggregates coal consumption into 

anthracite versus bituminous/lignite, OAQPS estimated each state’s coal consumption using the 

state’s 1999 proportion of anthracite versus bituminous/lignite to total coal consumption. 

Emissions were allocated to the county level as a proportion of state population in states that 

consume anthracite coal and bituminous and lignite coal for residential heating. OAQPS 

estimated that in 2000, 67,400 metric tons of anthracite coal and 343,000 metric tons of 

4-76
 

Figure 4-9. Congener group profile for air emissions from residential 

coal-fueled combustors. 

Source: U.S. EPA (1997a). 

bituminous and lignite coal were consumed for residential heating. Applying the TEQ emission 

factors of 2.1 ng I-TEQDF/kg anthracite coal combusted and 7.5 ng I-TEQDF/kg bituminous coal 

combusted (Eduljee and Dyke, 1996) to these production factors yields preliminary estimates of 

annual emissions of 0.14 g I-TEQDF of anthracite coal and 2.6 g I-TEQDF of bituminous/subbituminous coal in 2000. These emission estimates are assigned a Category D uncertainty rating 

because the emission factor has been judged to be clearly nonrepresentative of residential coal 

combustion. 

4.4.3. Solid Wastes from Coal Combustion 

A limited amount of CDD/CDF concentration data have been developed for utility 

industry solid wastes (U.S. EPA, 1999c), and these data are for wastes that are comanaged (i.e., 

combinations of fly ash, bottom ash, boiler slag, and flue gas desulfurization wastes). A total of 

15 samples were taken from 11 disposal sites. The average concentration for each of the CDD 

4-77
 

and CDF congeners is presented in the second column of Table 4-27. It should be noted that most 

of the concentration values shown in Table 4-27 represent DLs. Consequently, the values 

overestimate actual concentrations. 

U.S. EPA (1999d) indicates that approximately 63 million tons (assumed to be short tons, 

i.e., 2,000 pounds) of large-volume utility coal combustion solid wastes were produced in 1995. 

Of this amount, about 67% was landfilled and the balance was disposed of in surface 

impoundments. The concentration data presented in Table 4-27 are for only the 53 million tons 

that were comanaged (about 84% of the total wastes). For purposes of this analysis it is assumed 

that the CDD/CDF concentrations in the comanaged wastes are the same as for the entire waste 

quantity. Combining the concentration data with the 63 million tons of total waste yields the total 

quantities of each congener disposed of in 1995. These data are presented in the fourth column of 

Table 4-27. As indicated in Section 4.4 of this document, total consumption of coal for electric 

utility boilers in 1987 was 98.4% of 1995 consumption. Consequently, the quantities of 

CDDs/CDFs disposed of in 1987 is assumed to be 98.4% of the 1995 values. These values are 

presented in column 3 of Table 4-27. The 1995 congener quantities are converted into I-TEQDF 

and TEQDF-WHO98 values in columns 5 and 6. Because disposal of these wastes does not 

constitute an environmental release, the values are not included in the inventory. 

4-78
 

Table 4-27. CDD/CDF concentrations and amounts disposed of in solid wastes 

from coal-fired utilities 

Congener 

Mean 

concentrationa 

(ng/kg) 

Disposed of 

in solid 

waste in 

1987 (g/yr)

b 

Disposed of 

in solid waste 

in 1995 (g/yr)

c 

I-TEQDF 

/yr 1995 

(g) 

TEQDF ­

WHO98/yr 

1995 (g) 

2,3,7,8-TCDDd 0.17 10 10 9.72 9.72 

1,2,3,7,8-PeCDDd 0.25 14 14 7.15 14.3 

1,2,3,4,7,8-HxCDDd 0.35 20 20 2 2 

1,2,3,6,7,8-HxCDDd 0.28 16 16 1.6 1.6 

1,2,3,7,8,9-HxCDDe 0.3 17 17 1.72 1.72 

1,2,3,4,6,7,8-HpCDDf 0.59 33 34 0.34 0.34 

OCDDg 10.54 593 603 0.6 0.6 

2,3,7,8-TCDFh 0.19 11 11 1.09 1.09 

1,2,3,7,8-PeCDFd 0.17 10 10 0.49 0.49 

2,3,4,7,8-PeCDFd 0.17 10 10 4.86 4.86 

1,2,3,4,7,8-HxCDFe 0.25 14 14 1.43 1.43 

1,2,3,6,7,8-HxCDFd 0.18 10 10 1.03 1.03 

2,3,4,6,7,8-HxCDFd 0.28 16 16 1.6 1.6 

1,2,3,7,8,9-HxCDFd 0.24 14 14 1.37 1.37 

1,2,3,4,6,7,8-HpCDFe 0.29 16 17 0.17 0.17 

1,2,3,4,7,8,9-HpCDFd 0.35 20 20 0.2 0.2 

OCDFi 0.59 33 34 0.03 <0.01 

TOTAL 35.41 41.98 a

Source: U.S. EPA (1999c, Table 2-9). b

Assumes that solid waste quantity for 1987 was 98.4% of 1995 quantity, based on total utility coal use in those years

 (see Section 4.4). c

Based on EPRI estimate of 63 million tons/yr of large-volume utility coal combustion solid wastes. See Section

 3.3 of U.S. EPA (1999d). Assumes all waste characteristics were the same as for comanaged wastes. d

All 17 analyses were nondetects. e

Sixteen of the 17 analyses were nondetects. f

Eleven of the 17 analyses were nondetects. g

Five of the 17 analyses were nondetects. h

Fourteen of the 17 analyses were nondetects. i

Fifteen of the 17 analyses were nondetects. 

4-79
 

5. COMBUSTION SOURCES OF CDDs/CDFs: OTHER 

HIGH-TEMPERATURE SOURCES 

5.1. CEMENT KILNS 

This section addresses CDD/CDF emissions from Portland cement kilns. These facilities 

use high temperatures to convert mineral feedstocks into Portland cement and other types of 

construction materials. For purposes of this analysis, cement kilns are divided into two 

categories: those that burn hazardous waste as a supplementary fuel and those that do not. For 

the 1987 and 1995 estimates, the hazardous waste-burning cement kiln category was further 

divided into kilns with inlet air pollution control device (APCD) temperatures above 232°C and 

those with APCD temperatures below 232°C. Cement kiln technology, the derivation of TEQ 

emission factors for cement kilns that burn hazardous waste as supplemental fuel and those that 

do not, and the derivation of annual TEQ air emissions (g/yr) for reference years 1987, 1995, and 

2000 are discussed below. 

5.1.1. Process Description of Portland Cement Kilns

 In the United States, the primary cement product is Portland cement. Portland cement is 

a fine, gray powder consisting of a mixture of four basic materials: lime, silica, alumina, and iron 

compounds. Cement production involves heating (pyroprocessing) the raw materials to a very 

high temperature in a rotary (rotating) kiln to induce chemical reactions that produce a fused 

material called clinker. The cement clinker is then ground into a fine powder and mixed with 

gypsum to form the Portland cement. 

The cement kiln is a large, steel, rotating cylindrical furnace lined with refractory 

material. The kiln is aligned on a slight angle, usually a slope of 3 to 6 degrees, which allows the 

materials to pass through the kiln by gravity. The kiln rotates at about 50 to 70 revs/hr, and the 

rotation induces mixing and the downward movement of mixed materials. The upper end of the 

kiln, known as the cold end, is generally where the raw materials, or meal, are fed into the kiln. 

Midpoint injection is practiced at some facilities. The lower end of the kiln, known as the hot 

end, is where the combustion of primary fuels (usually coal and petroleum coke) occurs, 

producing a high temperature. The cement kiln operates using countercurrents: hot combustion 

gases are convected up through the kiln while the raw materials pass down toward the lower end. 

As the meal moves through the cement kiln and is heated by hot combustion gases, water is 

vaporized and pyroprocessing of materials occurs. 

The cement kiln consists of three thermal zones to produce cement clinker. Zone 1 is at 

the upper end of the kiln where the raw meal is added. Temperatures in this zone typically range 

from ambient up to 600°C. In this area of the kiln, moisture is evaporated from the raw meal. 

5-1
 
Zone 2 is known as the calcining zone. Calcining occurs when the hot gases from the 

combustion of primary fuels dissociate calcium dioxide from the limestone and form calcium 

oxide. In this zone, temperatures range from 600 to 900°C. Zone 3, the burning or sintering 

zone, is the lowest and hottest region of the kiln. Here, temperatures in excess of 1,500°C induce 

the calcium oxide to react with silicates, iron, and aluminum in the raw materials to form the 

cement clinker. The formation of clinker actually occurs close to the combustion of primary fuel. 

The chemical reactions that occur in zone 3 are referred to as pyroprocessing. 

The cement clinker, which leaves the kiln at the hot end, is a gray, glass-hard material 

consisting of dicalcium silicate, tricalcium silicate, calcium aluminate, and tetracalcium 

aluminoferrite. At this point, the temperature of the clinker is about 1,100°C. The hot clinker is 

then dumped onto a moving grate, where it cools as it passes under a series of cool-air blowers. 

After it is cooled to ambient temperature, the clinker is ground into a fine powder and mixed with 

gypsum to produce the Portland cement product. 

Cement kilns can be either wet process or dry process. In the wet process, the raw 

materials are ground and mixed with water to form a slurry, which is fed into the kiln through a 

pump. This is an older process. A greater amount of heat energy is needed in the wet process 

kiln than in other types of kilns. These kilns consume about 5 to 7 trillion Btu per ton of clinker 

product to evaporate the additional water. In the dry process, a preheater is used to dry the raw 

meal before it enters the kiln. A typical preheater consists of a vertical tower containing a series 

of cyclone-type vessels. Raw meal is added at the top of the tower and hot exhaust gases from 

the kiln operation preheat the meal, thus lowering the fuel consumption of the kiln. Dry kilns are 

now the most popular type of cement kiln. 

Portland cement clinker production in the United States is estimated to have been 52 

billion kg in 1987 (U.S. DOC, 1996), 67.6 billion kg in 1995 (U.S. DOC, 1996), and 75.2 billion 

kg in 2000 (PCA, 2001). The 2000 estimate is based on the assumption that of the annual 

maximum clinker capacity reported for that reference year (PCA, 2001), only 90% was actually 

produced (e-mail dated January 31, 2003, from Garth Hawkins, Portland Cement Association, to 

Karie Riley, Versar, Inc.). 

5.1.2. Cement Kilns That Burn Hazardous Waste 

The high temperatures achieved in cement kilns make the kilns an attractive technology 

for combusting hazardous waste as supplemental fuel. Sustaining the relatively high combustion 

temperatures that are needed to form cement clinker (1,100 to 1,500°C) requires the burning of a 

fuel with a high energy output. Therefore, coal or petroleum coke is typically used as the primary 

fuel source. Because much of the cost of operating the cement kiln at high temperatures is 

associated with the consumption of fossil fuels, some cement kiln operators burn hazardous 

5-2
 
liquid and solid waste as supplemental fuel. In 2000, approximately 60% of all facilities burned 

hazardous waste as the primary fuel to offset the amount of coal/coke purchased and burned by 

the kiln (PCA, 2001). Organic hazardous waste may have a heating value similar to that of coal 

(9,000 to 12,000 Btu/lb for coal). The kiln operator may charge the waste generator a disposal 

fee to combust the hazardous waste; this fee also offsets the cost of kiln operation. The highenergy and ignitable wastes include diverse substances, such as waste oils, spent organic 

solvents, sludges from the paint and coatings industry, waste paints and coatings from auto and 

truck assembly plants, and sludges from the petroleum refining industry (Greer et al., 1992). 

The conditions in the cement kiln mimic conditions of hazardous waste incineration. For 

example, the gas residence time in the burning zone is typically 3 sec at temperatures in excess of 

1,500°C (Greer et al., 1992). The method of introducing liquid and solid hazardous waste into 

the kiln is a key factor in the complete consumption of the waste during the combustion of the 

primary fuel. Liquid hazardous waste is either injected separately or blended with the primary 

fuel (coal). Solid waste is mixed and burned along with the primary fuel. 

Trial burns have consistently shown that destruction and removal efficiencies of 99.99 to 

99.9999% can be achieved for very stable organic wastes using cement kilns (Greer et al., 1992). 

Hazardous waste was combusted at 34 of the 212 kilns operating in 1995 (Federal Register, 

1996a) and at 33 of the 201 kilns operating in 2000 (e-mail correspondence dated February 24, 

2003, between M. Benoit, Cement Kiln Recycling Coalition, and K. Riley, Versar, Inc.; PCA, 

2001). Other types of supplemental fuel used by these facilities include natural gas, fuel oil, 

automobile tires, used motor oil, sawdust, and scrap wood chips. 

5.1.3. Air Pollution Control Devices 

The pyroprocessing of raw meal in a cement kiln also produces fine particulates, referred 

to as cement kiln dust (CKD). CKD is collected and controlled with fabric filters (FFs), 

electrostatic precipitators (ESPs), or both. Acid gases such as sulfur dioxide (SO2) can be formed 

during pyroprocessing of the sulfur-laden minerals and fuels, but the minerals have high 

alkalinity, which partially neutralizes SO2 gases. Most APCDs used at cement kilns in 1987 and 

1995 were considered to be hot-sided control devices. A hot-sided control device is one that 

operates at kiln exhaust gas temperatures above 232°C (some EPA rules use different definitions 

for hot-sided control devices for different industries). Most APCDs currently used at cement 

kilns are cold-sided devices (i.e., they operate at kiln exhaust gas temperatures below 232°C. 

Reducing the temperature at the inlet of the APCD is one factor that has been shown to 

have a significant impact on limiting dioxin formation and emissions at cement kilns (U.S. EPA, 

1997c). Emissions testing at a Portland cement kiln showed that CDDs/CDFs were almost 

entirely absent at the inlet to a hot-sided ESP, but measurements taken at the exit showed 

5-3
 
conclusively that dioxins were formed within the hot-sided ESP (U.S. EPA, 1997c). Reducing 

the kiln exhaust gas temperature in the APCD to below 232°C has been shown to substantially 

limit CDD/CDF formation. Lower temperatures are believed to prevent the post-combustion 

catalytic formation of CDDs/CDFs. Consequently, a number of cement kilns have added exhaust 

gas-quenching units upstream of the APCD to reduce the inlet APCD temperature, thereby 

reducing CDD/CDF stack concentrations. A quenching unit usually consists of a water spray 

system within the flue duct. 

5.1.4. CDD/CDF Emissions Data 

The general strategy used to derive emission factors for this report was to divide each 

source category on the basis of design and operation. However, because cement kilns are 

relatively uniform in terms of kiln design, raw feed material, operating temperatures, and 

APCDs, they have been categorized, as noted above, only on the basis of whether or not 

hazardous waste is burned as a supplementary fuel. 

CDD/CDF emissions data from tests conducted between 1989 and 1996 were obtained 

for 16 cement kilns burning hazardous waste and 15 cement kilns burning nonhazardous waste 

(U.S. EPA, 1996b). More recent CDD/CDF emissions data were also obtained from tests 

conducted in 2000 at 3 cement kilns burning hazardous waste (U.S. EPA, 2002a) and from tests 

conducted in June and July of 1999 at one facility burning nonhazardous waste (Bell, 1999). The 

majority of stack emissions data from cement kilns burning hazardous waste were derived during 

trial burns and may overestimate the CDD/CDF emissions that most kilns achieve during normal 

operations. Stack emissions data from kilns burning nonhazardous waste were derived from 

testing during normal operations. 

5.1.4.1. Emissions Data for 1989 Through 1996 (U.S. EPA, 1996b) 

The average TEQ emission factors for this period, based on the data reported by EPA in 

1996 (U.S. EPA, 1996b), were 0.000941 to 232 ng TEQDF-WHO98/kg (average of 22.48 ng 

TEQDF-WHO98/kg [20.91 ng I-TEQDF/kg]) clinker produced for cement kilns burning hazardous 

waste and 0.000012 to 2.76 ng TEQDF-WHO98/kg (average of 0.29 ng TEQDF-WHO98/kg [0.27 ng 

I-TEQDF/kg]) clinker produced for cement kilns burning nonhazardous waste. 

These data show that the average emission factor for kilns burning hazardous waste was 

about 90 times greater than that for kilns burning nonhazardous waste. However, it should be 

noted that the average emission factor for kilns burning hazardous waste was derived from “near 

worst case” testing of hazardous waste-burning kilns. As discussed in Section 5.1.8, a 

comparison of CDD/CDF concentrations in CKD samples shows a similar relationship (i.e., the 

CDD/CDF TEQ concentration of the CKD from kilns burning hazardous waste was about 100 

5-4
 
times higher than that of the dust from kilns burning nonhazardous waste). Although the average 

emission factors for the two groups of kilns differ substantially, the emission factors for 

individual kilns in the two groups overlap. Therefore, other aspects of the design and operation 

of the kilns—in particular, the temperature of the APCD equipment (as discussed in Section 

5.1.3)—were likely affecting CDD/CDF emissions. 

Previous attempts to understand these differences using parametric testing of cement 

kilns yielded mixed results. EPA (U.S. EPA, 1997c) conducted a limited comparison of 

CDD/CDF TEQ stack gas concentrations (ng TEQ/dscm) between cement kilns burning 

hazardous wastes and those not burning hazardous wastes. Those comparisons were made at 14 

cement kilns. With the exception of the fuel being burned, operating conditions (e.g., APCD 

temperature) were the same or similar for each set of comparisons. Baseline conditions used coal 

as the only primary fuel. The results of these comparisons found: 

•	 seven kilns in which the baseline (i.e., no combustion of hazardous waste) CDD/CDF 

TEQ stack gas concentrations were about the same as those for the burning of 

hazardous wastes, 

•	 two kilns in which the baseline CDD/CDF I-TEQDF stack gas concentrations were 

about double those for the burning of hazardous wastes, and 

•	 five kilns in which the hazardous waste CDD/CDF I-TEQDF stack gas concentrations 

were substantially greater (3- to 29-fold greater) than those for the baseline operating 

conditions. 

Subsequently, EPA’s Office of Research and Development (ORD) conducted analyses of 

the available emissions data to evaluate, on a congener-specific basis, whether there were 

significant differences in emission factors between (a) kilns burning hazardous waste and those 

burning nonhazardous waste, (b) kilns with APCD inlet temperatures greater than 232°C and 

those with temperatures less than 232°C, (c) hazardous waste-burning and nonhazardous wasteburning facilities with APCD inlet temperatures greater than 232°C, (d) hazardous waste-burning 

and nonhazardous waste-burning facilities with APCD inlet temperatures less than 232°C, (e) 

hazardous waste-burning facilities with APCD inlet temperatures less than and greater than 

232°C, and (f) nonhazardous waste-burning facilities with APCD inlet temperatures less than and 

greater than 232°C. The results of all analyses showed significant differences in the sample 

mean values (p<0.05). 

Given the strong empirical evidence that real differences existed, ORD decided to address 

the kilns burning hazardous waste separately from those burning nonhazardous waste to develop 

a CDD/CDF emissions inventory and to subdivide the hazardous waste-burning category into 

5-5
 
subcategories by APCD inlet temperature (i.e., less than 232°C and greater than 232°C). APCD 

inlet temperature data were available for 88 test runs at 14 cement kilns. The number of test runs 

conducted at individual kilns ranged from 1 to 26. Each test run was treated as an individual 

facility and each was classified according to APCD inlet temperature and whether or not 

hazardous waste was burned. The emission factor for each cement kiln test run was calculated 

using eq 5-1. 

= C × Fv EF (5-1) CK ______ 

Icl 

where: 

EFCK = cement kiln emission factor (burning or not burning hazardous waste) 

 (ng TEQ/kg of clinker produced) 

C = TEQ or CDD/CDF concentration in kiln exhaust gases (ng TEQ/dscm)

 (20°C, 1 atm; adjusted to 7% O2) 

Fv = volumetric kiln exhaust gas flow rate (dscm/hr) (20°C, 1 atm; adjusted to 7% O2) 

Icl = average cement kiln clinker production rate (kg/hr) 

After developing the emission factor for each cement kiln test run, the overall average congenerspecific emission factor was derived for all test runs in each subcategory using eq 5-2. 

EFCK1 + EFCK2 + EFCK3 + ....... + EFCKN 

EFavgCK 

_______________________________ (5-2)

= N 

where: 

EFavgCK = average emission factor of tested cement kilns burning hazardous 

 waste as supplemental fuel and with APCD inlet temperatures 

 either greater than or less than 232°C (ng TEQ/kg clinker) 

N = number of cement kiln test runs 

TEQ emission values for hazardous waste-burning cement kilns with APCD inlet temperatures 

greater than 232°C and less than 232°C were 30.7 and 1.11 ng TEQDF-WHO98/kg clinker 

produced, respectively. 

5-6
 
5.1.4.2. Emissions Data for 1999 and 2000 (Bell, 1999; U.S. EPA, 2002a) 

The results of a test conducted in 1999 for a cement kiln burning nonhazardous waste 

(Bell, 1999) showed average TEQDF-WHO98 and I-TEQ emission factors of 0.14 ng/kg clinker 

produced. This value is within the range of emission factor values developed using the 1989 

through 1996 data (U.S. EPA, 1996b). 

In 1999, EPA’s Office of Solid Waste and Emergency Response (OSWER), under the 

National Emission Standards for Hazardous Air Pollutants (NESHAPs), promulgated Final 

Standards for Hazardous Air Pollutants for Hazardous Waste Combustors (Federal Register, 

1999a). During the development of this rulemaking, a database of national emission estimates 

for hazardous waste combustors (HWCs), including hazardous waste-burning cement kilns, was 

established. In this database EPA identified 30 hazardous waste cement kilns operating in 1997. 

Twenty-four of the 30 kilns had provided emissions sampling data to EPA, including dioxin 

concentrations. Using the most recent sampling data for each kiln, along with imputed release 

estimates for the missing six kilns and the assumption that cement kilns were operating 24 

hr/day, 365 days/yr, OSWER developed a national emissions estimate of 13 g TEQ/yr for 1997. 

A number of parties sought judicial review of the 1999 rule, and in 2001 the standards 

were vacated. Interim standards were issued in 2002 until EPA could issue revised final 

standards in 2004. The existing national emissions database was updated to include new 

sampling data and data for newly operational HWCs and to remove eight of the cement kilns that 

were identified as no longer burning hazardous waste. From this updated database (U.S. EPA, 

2002a), EPA was able to develop new emission estimates for 2000. 

5.1.4.3. Emission Factor Estimates for Cement Kilns Burning Hazardous Waste 

For reference years 1987 and 1995, EPA estimated the TEQ emission factor by 

subdividing the emissions data reported in 1996 (i.e., for 1989 through 1996) (U.S. EPA, 1996b) 

by inlet APCD temperature above and below 232°C. For cement kilns operating at temperatures 

above 232°C, the TEQ emission factor was 30.7 ng TEQDF-WHO98/kg clinker produced, and for 

cement kilns operating at temperatures below 232°C, the TEQ emission factor was 1.11 ng 

TEQDF-WHO98/kg clinker produced. These emission factors are presented in Table 5-1 and the 

average congener profile is presented in Figure 5-1. 

 Because a vast majority of the facilities had reduced their APCD inlet temperature to 

below 232°C by 2000, and because only a few new test reports applicable to reference year 2000 

were available, EPA removed the 232°C divider and combined the emission factor results (i.e., 

1989 through 1996 data) (U.S. EPA, 1996b) for facilities that were still operating in 2000 with 

5-7
 
Table 5-1. CDD/CDF emission factors (ng/kg clinker produced) for cement 

kilns burning hazardous waste for reference years 1987 and 1995 

Congener/ 

congener group 

Mean emission factor 

(nondetect values set equal to zero) 

APCD inlet temperature 

>232°C 

APCD inlet temperature 

<232°C 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

3.38 

4.28 

4.85 

6.93 

9.55 

27.05 

18.61 

0.02 

0.13 

0.29 

0.42 

0.4 

3.16 

1.08 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

36.26 

13.36 

23.48 

22.24 

8.46 

0.96 

13.33 

7.73 

2.16 

2.51 

3.24 

0.23 

0.65 

0.55 

0.27 

0.06 

0.52 

0.34 

0.16 

0.37 

Total I-TEQDF 

Total TEQDF-WHO98 

28.58 

30.7 

1.04 

1.10 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

406.76 

608.65 

845.99 

192.99 

18.61 

295.72 

127.99 

50.75 

8.36 

2.51 

1.78 

0.89 

0.69 

0.42 

1.08 

11.52 

3.83 

1.88 

0.47 

0.37 

Total CDD/CDF 2,558.33 22.93

 APCD = Air pollution control device

 Source: U.S. EPA (1996b). 

the newer data reported (U.S. EPA, 2002a). Therefore, emission tests from five facilities (U.S. 

EPA, 1996b) were not used to estimate the 2000 emission factor because the facilities no longer 

burned hazardous waste in 2000. Using this approach, a conservative TEQ emission estimate of 

5-8
 
Ratio (congener emission factor/total CDD/CDF emission factor) 

0a 0.005 0.01 0.015 0.02 

Figure 5-1. Congener profile for air emissions from cement kilns burning 

hazardous waste for reference years 1987 and 1995 (nondetect set equal to 

zero). 

5.95 ng TEQDF-WHO98/kg (5.49 ng I-TEQ/kg) clinker produced was developed for reference year 

2000. The congener-specific emission factors are presented in Table 5-2 and the average 

congener and congener group profiles are presented in Figure 5-2. 

5.1.4.4. Emission Factor Estimates for Cement Kilns Burning Nonhazardous Waste 

Because only one test report applicable to reference year 2000 was located for a cement 

kiln burning nonhazardous waste (Bell, 1999), and the results from the tests were similar to the 

results reported by EPA in 1996 (U.S. EPA, 1996b), EPA combined the results from the two data 

sets to obtain a TEQ emission factor estimate of 0.27 ng TEQDF-WHO98/kg (0.26 ng I-TEQ/kg) 

clinker produced for reference years 1987, 1995, and 2000. The congener-specific emission 

factors are presented in Table 5-3 and the average congener and congener group profiles are 

presented in Figure 5-3. 

5-9
 
Table 5-2. CDD/CDF emission factors (ng/hr) for cement kilns burning 

hazardous waste for reference year 2000 

Congener/congener group Mean emission factor 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

9,533 

32,510 

48,483 

78,043 

66,369 

446,105 

142,108 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

63,866 

37,470 

77,981 

59,340 

48,179 

33,967 

47,458 

47,762 

15,098 

12,418 

Total I-TEQDF (g/yr) 

Total TEQDF-WHO98 (g/yr) 

16.6 

18.8 

Sources: U.S. EPA (1996b, 2002a). 

5.1.4.5. Confidence Ratings of Emission Factor Estimates 

The TEQ emission factors are given a low confidence rating for all subcategories and all 

years. The emission factor for nonhazardous waste-burning kilns was given a low rating because 

test data were available for only 16 facilities. The tested facilities may not be representative of 

routine CDD/CDF emissions from all kilns burning nonhazardous waste. Although a higher 

percentage of the kilns burning hazardous waste (with reported APCD temperature data) had 

been tested, greater uncertainty exists about whether the emissions are representative of normal 

operations because the tests used trial burn procedures and because a greater majority of the 

operating facilities had reduced their APCD temperatures to below 232°C. Accordingly, a low 

confidence rating is also assigned to the estimated emission factors for kilns burning hazardous 

waste. 

5.1.5. Activity Level Information 

In 1987, approximately 52 billion kg of cement clinker were produced in the United 

States (U.S. DOC, 1996). In 1995, approximately 67.6 billion kg of clinker were produced in the 

5-10
 
0.000 

4D 2378 

5D 12378 

6D 123478 

6D 123678 

6D 123789 

7D 1234678 

8D 

4F 2378 

5F 12378 

5F 23478 

6F 123478 

6F 123678 

6F 123789 

6F 234678 

7F 1234678 

7F 1234789 

8F 

Ratio (congener emission factor / 2378-CDD/CDF emission factor) 

0.010 0.020 0.030 0.040 0.050 0.060 

R a t io ( c o n g e n e r gr o u p e m is sio n f a c t o r /t o t a l C D D /C D F e m iss io n f a c t o r ) 

0 0 .1 0 . 2 0 .3 0 . 4 0 .5 

4 D T ot a l 

5 D T ot a l 

6 D T ot a l 

7 D T ot a l 

8D 

4 F T ot a l 

5 F T ot a l 

6 F T ot a l 

7 F T ot a l 

8F 

Figure 5-2. Congener profile for air emissions from cement kilns burning 

hazardous waste for reference year 2000. 

5-11
 
Table 5-3. CDD/CDF emission factors (ng/kg clinker produced) for cement kilns not 

burning hazardous waste for reference years 1987, 1995, and 2000a 

Congener/congener group 

Mean emission factor 

Nondetect set to zero 

Nondetect set to 1⁄2 detection 

limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.01 

0.03 

0.03 

0.04 

0.04 

0.39 

0.64 

0.02 

0.04 

0.04 

0.05 

0.06 

0.39 

0.64 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.73 

0.1 

0.22 

0.17 

0.05 

0.01 

0.08 

0.13 

0 

0.22 

0.73 

0.11 

0.23 

0.18 

0.06 

0.02 

0.08 

0.14 

0.02 

0.24 

Total I-TEQDF 

Total TEQDF-WHO98 

0.26 

0.27 

0.29 

0.31 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

1.89 

1.92 

5.51 

0.78 

0.64 

7.72 

2.06 

0.56 

0.23 

0.22 

1.89 

1.92 

5.51 

0.78 

0.64 

7.72 

2.06 

0.56 

0.23 

0.24 

Total CDD/CDF 21.53 21.55 a

The same CDD/CDF emission factor was assumed for all three years. 

Sources: U.S. EPA (1996b); Bell (1999). 

United States (U.S. DOC, 1996), and of this amount, 61.3 billion kg were produced by cement 

kilns burning nonhazardous waste (Memorandum dated August 23, 1995, from E. Heath, 

Research Triangle Institute, to J. Wood, U.S. EPA); therefore, approximately 6.3 billion kg were 

produced by cement kilns burning hazardous waste. Based on the fact that 9.3% of the clinker 

5-12
 
Ratio (congener emission factor/2378-CDD/CDF emission factor) 

0.000 0.050 0.100 0.150 0.200 0.250 0.300 

4D 2378 

5D 12378 

6D 123478 

6D 123678 

6D 123789 

7D 1234678 

8D 12346789 

4F 2378 

5F 12378 

5F 23478 

6F 123478 

6F 123678 

6F 123789 

6F 234678 

7F 1234678 

7F 1234789 

8F 12346789 

Ratio (congener group emission factor/total CDD/CDF emission factor) 

0.000 0.050 0.100 0.150 0.200 0.250 0.300 0.350 

4D Total 

5D Total 

6D Total 

7D Total 

8D 12346789 

4F Total 

5F Total 

6F Total 

7F Total 

8F 12346789 

Figure 5-3. Congener profile for air emissions from cement kilns
 

burning nonhazardous waste for reference years 1987, 1995, and 2000.
 

5-13
 
produced in 1995 was from cement kilns burning hazardous waste, it is assumed that 

approximately 4.8 billion kg of the clinker produced in 1987 were from cement kilns burning 

hazardous waste. 

In 2000, cement kilns produced approximately 75.2 billion kg of clinker. This amount 

is based on the assumption that cement kilns operated at 90% of the maximum annual clinker 

capacity of 83.6 billion kg (e-mail dated January 31, 2003, from Garth Hawkins, Portland 

Cement Association, to Karie Riley, Versar, Inc.). Based on the annual clinker capacities of 

individual cement kilns, approximately 11.5 billion kg of clinker (15%) were produced by 

cement kilns burning hazardous waste and approximately 63.7 billion kg of clinker (85%) were 

produced by cement kilns not burning hazardous waste (PCA, 2001). The activity level estimates 

for 1995 and 2000 are given a high confidence rating because they are based on comprehensive 

survey data, but the rating for 1987 is medium because of uncertainty concerning the proportion 

produced by hazardous waste-burning kilns (U.S. EPA, 1996b). 

5.1.6. National CDD/CDF Emission Estimates 

5.1.6.1. Estimates for Reference Years 1987 and 1995 

National estimates of CDD/CDF air emissions (g TEQ/yr) from all Portland cement kilns 

for reference years 1987 and 1995 were made by multiplying the average TEQ emission factors 

by an estimate of the annual activity level (cement clinker produced) for each of the three 

subcategories (hazardous waste-burning kilns with APCD inlet temperatures greater than 232°C 

and less than 232°C and kilns burning nonhazardous waste). Of the 10 hazardous waste-burning 

kilns with APCD temperature data, 8 facilities (80%) had APCD inlet temperatures greater than 

232°C and 2 facilities (20%) had APCD inlet temperatures less than 232°C. The percentages of 

hazardous waste-burning kilns with input temperatures less than and greater than 232°C were 

assumed to represent the actual distribution of activity level in the industry. These percentages, 

coupled with the TEQ emission factors presented in Table 5-1 and Table 5-3 (hazardous waste 

cement kilns and nonhazardous waste cement kilns, respectively) and the activity levels 

established in Section 5.1.5, were used to calculate the annual national TEQ emission estimates 

shown in Table 5-4. 

Overall, 131 g TEQDF-WHO98 (122 g I-TEQDF) were produced by cement kilns in 1987. 

Of this amount, 116.7 g TEQDF-WHO98 (108.6 g I-TEQDF) were produced by hazardous wasteburning cement kilns with inlet APCD temperatures greater than 232°C, 1.1 g TEQDF-WHO98 

(1 g I-TEQDF) were produced by cement kilns burning hazardous waste with inlet APCD 

temperatures less than 232°C, and 12.7 g TEQDF-WHO98 (12.3 g I-TEQDF) were produced by 

cement kilns burning nonhazardous waste. 

5-14
 
Table 5-4. National emission estimates for cement kilns for reference years 

1987 and 1995 

Category 

TEQ emission factor 

(ng/kg clinker) Activity 

level 

(billion kg 

clinker/yr) 

Annual TEQ emission 

(g/yr) 

I-TEQDF 

TEQDF ­

WHO98 I-TEQDF 

TEQDF ­

WHO98 

Reference year 1987 

Hazardous waste >232°C 

Hazardous waste <232°C 

Nonhazardous waste 

28.58 

1.04 

0.26 

30.7 

1.11 

0.27 

3.8 

1 

47.2 

108.6 

1 

12.3 

116.7 

1.1 

12.7 

TOTAL 52 122 131 

Reference year 1995 

Hazardous waste >232°C 

Hazardous waste <232°C 

Nonhazardous waste 

28.58 

1.04 

0.26 

30.7 

1.11 

0.27 

5.04 

1.26 

61.3 

144 

1.3 

15.9 

154.7 

1.4 

16.6 

TOTAL 67.6 161 173 

In 1995, a total of 173 g TEQDF-WHO98 (161 g I-TEQDF) were produced by cement kilns. 

Of this amount, 154.7 g TEQDF-WHO98 (144 g I-TEQDF) were produced by hazardous wasteburning cement kilns with inlet APCD temperatures greater than 232°C, 1.4 g TEQDF-WHO98 

(1.3 g I-TEQDF) were produced by cement kilns burning hazardous waste with inlet APCD 

temperatures less than 232°C, and 16.6 g TEQDF-WHO98 (15.9 g I-TEQDF) were produced by 

cement kilns burning nonhazardous waste. 

The overall rating for these emission estimates is low because the emission factors had a 

low confidence rating. 

5.1.6.2. Estimates for Reference Year 2000 

National estimates of CDD/CDF air emissions (g TEQ/yr) from all Portland cement kilns 

for reference year 2000 were made by multiplying the average TEQ emission factors by an 

estimate of the annual activity level (cement clinker produced) for the nonhazardous waste 

category. The TEQ emission factors presented in Table 5-3 (above) and the activity levels 

established in Section 5.1.5 were used to calculate the annual national TEQ emission estimates 

shown in Table 5-5. For 2000, 17.2 g TEQDF-WHO98 (16.6 g I-TEQDF) were emitted by cement 

kilns not burning hazardous waste. The emission estimates for cement kilns not burning 

hazardous waste in 2000 have an overall low confidence rating because of the small number of 

tested facilities providing the basis for the emission factor. The emission factor may not be 

5-15
 
representative of emissions from cement kilns not burning hazardous waste. However, the 

activity level is given a high confidence rating because it is based on measured data. 

In reference year 2000, 22 cement kilns burning hazardous waste as supplemental fuel 

were operational in the United States, and most had been stack tested and evaluated for dioxin 

emissions. Using the measured data from the stack tests, EPA estimated that 18.8 g TEQDFWHO98 (16.6 g I-TEQDF) were released from all cement kilns burning hazardous waste in 2000. 

The emission estimates for cement kilns burning hazardous waste in 2000 have an overall high 

confidence rating because the emission factors have a high confidence rating due to the large 

number of tested facilities providing the basis for the emission factor. The activity level is given 

a high confidence rating because the amount of clinker produced by cement kilns burning 

hazardous waste is known. 

5.1.7. EPA Regulatory Activities 

In 1999 EPA promulgated national emission standards for new and existing cement kilns 

burning and not burning hazardous waste (Federal Register, 1999a, b). A number of parties, 

representing both industrial and environmental communities, requested judicial review of the rule 

affecting cement kilns and incinerators burning hazardous waste and challenged EPA’s emission 

standards and several implementation provisions. On July 24, 2001, the U.S. Court of Appeals 

for the District of Columbia Circuit vacated the emission standards. In response to this action, 

EPA proposed new standards affecting cement kilns and incinerators burning hazardous waste on 

April 20, 2004 (Federal Register, 2004), which were made final in December 2005 (Federal 

Register, 2005). 

With regard to dioxin emissions, the regulations are specific to the I-TEQ concentration 

in the combustion gases leaving the stack. Existing and new cement kilns either combusting or 

not combusting hazardous waste as supplemental fuel cannot emit more than 0.2 ng I-TEQ/dscm. 

In addition, the temperature of the combustion gases measured at the inlet to the APCD cannot 

exceed 232oC. The rule required owners or operators of facilities to test for CDDs/CDFs every 

21⁄2 years. EPA expects this rule to reduce I-TEQDF emissions from existing and new facilities by 

more than 30% over the next several years (Federal Register, 1999a, b, 2004, 2005). 

5.1.8. Solid Waste from Cement Manufacturing: Cement Kiln Dust 

EPA characterized CKD (the solid residual material generated during the manufacturing 

of cement) in a report to Congress (U.S. EPA, 1993a) that was based in part on a 1991 survey of 

cement manufacturers conducted by the Portland Cement Association (PCA). Survey responses 

were received from 64% of the active cement kilns in the United States. On the basis of the 

5-16
 
Table 5-5. National emission estimates for cement kilns for reference 

year 2000 

Tested facility ID 

Hazardous waste 

burning kilns Estimate in g I-TEQ/yr Estimate in g WHO-TEQ/yr 

200 0.61 0.67 

201 0.01 0.01 

203 3.77 4.24 

204 3.15 3.63 

205 0.14 0.16 

206 1.12 1.39 

207 0.05 0.06 

208 0.88 1.26 

228 0.06 0.07 

300 3.45 3.53 

302 0.23 0.27 

303 0.13 0.16 

318 0.27 0.30 

319 0.50 0.59 

322 0.04 0.04 

323 0.06 0.07 

403 0.32 0.38 

404 0.06 0.07 

473 0.11 0.12 

491 1.59 1.75 

680 0.01 0.01 

681 0.06 0.07 

Total Hazardous 

waste burning kilns 

16.6 18.8 

Total Non-hazardous 

waste burning kilns 

16.6 17.2 

All cement kilns 33.2 36.0 

Source: U.S. EPA (2002a). 

5-17
 
survey responses, EPA estimated that in 1990 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 were land-disposed. The material collected by the APCD system is called gross CKD 

(or as-generated CKD); it is either recycled back into the kiln system or removed from the system 

for disposal (becoming net CKD or as-managed CKD). As discussed below, low levels of dioxin 

have been measured in CKD. This material is disposed of in permitted landfills and therefore is 

not considered to be an environmental release and is not included in the inventory of dioxin 

releases presented in this report. However, for informational purposes only, estimates of the 

amount of dioxin in CKD were developed for the reference years 1987, 1995, and 2000. 

The PCA recently provided current estimates of the amount of CKD removed from the 

manufacturing process for beneficial reuse and long-term management units (i.e., landfill 

disposal) in 1990, 1995, and 2000 (e-mails dated January 31 and March 18, 2003, from Garth 

Hawkins, Portland Cement Association, to Karie Riley, Versor, Inc.). Possible beneficial reuses 

include municipal waste daily cover material, municipal waste landfill final cover material, soil 

stabilization for roadways or other structures, waste neutralization/stabilization/solidification 

(food wastes, hazardous wastes, etc.), and agricultural soil amendment. The PCA estimated that 

the amount of CKD beneficially reused on or off site was 752 million kg in 1990, 652 million kg 

in 1995, and 575 million kg in 2000. The amount of CKD disposed of annually in landfills was 

estimated to be 2.7 billion kg in 1990, 3.1 billion kg in 1995, and 2.2 billion kg in 2000. 

In its report to Congress (U.S. EPA, 1993a), EPA also included the results of sampling 

and analysis of CKD and clinker conducted in 1992 and 1993. The purposes of the sampling and 

analysis efforts were to (a) characterize the CDD/CDF content of clinker and CKD, (b) determine 

the relationship, if any, between the CDD/CDF content of CKD and the use of hazardous waste 

as fuel, and (c) determine the relationship, if any, between the CDD/CDF content of CKD and 

the use of wet-process and dry-process cement kilns. 

Clinker samples were collected from five cement kilns burning nonhazardous waste and 

six kilns burning hazardous waste. CDDs/CDFs were not detected in any of the samples. Tetrathrough octa-chlorinated CDDs/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 ranged from 0.008 to 247 ng I-TEQDF/kg for gross CKD and from 0.045 to 195 ng I

TEQDF/kg for net CKD. Analyses for seven PCB congeners were also conducted, but no 

congeners were detected in any clinker or CKD sample. 

Mean CDD/CDF concentrations in net CKD generated by the kilns burning hazardous 

waste were higher (35 ng I-TEQDF/kg) than in net CKD generated by the facilities burning 

nonhazardous waste (0.003 ng I-TEQDF/kg). These calculations of mean values treated nondetect 

values as zero. If the nondetects had been excluded from the calculation of the means, the mean 

­

5-18
 
for net CKD from kilns burning hazardous waste would increase by a factor of 1.2 and the mean 

for net CKD from kilns burning nonhazardous waste would increase by a factor of 1.7. One 

sampled kiln had a net CKD TEQ concentration of more than two orders of magnitude greater 

than the TEQ levels found in samples from any other kiln. If this kiln were considered atypical 

of the industry (U.S. EPA, 1993a) and were not included in the calculation, then the mean net 

CKD concentration for hazardous waste-burning kilns would decrease to 2.9 ng I-TEQDF/kg. 

CDD/CDF congener data for CKD from Holnam, Inc., Seattle, WA, were presented in a 

report by the Washington State Department of Ecology (1998). The data were compiled and 

evaluated to determine total I-TEQ concentrations and loadings. Nondetect values were included 

as either zero, one-half of the detection limit (DL), or at the DL. The results of three separate 

tests of CKD were as follows, assuming that nondetect values were zero: 

I-TEQ I-TEQ

 Date (ng/kg) (mg/day) 

05/15/96 0.038 0.0038 

10/21/97 0.67 0.0674 

10/21/97 0.95 0.0948 

EPA provided data for ashes from an ESP connected to a cement kiln and an FF 

connected to a lightweight aggregate (LWA) kiln (U.S. EPA, 1999e). The average congener 

concentrations for the ash samples are listed in Table 5-6. The average concentrations for the 

cement kiln were determined from two different waste streams, each with five sample burns. 

The average concentrations for the LWA kiln were determined using one waste stream with three 

sample burns. 

The amount of CDDs/CDFs associated with CKD was calculated for informational 

purposes only. National estimates were divided among cement kilns burning hazardous waste 

and those burning nonhazardous waste for both CKD that was beneficially reused and CKD that 

was sent landfills. The activity levels used in the estimates were those provided by the PCA (email dated March 18, 2003, from Garth Hawkins, Portland Cement Association, to Karie Riley, 

Versar, Inc.). The 1990 activity levels provided by PCA were used for reference year 1987. The 

CDD/CDF concentrations in CKD used in the estimates were 35 ng I-TEQDF/kg for cement kilns 

burning hazardous waste (which includes the high value discussed above) and 0.003 ng I

TEQDF/kg for cement kilns burning nonhazardous waste. 

As shown in Table 5-7, by combining the appropriate activity levels and CDD/CDF 

concentrations, national estimates of CDDs/CDFs in CKD were developed for reference years 

1987, 1995, and 2000. For cement kilns burning hazardous waste, approximately 4.2 g I-TEQDF 

5-19
 

­
Table 5-6. CDD/CDF concentrations (ng/kg) in ash samples from cement 

kiln electrostatic precipitator and lightweight aggregate (LWA) kiln fabric 

filter 

Congener 

Cement 

kiln 

LWA 

kiln Cement kiln LWA kiln 

Avg. 

conc. 

Avg. 

conc. I-TEQ 

WHOTEQ I-TEQ WHO-TEQ 

2,3,7,8-TCDD 

Total TCDD 

1,2,3,7,8-PeCDD 

Total PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

Total HxCDD 

12,2,3,4,6,7,8-HpCDD 

Total HpCDD 

OCDD 

0.429 

36.1 

0.886 

54.9 

1.03 

2.36 

2.47 

173 

17.7 

55.2 

21 

3.97 

333 

17.3 

467 

15.4 

35.6 

56.6 

500 

133 

300 

133 

0.429

 – 

0.443

 – 

0.103 

0.236 

0.247

 – 

0.177

 – 

0.021 

0.429

 – 

0.886

 – 

0.103 

0.236 

0.247

 – 

0.177

 – 

0.0021 

3.97

 – 

8.65

 – 

1.54 

3.56 

5.66

 – 

1.33

 – 

0.133 

3.97

 – 

17.3

 – 

1.54 

3.56 

5.66

 – 

1.33

 – 

0.133 

Total TCDD TEQs 1.66 2.08 24.8 33.5 

2,3,7,8-TCDF 

Total TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

Total PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

Total HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

Total HpCDF 

OCDF 

4.65 

18.1 

1.04 

2.59 

31.8 

2.13 

0.869 

0.523 

2.14 

9.26 

1.84 

0.739 

3.06 

1.43 

833 

4630 

100 

267 

2930 

267 

100 

7.8 

133 

1230 

167 

22.6 

2670 

39.2 

0.465

 – 

0.0518 

1.3

 – 

0.213 

0.0869 

0.0523 

0.214

 – 

0.0184 

0.00739

 – 

0.00143 

0.465

 – 

0.0518 

1.3

 – 

0.213 

0.869 

0.0523 

0.214

 – 

0.0184 

0.00739

 – 

0.000143 

83.3

 – 

5 

133

 – 

26.7 

10 

0.780 

13.3

 – 

1.67 

0.226

 – 

0.0392 

83.3

 – 

5 

133

 – 

26.7 

10 

0.780 

13.3

 – 

1.67 

0.226

 – 

0.00392 

Total TCDF TEQs 2.41 3.19 274 274 

– = No data available 

in 1987, 3.6 g I-TEQDF in 1995, and 3.3 g I-TEQDF in 2000 were produced from CKD that was 

beneficially reused, and approximately 14.9 g I-TEQDF in 1987, 17.7 g I-TEQDF in 1995, and 12.8 

g I-TEQDF in 2000 were produced from CKD that was disposed of in a landfill. For cement kilns 

burning nonhazardous waste, approximately 0.0019 g I-TEQDF in 1987, 0.0016 g I-TEQDF in 

1995, and 0.0014 g I-TEQDF in 2000 were produced from CKD that was beneficially reused, and 

5-20
 
Table 5-7. CDD/CDF estimates in cement kiln dust (CKD) for reference 

years 1987, 1995, and 2000 

Category 

CDD/CDF 

concentration 

(ng I-TEQDF/kg 

of CKD) 

CKD beneficially reused on or 

off site 

CKD sent to a landfill for 

disposal 

Activity level 

(million kg 

tons CKD/yr) 

Annual TEQ 

CDD/CDF 

concentration 

(g/yr) 

Activity 

level 

(million kg 

tons 

CKD/yr) 

Annual TEQ 

CDD/CDF 

concentration 

(g/yr) 

Reference year 1987 

HW kilns 

NHW kilns 

35 

0.003 

120 

632 

4.2 

0.0019

 426 

2,230 

14.9 

0.0067 

Reference year 1995 

HW kilns 

NHW kilns 

35 

0.003 

104 

547 

3.6 

0.0016

 505 

2,642 

17.7 

0.0079 

Reference year 2000 

HW kilns 

NHW kilns 

35 

0.003 

94 

480 

3.3 

0.0014

 365 

1,858 

12.8 

0.0056 

HW = Hazardous waste 

NHW = Nonhazardous waste 

approximately 0.0067 g I-TEQDF in 1987, 0.0079 g I-TEQDF in 1995, and 0.0056 g I-TEQDF in 

2000 were produced from CKD that was disposed of in a landfill. 

EPA is currently developing CKD storage and disposal requirements. In 1999, a 

proposed rule for the standards for the management of CKD was developed by EPA (Federal 

Register, 1999a). Under the rule, CKD would remain a nonhazardous waste, provided that 

proposed management standards are met, which would protect groundwater and control releases 

of fugitive dust. Additionally, the rule proposes concentration limits on various pollutants in 

CKD used for agricultural purposes (Federal Register, 1999c). 

5.2. LIGHTWEIGHT AGGREGATE KILNS 

LWA kilns heat raw materials such as clay, shale, or slate to expand the particles to form 

lightweight materials for use in concrete products. In 1995, only 5 of the more than 36 LWA 

kilns in the United States were burning hazardous waste; in 2000, 9 LWA kilns were burning 

hazardous waste. LWA kilns are estimated to have emitted 3.3 g I-TEQDF to air in 1990 (Federal 

Register, 1998a) and 2.4 g I-TEQDF in 1997 (Federal Register, 1999a); these estimates are used in 

this report for reference years 1987 and 1995, respectively. 

5-21
 
The CDD/CDF emission factors for 2000 are based on the data for five LWA kilns tested 

in 2000 (U.S. EPA, 2002a). They were calculated using the process described in Section 3.2.3. 

The average emission factor for the LWA kilns was 1.986 ng TEQDF-WHO98/kg (2.063 ng 

I-TEQDF/kg) of waste feed, assuming nondetect values of zero. These are assigned a low 

confidence rating because the emission factor may not be representative of emissions from the 

source category. 

The amount of hazardous waste combusted using LWA kilns in 2000 was conservatively 

estimated to be 903,000 metric tons, based on estimated activity levels derived for each halogen 

acid furnace (HAF) in 2000. Data were available for all of the nine facilities operating in 2000. 

A conservative estimate for the average annual quantity burned per HAF (100,280 metric tons/yr) 

was derived by assuming that plants operate continuously throughout the year and are always 

running at 80% of capacity. This quantity, multiplied by the total universe of nine facilities, 

yielded the final estimate. Because the activity level was not derived from a survey but was 

estimated, it is given a low confidence rating. 

Equation 3-5 (Chapter 3, Section 3.2.4), used to calculate annual TEQ emissions for 

dedicated hazardous waste incinerators, was also used to calculate annual TEQ emissions for 

LWA kilns. Multiplying the average TEQ emission factors by the total estimated amount of 

liquid hazardous waste burned in 2000 yields an annual emissions estimate. From this 

procedure, the emissions from all LWA kilns burning hazardous waste as supplemental fuel were 

estimated as 1.86 g TEQDF-WHO98 (1.79 g I-TEQDF) for 2000. Because of the low confidence 

rating for the emission factor, the overall confidence rating for the emission estimates is low. 

5.3. ASPHALT MIXING PLANTS 

Asphalt consists of an aggregate of gravel, sand, and filler mixed with liquid asphalt 

cement or bitumen. Filler typically consists of limestone, mineral stone powder, and sometimes 

ash from power plants and municipal waste combustors. The exact composition of an asphalt 

formulation depends on how it will be used. The aggregate typically constitutes more than 92% 

by weight of the total asphalt mixture. The components of the aggregate are dried, heated to a 

temperature ranging from 135 to 163°C, and then mixed and coated with the bitumen at an 

asphalt mixing installation. “Old” asphalt (i.e., asphalt from dismantled bridges and roads) can 

be heated and disaggregated to its original components and reused in the manufacture of new 

asphalt (U.S. EPA, 1996c). “Hot mix” asphalt paving materials can be manufactured by batch 

mix plants, continuous mix plants, parallel-flow drum mix plants, and counterflow drum mix 

plants (U.S. EPA, 1998c). 

Bremmer et al. (1994) reported the CDD/CDF emissions factor for an asphalt mixing 

plant in the Netherlands as 47 ng I-TEQDF per metric ton of produced asphalt. No congener5-22
 
specific emission factors were reported. The mixing plant heated old asphalt to about 150°C in 

an individual recycling drum with kiln exhaust gases that were mixed with ambient air and 

heated to a temperature of 300 to 400°C. Parallel to this recycling drum was the main drum, 

which dried and heated the aggregate (sand and gravel/granite chippings) to a temperature of 

about 220°C. The kiln exhaust gases leaving the recycling drum were led along the main burner 

of the main drum for incineration. The old asphalt, the minerals from the main drum, and new 

bitumen from a hot storage tank (about 180°C) were mixed in a mixer to form new asphalt. 

Natural gas fueled the plant during the sample collection period, and 46% of the feed was old 

asphalt. The plant’s APCD system consisted of cyclones and an FF. 

Umweltbundesamt (1996) reported lower emission factors for three tested facilities in 

Germany that were also equipped with FFs. These three facilities were fueled by oil or butane 

gas and used old asphalt at rates ranging from 30 to 60% of the feed. The emission factors 

calculated from the stack gas concentrations, gas flow rates, and hourly throughputs for these 

three facilities were 0.2, 3.5, and 3.8 ng I-TEQDF/metric ton of asphalt produced, respectively. 

EPA conducted stack emissions testing at two hot-mix asphalt production plants in the 

United States (U.S. EPA, 2000g). The Midwest Research Institute performed emission tests on 

outlet of FFs that controlled emissions from the counter-flow rotary dryer process used at the 

asphalt plant in Clayton, NC, and from the parallel-flow rotary dryer process used at the asphalt 

plant in Cary, NC. In both processes, virgin aggregate of various sizes was fed to the drum by 

cold-feed controls in proportions dictated by the final mix specifications. Aggregate was 

delivered at the opposite end of the burner in the counter-flow continuous drum mix process and 

at the same end as the burner in the parallel-flow continuous drum mix process. EPA developed 

average TEQ emission factors from the tested facilities to represent hot-mix asphalt plants in 

general (U.S. EPA, 2000g). The average TEQ emission factor for hot-mix asphalt plants was 

1.40e–03 ng I-TEQ (1.46e–3 ng TEQDF-WHO98/kg) asphalt produced. Table 5-8 summarizes the 

TEQ emission factors for hot-mix asphalt plants. 

Because only two U.S. facilities have been tested, these emission factors are given a low 

confidence rating. It is likely that these emission factors are not representative of CDD/CDF 

emissions from all types of asphalt production facilities in the United States; nevertheless, they 

are sufficient to derive a preliminary estimate of annual emissions from asphalt production 

facilities in the United States for 2000. 

Approximately 500 million tons of hot-mix asphalt paving materials were produced at 

approximately 3,600 active asphalt plants in the United States in 1996 (U.S. EPA, 1998c). This 

activity level was used to represent reference year 2000. The activity level is given a high 

confidence level because it was based on a comprehensive survey. A preliminary estimate of 

annual TEQ air releases for 2000 was calculated by multiplying the activity level (in kilograms) 

5-23
 
Table 5-8. Congener-specific emission factors (ng/kg asphalt produced) 

derived from the testing of air emissions at two hot-mix asphalt plants in the 

United States 

Congener 

Mean emission factor 

I-TEQDF TEQDF-WHO98 

2,3,7,8-TCDD 

1,2,3,7,8-PeCD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

9.53e–05 

7.03e–05 

1.91e–05 

5.90e–05 

4.45e–05 

2.18e–05 

1.13e–05 

9.53e–05 

1.41e–04 

1.91e–05 

5.90e–05 

4.45e–05 

2.18e–05 

1.13e–06 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

4.40e–05 

9.75e–05 

1.91e–04 

1.81e–04 

5.44e–05 

8.62e–05 

3.81e–04 

2.95e–05 

1.22e–05 

2.18e–06 

4.40e–05 

9.75e–05 

1.91e–04 

1.81e–04 

5.44e–05 

8.62e–05 

3.81e–04 

2.95e–05 

1.22e–05 

2.18e–07 

Total TEQ 1.40e–03 1.50e–03 

Source: U.S. EPA (2000g). 

by the average emission factor of 1.40e–03 ng I-TEQ (1.46e–3 ng TEQDF-WHO98/kg) asphalt 

produced. Approximately 0.70 g I-TEQ (0.73 g TEQDF-WHO98) were emitted into the air from 

asphalt plants in the United States in 2000. This estimate is assigned a low confidence level 

because the emission factors are given a low confidence level rating. 

5.4. PETROLEUM REFINING CATALYST REGENERATION 

Regeneration of spent catalyst from the reforming process at petroleum refineries is a 

potential source of CDDs/CDFs, according to limited testing conducted in the United States 

(Amendola and Barna, 1989; Kirby, 1994), Canada (Maniff and Lewis, 1988; Thompson et al., 

1990), and the Netherlands (Bremmer et al., 1994). This section summarizes the catalyst 

regeneration process, relevant studies performed to date, and the status of EPA regulatory 

investigations of this source. 

5-24
 
Catalytic reforming is the process used to produce high-octane reformates from loweroctane reformates for blending of high-octane gasolines and aviation fuels. The reforming 

process occurs at high temperature and pressure and requires the use of a platinum or 

platinum/rhenium catalyst. During the reforming 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 at temperatures of 

399 to 454°C and then reactivating the catalyst at elevated temperatures (454 to 538°C) using 

chlorine or chlorinated compounds (e.g., methylene chloride, 1,1,1-trichloroethane, and ethylene 

dichloride; most refineries use chlorine of perchloroethylene). Burning of the coke produces kiln 

exhaust gases that can contain CDDs and CDFs along with other combustion products. Because 

kiln exhaust gases, if not vented directly to the atmosphere, may be scrubbed with caustic or 

water, internal effluents may become contaminated with CDDs/CDFs (Kirby, 1994; SAIC, 

1994). 

Three basic catalyst regeneration processes are used: semi-regenerative, cyclic, and 

continuous. During the semi-regenerative process, the entire catalytic reformer is taken off line. 

In the cyclic process, one of two (or more) reforming reactors is taken off line for catalyst 

regeneration; the remaining reactor(s) remains on line so that reforming operations continue. In 

the continuous process, aged catalyst is continuously removed from one or more on-line stacked 

or side-by-side reactors, regenerated in an external regenerator, and then returned to the system; 

the reforming system, consequently, never shuts down (SAIC, 1994). 

In 1988, a study by the Canadian Ministry of the Environment (Maniff and Lewis, 1988) 

detected concentrations of CDDs ranging from 1.8 to 22.2 μg/L and CDFs ranging from 4.4 to 

27.6 μg/L in an internal waste stream of spent caustic at a petroleum refinery. The highest 

concentration of 2,3,7,8-TCDD was 0.0054 μg/L. CDDs were also observed in the refinery’s 

biological sludge at a maximum concentration of 74.5 μg/kg, and CDFs were observed at a 

maximum concentration of 125 μg/kg. The concentration of CDDs/CDFs in the final combined 

refinery plant effluent was below the DLs. 

Amendola and Barna (1989) reported detecting trace levels of hexa- to octa-CDDs and 

CDFs in untreated wastewaters (up to 2.9 pg I-TEQDF/L) and wastewater sludges (0.26 to 2.4 ng 

I-TEQDF/kg) at a refinery in Ohio. The levels of detected total CDDs/CDFs in the wastewater 

and sludge were much lower (<3 ng/L and <1 μg/kg, respectively) than the levels reported by 

Maniff and Lewis (1988). No CDDs/CDFs were detected in the final treated effluent (less than 

0.2 ng I-TEQDF/L). The data collected in the study were acknowledged to be too limited to 

enable identifying the source(s) of the CDDs/CDFs within the refinery. The study authors also 

presented in an appendix to their report the results of analyses of wastewater from the catalyst 

5-25
 
regeneration processes at two other U.S. refineries. In both cases, untreated wastewaters 

contained CDDs/CDFs at levels ranging from high picograms to low nanograms per liter (results 

were reported for congener group totals, not specific congeners). However, CDDs/CDFs were 

not detected in the only treated effluent sample collected at one refinery. 

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 a Canadian petroleum refinery’s reforming 

operation. They also observed CDDs/CDFs in the picogram- to nanogram-per-liter range in the 

internal washwater from a scrubber of a periodic/cyclic regenerator. 

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 kiln exhaust gas from burning coked catalysts revealed the presence of 

unchlorinated dibenzofuran in quantities up to 220 μg/kg 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 seemed to be the most likely chlorinating agent. 

In May 1994, EPA’s Office of Water conducted a sampling and analytical study of 

catalyst regeneration wastewater for CDDs/CDFs at three petroleum refining plants (Kirby, 

1994). The study objectives were to determine the analytical method best suited for determining 

CDDs/CDFs in refinery wastewater and to screen and characterize wastewater discharges from 

several types of reforming operations for CDDs/CDFs. The report for this study (Kirby, 1994) 

also presented results submitted voluntarily to EPA by two other facilities. The sampled internal 

untreated wastewaters and spent caustics were found to contain a wide range of CDD/CDF 

concentrations, 0.1 pg I-TEQDF/L to 57.2 ng I-TEQDF/L. The study results also showed that 90% 

of the TEQ was contained in the wastewater treatment sludges generated during the treatment of 

wastewater and caustic from the regeneration process. 

In 1995, EPA issued a notice of its proposed intent to not designate spent reformer 

catalysts as a listed hazardous waste under RCRA (Federal Register, 1995b). The final rule was 

issued in August 1998 (Federal Register, 1998b). The Agency’s assessment of current 

management practices associated with recycling of reforming catalyst found no significant risks 

to human health or the environment. The Agency estimated that 94% of the approximately 3,600 

metric tons of spent reformer catalyst sent off site by refineries were being recycled for their 

precious metal content. EPA made no determination of the “listability” of spent caustic residuals 

formed during regeneration of spent reforming catalyst, but it did identify the potential air 

releases from the combustion of the reforming catalyst prior to reclamation as possibly being of 

concern. The Agency requested comments on (a) opportunities for removing dioxin prior to 

5-26
 
discharge of scrubber water into the wastewater treatment system, (b) opportunities to segregate 

this wastestream, and (c) potential health risks associated with insertion of dioxin-contaminated 

media back into the refinery process (such as the coker). In this proposed rulemaking, EPA also 

noted the possibility of dioxin releases to air during regeneration operations. 

As part of its regulatory investigation under RCRA, EPA’s Office of Solid Waste 

commissioned a study to analyze and discuss existing data and information concerning 

CDD/CDF formation in the treatment of catalytic reformer wastes. This report (SAIC, 1994) 

also identified potential process modifications that may prevent the formation of CDDs/CDFs. 

The report’s authors concluded that, although the available data indicate that CDDs/CDFs can be 

generated during the catalyst regeneration process, the available data indicate that CDD/CDF 

concentrations in treated wastewater and in solid waste are minimal. Releases to air could result 

from vented kiln exhaust gases at some facilities. In addition, the CDDs/CDFs formed could 

possibly be reintroduced into other refining operations (e.g., the coker) and resulting products. 

In 1998, emissions from the caustic scrubber used to treat gases from the external 

regeneration unit of a refinery in California were tested (CARB, 1999). This facility uses a 

continuous regeneration process. The reactor is not taken off line during regeneration; rather, 

small amounts of catalyst are continuously withdrawn from the reactor and are regenerated. The 

emissions from the regeneration unit are neutralized by a caustic scrubber before being vented to 

the atmosphere. The catalyst recirculation rate during the three tests ranged from 733 to 1,000 

lb/hr. 

All 2,3,7,8-substituted CDDs/CDFs were detected in each of the three samples collected. 

The average emission factors in units of nanograms per barrel of reformer feed are presented in 

Table 5-9. The congener profile is presented in Figure 5-4. The samples showed a wide range in 

concentrations of the CDD/CDF congeners (up to a fivefold difference); however, the congener 

profile was consistent in all samples. The concentrations of the individual furan congener groups 

were always higher than the concentrations of the corresponding dioxin congener group. The 

average TEQDF-WHO98 emission factor for these three tests was 3.18 ng TEQ/barrel and the 

average I-TEQDF was 3.04 ng TEQ/barrel. 

In 1991, stack testing was performed on the exhaust from one of the three semiregenerative catalytic reforming units of a refinery in California (Radian Corporation, 1991). At 

these units, a caustic solution is introduced to the exhaust to neutralize HCl emissions from the 

catalyst beds prior to release to the atmosphere. The tested unit was considered to be 

representative of the other units. Each unit is periodically taken off line (approximately once a 

year) so the catalyst beds can be regenerated. The tested unit has a feed capacity of 7,000 

barrels/day. Approximately 59,500 pounds of catalyst were regenerated during the tested 

regeneration cycle, which lasted for 62 hr. 

5-27
 
The average emission factors for this facility (in units of nanogams per barrel of reformer 

feed) are presented in Table 5-9 and the congener profile is presented in Figure 5-4. The 

majority of the 2,3,7,8-substituted CDD congeners were not detected during testing. In contrast, 

the majority of the 2,3,7,8-substituted CDF congeners were detected. The average TEQDFWHO98 emission factor (assuming nondetect values were zero) was 1.04e–03 ng TEQ/barrel and 

the average I-TEQDF emission factor was 1.01e–03 ng TEQ/barrel. These values are three orders 

of magnitude less than the emission factor reported in CARB (1999). The calculation of these 

emission factors involved several assumptions: the unit is regenerated once per year, the unit 

operates at capacity (7,000 barrels/day), and the facility operates 362 days/yr. 

The average of the two facility emission factors, 1.59 ng TEQDF-WHO98/barrel (1.52 ng I

TEQDF/barrel) of reformer feed, is assumed to apply to all reference years (1987, 1995, and 2000) 

and is assigned a low confidence rating. Only one continuous and one semiregenerative unit in 

the United States have been tested. Combined, these two facilities represent less than 1% of the 

catalytic reforming capacity in U.S. petroleum refineries in 1987 (3.805 million barrels/day), 

1995 (3.867 million barrels/day), and 2000 (3.770 million barrels/day) (EIA, 2002a). The 

average emission factor developed above assumes that emissions are proportional to reforming 

capacity; however, they may be more related to the amount of coke burned, the APCD equipment 

present, or other process parameters. 

The national daily average catalytic reforming capacities in the United States were 3.805, 

3.867, and 3.770 million barrels per day for 1987, 1995, and 2000, respectively (EIA, 2002). 

These were assigned a high confidence rating because they are based on comprehensive surveys 

of industry. If it is conservatively assumed that all units operated at full capacity in all three 

years, then applying the average emission factors of TEQ/barrel yields annual emissions of 2.21 g 

TEQDF-WHO98 (2.11 g I-TEQDF) in 1987, 2.24 g TEQDF-WHO98 (2.14 g I-TEQDF) in 1995, and 

2.19 g TEQDF-WHO98 (2.09 g I-TEQDF) in 2000. These emissions have a low confidence rating 

because they are based on an emission factor with a low confidence rating. 

5.5. CIGARETTE SMOKING 

Bumb et al. (1980) were the first to report that cigarette smoking is a source of CDD 

emissions. Subsequent studies by Muto and Takizawa (1989), Ball et al. (1990), and Löfroth and 

Zebühr (1992) also reported the presence of CDDs as well as CDFs in cigarette smoke. A study 

by Matsueda et al. (1994) reported on the CDD/CDF content of tobacco from 20 brands of 

cigarettes from seven countries. Although a wide range in the concentrations of total 

CDDs/CDFs and total TEQs were reported in these studies, similar congener profiles and 

patterns were reported. The findings of each of these studies are described in this section. 

­

5-28
 
Table 5-9. CDD/CDF emission factors (ng/barrel)a

 for petroleum catalytic 

reforming units 

Congener/congener 

group 

Semiregenerative unit Continuous regeneration unit 

Nondetect set 

to zero 

Nondetect set to 

1⁄2 detection limit 

Nondetect set 

to zero 

Nondetect set to 

1⁄2 detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND 

5.69e!05 

4.22e!05 

ND 

ND 

7.02e!04 

2.55e!03 

2.35e!05 

9.58e!05 

8.09e!05 

5.52e!05 

5.10e!05 

7.02e!04 

2.55e!03 

1.61e!02 

2.87e!01 

3.47e!01 

8.45e!01 

5.56e!01 

3.02e+00 

1.71e+00 

1.61e!02 

2.87e!01 

3.47e!01 

8.45e!01 

5.56e!01 

3.02e+00 

1.71e+00 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

2.32e!04 

4.68e!04 

1.09e!03 

1.06e!03 

1.07e!03 

ND 

1.24e!03 

2.94e!03 

8.32e!04 

1.01e!03 

2.32e!04 

4.68e!04 

1.09e!03 

1.06e!03 

1.07e!03 

6.82e!05 

1.24e!03 

2.94e!03 

8.32e!04 

1.01e!03 

6.10e!01 

1.72e+00 

2.33e+00 

4.70e+00 

3.58e+00 

4.34e!01 

3.10e+00 

1.59e+01 

1.45e+00 

3.75e+00 

6.10e!01 

1.72e+00 

2.33e+00 

4.70e+00 

3.58e+00 

4.34e!01 

3.10e+00 

1.59e+01 

1.45e+00 

3.75e+00 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

3.35e!03 

9.94e!03 

1.01e!03 

1.03e!03 

3.56e!03 

1.00e!02 

1.07e!03 

1.12e!03 

6.78e+00 

3.76e+01 

3.04e+00 

3.18e+00 

6.78e+00 

2.18e+01 

3.04e+00 

3.18e+00 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

ND 

3.56e!04 

1.28e!03 

1.39e!03 

2.55e!03 

2.70e!03 

5.12e!03 

7.85e!03 

4.88e!03 

1.01e!03 

2.35e!05 

3.56e!04 

1.28e!03 

1.39e!03 

2.55e!03 

2.70e!03 

5.12e!03 

7.85e!03 

4.88e!03 

1.01e!03 

6.84e+00 

5.61e+00 

8.18e+00 

6.58e+00 

1.71e+00 

4.68e+01 

3.30e+01 

2.96e+01 

2.11e+01 

3.75e+00 

6.84e+00 

5.61e+00 

8.18e+00 

6.58e+00 

1.71e+00 

4.68e+01 

3.30e+01 

2.96e+01 

2.11e+01 

3.75e+00 

Total CDD/CDF 2.71e!02 2.72e!02 1.63e+02 1.63e+02 

a

One barrel assumed to be equivalent to 139 kg.

 ND = Not detected

 Sources: Radian Corporation (1991); CARB (1999). 

5-29 
Ratio (mean congener emission factor/total CDD/CDF emission factor) 

0a 0.02 0.04 0.06 0.08 0.1 0.12

 Ratio (mean congener emission factor/total CDD/CDF emission factor) 

0a 0.05 0.1 0.15 0.2 0.25 0.3 

Figure 5-4. Congener and congener group profiles for air emissions from 

petroleum catalytic reforming units (nondetect set equal to zero). 

Sources: CARB (1999); Radian Corporation (1991). 

5-30
 

0.35 
No studies published to date have demonstrated a mass balance, and it is not known 

whether the CDDs/CDFs measured in cigarette smoke are the result of formation during tobacco 

combustion, volatilization of CDDs/CDFs present in the unburned tobacco, or a combination of 

these two source mechanisms. The combustion processes operating during cigarette smoking are 

complex and could be used to justify both source mechanisms. As reported by Guerin et al. 

(1992), during a puff on a cigarette, gas-phase temperatures reach 850°C at the core of the 

firecone, and solid-phase temperatures reach 800°C at the core and 900°C or greater at the char 

line. Thus, temperatures are sufficient to cause at least some destruction of CDDs/CDFs initially 

present in the tobacco. Both solid- and gas-phase temperatures rapidly decline to 200 to 400°C 

within 2 mm of the char line. 

Formation of CDDs/CDFs has been reported in combustion studies with other media in 

this temperature range of 200 to 900°C. However, it is known that a process likened by Guerin 

et al. (1992) to steam distillation takes place in the region behind the char line because of high, 

localized concentrations of water and temperatures of 200 to 400°C. At least 1,200 tobacco 

constituents (e.g., nicotine, n-paraffin, some terpenes) are transferred intact from the tobacco into 

the smoke stream by distillation in this region, and it is plausible that CDDs/CDFs present in the 

unburned tobacco would be subject to similar distillation. 

Bumb et al. (1980), using low-resolution mass spectrometry, analyzed the CDD content 

of mainstream smoke from the burning of a U.S. brand of unfiltered cigarette. A package of 20 

cigarettes was combusted in each of two experiments. Approximately 20 to 30 puffs of 2 to 3 sec 

duration were collected from each cigarette on a silica column. Hexa-, hepta-, and octa-CDDs 

were detected at levels of 0.004 to 0.008, 0.009, and 0.02 to 0.05 ng/g, respectively. 

Muto and Takizawa (1989) employed a continuous smoking apparatus to measure CDD 

congener concentrations in the mainstream smoke generated from the combustion of one kind of 

filtered cigarette (brand not reported). The apparatus pulled air at a constant continuous rate 

(rather than a pulsed rate) through a burning cigarette and collected the smoke on a series of traps 

(glass fiber filter, polyurethane foam, and XAD-II resin). The CDD content of the smoke as well 

as the CDD content of the unburned cigarette and the ash from the burned cigarettes were also 

analyzed using low-resolution mass spectrometry. The results are presented in Table 5-10, and 

the congener group profiles are presented in Figure 5-5. Table 5-11 and Figure 5-6 present the 

mainstream smoke results on a mass-per-cigarette basis to enable comparison with the results of 

other studies. 

The major CDD congener group found was HpCDD, which accounted for 84% of total 

CDDs found in the cigarette, 94% of total CDDs found in smoke, and 99% of total CDDs found 

in the ash. The 2,3,7,8-HpCDDs also accounted for the majority of the measured TEQ in the 

5-31
 
Table 5-10. CDD concentrations in Japanese cigarettes, smoke, and ash 

Congener/congener group 

Cigarette 

(pg/g) 

Concentrations 

Mainstream smoke 

(ng/m3) 

Ash 

(pg/g) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND (0.5) 

ND (0.5) 

2.01a 

a 

a 

1343 

257 

ND (0.22) 

0.43 

2.15a 

a 

a 

783 

240 

ND (0.5) 

ND (0.5) 

0.56a 

a 

a 

ND (0.5) 

ND (0.5) 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

1602.01 

NR 

13.9 

13.7 

1025.58 

NR 

8.5 

8.3 

0.56 

NR 

0.06 

0.06 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

44.9 

ND (0.5) 

13.41 

1629 

257 

NR 

NR 

NR 

NR 

NR 

68 

1.51 

7.51 

4939 

240 

NR 

NR 

NR 

NR 

NR 

4.63 

ND (0.5) 

5.01 

3211 

ND (0.5) 

NR 

NR 

NR 

NR 

NR 

Total CDD/CDF 1944 5256 3221 a

Value reported only for total 2,3,7,8-substituted HxCDDs.

 ND = Not detected (value in parenthesis is the detection limit)

 NR = Not reported

 Source: Muto and Takizawa (1989). 

5-32
 
Figure 5-5. CDD profiles for Japanese cigarettes, smoke, and ash. 

Source: Matsueda et al. (1994). 

cigarettes and smoke; however, none were measured in the ash. Although no PeCDDs were 

detected in the cigarette, they were detected at low levels in the smoke, indicating probable 

formation during combustion. On the basis of the similarities in the congener group profiles for 

the three media, the study authors concluded that most of the CDDs found in the cigarette smoke 

result from volatilization of CDDs/CDFs present in the unburned cigarette rather than being 

formed during combustion. 

Ball et al. (1990) measured the CDD/CDF content of mainstream smoke for the 10 bestselling German cigarette brands. The international test approach (1 puff/min; puff flow rate of 

35 mL/2 sec) was employed with an apparatus that smoked 20 cigarettes at a time in three 

successive batches and had a large collection device. The average TEQ content (on both an I

TEQDF and a TEQDF-WHO98 basis) in mainstream smoke for the 10 brands tested, normalized to 

a mass-per-cigarette basis, was 0.09 pg/cigarette (i.e., 16.5 times less than the value reported by 

Muto and Takizawa, 1989, for a Japanese cigarette brand). However, the congener group 

5-33
 

­
Table 5-11. CDD/CDF concentrations (pg/cig) in cigarette smoke, 

normalized to a per-cigarette basis

a 

Congener/congener 

group 

Muto and 

Takizawa (1989) 

(1 Japanese brand) 

(mainstream 

smoke) 

Ball et al. (1990) 

(avg. of 10 German 

brands) 

(mainstream 

smoke) 

Löfroth and 

Zebühr (1992) 

(1 Swedish brand) 

(mainstream 

smoke) 

Löfroth and 

Zebühr (1992) 

(1 Swedish brand) 

(sidestream smoke) 

2,3,7,8-TCDD ND (0.04) ND (0.03) 0.028 0.07 

1,2,3,7,8-PeCDD 0.075 ND (0.03) 0.15 0.32 

1,2,3,4,7,8-HxCDD 0.376 0.06 0.1 0.19 

1,2,3,6,7,8-HxCDD b 0.05 0.34 0.6 

1,2,3,7,8,9-HxCDD b 0.04 0.25 0.55 

1,2,3,4,6,7,8-HpCDD 137 1.3 6.05 12.2 

OCDD 42 3.4 22.1 38.8 

2,3,7,8-TCDF NR 0.19 1.2c 2.1c 

1,2,3,7,8-PeCDF NR 0.13 0.34c 0.8c 

2,3,4,7,8-PeCDF NR 0.04 0.34 0.6 

1,2,3,4,7,8-HxCDF NR ND (0.03) 1.3c 3.8c 

1,2,3,6,7,8-HxCDF NR 0.03 0.48 1.2 

1,2,3,7,8,9-HxCDF NR 0.03 0.14 0.39 

2,3,4,6,7,8-HxCDF NR 0.05 0.21 0.5 

1,2,3,4,6,7,8-HpCDF NR 0.16 10 23.5 

1,2,3,4,7,8,9-HpCDF NR 0.03 2.6 5 

OCDF NR 0.11 3.2 10.7 

Total 2,3,7,8-CDD 179.45 4.85 29.02 52.7 

Total 2,3,7,8-CDF NR 0.77 19.81 48.6 

Total I-TEQDF 1.49 0.09 0.9 1.96 

Total TEQDF-WHO98 1.49 0.09 0.96 2.08 

Total TCDD 11.9 0.51 0.61 0.67 

Total PeCDD 0.264 0.14 1.07 2.14 

Total HxCDD 1.31 0.53 2.52 5.2 

Total HpCDD 864 2.9 12.3 21.3 

Total OCDD 42 3.4 22.1 38.8 

Total TCDF NR 1.41 4.5 5.75 

Total PeCDF NR 0.83 3.23 6.35 

Total HxCDF NR 0.35 5.3 12.9 

Total HpCDF NR 0.27 19.8 47.8 

Total OCDF NR 0.11 3.2 10.7 

Total CDD/CDF 919.47 10.45 74.63 151.6 

a

Emissions calculated assuming 0.0035 m3 of smoke are inhaled per 20 cigarettes smoked (Muto and Takizawa,

 1992). 

b

Muto and Takizawa (1989) reported a value only for total 2,3,7,8-HxCDDs (0.38 pg/cig). c

Concentrations listed include the contribution of a coeluting non-2,3,7,8-substituted congener.

 ND = Not detected (value in parenthesis is the detection limit)

 NR = Not reported 

5-34
 
1 

0.9 

0.8 

0.7 

0.6 

0.5 

0.4 

0.3 

0.2 

0.1
 

0
 TCDD Pe CDD HxCDD Hp CDD OCDD TCDF Pe CDF HxCDF Hp CDF OCDF 

Jap ane se Mainstre am (row 1) Ge rman Mainstre am (row 2) 

Swe d ish Mainstre am (row 3) Swe d ish Sid e stre am (row 4) 

Figure 5-6. Congener group profiles for mainstream and sidestream 

cigarette smoke. 

Source: Matsueda et al. (1994). 

profiles were similar to those reported by Muto and Takizawa, with HpCDD and OCDD the 

dominant congener groups found. 

Löfroth and Zebühr (1992) measured the CDD/CDF content of mainstream and 

sidestream smoke from one common Swedish cigarette brand. The cigarette brand was labeled 

as giving 17 mg carbon monoxide, 21 mg tar, and 1.6 mg nicotine. The international test 

approach was used, and the smoke was collected on glass fiber filters followed by two 

polyurethane plugs. The analytical results for mainstream and sidestream smoke are presented in 

Table 5-11. The TEQ content in mainstream smoke, normalized to a mass-per-cigarette basis, 

was 0.96 pg TEQDF-WHO98/cigarette (0.9 pg I-TEQDF/cigarette) (i.e., about two times less than 

the value reported by Muto and Takizawa, 1989, and 10 times greater than the average value 

reported by Ball et al., 1990). As in the Muto and Takizawa and Ball et al. studies, the dominant 

congener groups were HpCDDs and OCDD; however, HpCDFs were also relatively high in 

5-35
 
comparison with the other congener group totals. The sidestream smoke contained 2.08 pg 

TEQDF-WHO98/cigarette (1.96 pg I-TEQDF/cigarette), or twice that of mainstream smoke. 

Using high-resolution mass spectrometry, Matsueda et al. (1994) analyzed the CDD/CDF 

content of tobacco from 20 brands of commercially available cigarettes collected in 1992 from 

Japan, the United States, Taiwan, China, the United Kingdom, Germany, and Denmark. Table 5­

12 presents the study results. The total CDD/CDF content ranged from 109 to 1,136 pg/pack, 

and total TEQ ­ DF-WHO98 content ranged from 1.9 to 14 pg/pack (1.4 to 12.6 pg/pack on an I

TEQDF basis). The Chinese cigarette brand contained significantly lower CDDs/CDFs and TEQs 

than did any other brand of cigarette. Figure 5-7 depicts the congener group profiles for the 

average results for each country. A high degree of similarity is seen among the CDF congener 

group profiles of the tested cigarette brands. The Japanese and Taiwanese cigarettes show CDD 

congener group profiles different from those of the other countries’ cigarettes. 

Brown (2002) estimated that 440 billion cigarettes were consumed in the United States in 

2000. In 1995, approximately 487 billion cigarettes were consumed in the United States and by 

U.S. overseas armed forces personnel. In 1987, approximately 575 billion cigarettes were 

consumed. According to The Tobacco Institute (1995), per capita U.S. cigarette consumption, 

based on the total U.S. population aged 16 and over, was a record high of 4,345 in 1963, 

declining to 2,415 in 1995 and 1,563 in 2000 (USDA, 1997; U.S. Census Bureau, 2000). The 

activity level estimates by Brown (2002) were adopted, and a high confidence rating is assigned 

because they are based on known consumption rates. 

The available emission factor data presented above provide the basis for two methods of 

estimating the amount of TEQs that may have been released to the air in the United States in 

2000, 1995, and 1987 from the combustion of cigarettes. The confidence rating assigned to the 

emission factor is low because of the very limited amount of testing performed to date. First, an 

annual emission estimate for 2000 of 0.19 g TEQ (on a TEQDF-WHO98 or I-TEQDF basis) is 

obtained if it is assumed that (a) the average TEQ content of seven brands of U.S. cigarettes 

reported by Matsueda et al. (1994)—6.3 pg TEQDF-WHO98/pack (5.5 pg I-TEQDF/pack)—is 

representative of cigarettes smoked in the United States, (b) CDDs/CDFs are not formed and the 

congener profile reported by Matsueda et al. (1994) is not altered during combustion of 

cigarettes, and (c) all CDDs/CDFs contributing to the TEQ are released from the tobacco during 

smoking. 

The available emission factor data presented above provide the basis for two methods of 

estimating the amount of TEQs that may have been released to the air in the United States in 

2000, 1995, and 1987 from the combustion of cigarettes. The confidence rating assigned to the 

emission factor is low because of the very limited amount of testing performed to date. First, an 

5-36
 
Table 5-12. CDD/CDF concentrations (pg/pack) in tobacco cigarette brands 

from various countries 

U.S. 

(avg. of 7 

brands) 

Japan 

(avg. of 6 

brands) 

United 

Kingdom 

(avg. of 3 

brands) 

Taiwan 

(1 brand) 

China 

(1 brand) 

Denmark 

(1 brand) 

Germany 

(1 brand) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCD 

D 

OCDD 

1.2 

1.6 

6.9 

a 

a 

52.7 

589.3 

0.5 

1.4 

4.8 

a 

a 

17.8 

244 

1.7 

3.1 

6.1 

a 

a 

23.9 

189.5 

1 

3.3 

12.2 

a 

a 

26.4 

272.7 

ND 

1.1 

1.1 

a 

a 

2.2 

28.2 

0.5 

0.8 

6.2 

a 

a 

53.3 

354.3 

1.1 

3.3 

5.7 

a 

a 

32.7 

288.6 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

18.2 

8.7 

b 

8.1 

c 

c 

c 

17.6 

d 

24.6 

4.8 

5.3 

b 

8.1 

c 

c 

c

 11.1 d 

10.5 

15.6 

21.2 

b 

17 

c 

c 

c 

13.6 

d 

8.3 

11 

16 

b 

12.9 

c 

c 

c 

13.2 

d 

13.9 

1.2 

1.5 

b 

2.2 

c 

c 

c 

1.5 

d 

0.5 

2.2 

4.3 

b 

4.3 

c 

c 

c 

7 

d 

10.5 

7.9 

14.4 

b 

13.2 

c 

c 

c 

12.9 

d 

13.9 

Total 2,3,7,8-CDD 651.7 268.5 224.3 315.6 32.6 415.1 331.4 

Total 2,3,7,8-CDF 77.2 39.8 75.7 67 6.9 28.3 62.3 

Total I-TEQDF 7.1 3.8 8.8 7.7 1.1 3.4 6.9 

Total TEQDF-WHO98 7.3 4.3 10.1 9.1 1.7 3.4 8.3 

Total TCDD 47.1 296.3 85.1 329 9.7 17 49.5 

Total PeCDD 27.6 33.6 62.9 150.5 5.2 9.8 40.8 

Total HxCDD 40.6 29.2 49.2 99.4 5.4 26.7 40.6 

Total HpCDD 108.7 40 47.7 62 3.8 93.1 60.2 

Total OCDD 589.3 244 189.5 272.7 28.2 354.3 288.6 

Total TCDF 183.8 102.1 348.9 372.1 35.4 97.8 233.4 

Total PeCDF 57.7 45.9 134.5 149.1 11.2 35.5 97.5 

Total HxCDF 29.1 26.4 51.3 45.8 7.8 18.1 40.8 

Total HpCDF 27.3 16.6 19 18.5 1.7 11.1 21.2 

Total OCDF 24.6 10.5 8.3 13.9 0.5 10.5 13.9 

Total CDD/CDF 1,135.8 844.6 996.4 1513 108.9 673.9 886.5 

a

Value reported only for total 2,3,7,8-substituted HxCDDs. b

Value reported only for total 2,3,7,8-substituted PeCDFs. c

Value reported only for total 2,3,7,8-substituted HxCDFs. d

Value reported only for total 2,3,7,8-substituted HpCDFs.

 Source: Matsueda et al. (1994). 

annual emission estimate for 2000 of 0.19 g TEQ (on a TEQDF-WHO98 or I-TEQDF basis) is 

obtained if it is assumed that (a) the average TEQ content of seven brands of U.S. cigarettes 

5-37
 
0.6 

0.5 

0.4 

0.3 

0.2 

0.1 

0 TCDD PeCDD HxCDD HpCDD OCDD TCDF PeCDF HxCDF HpCDF OCDF 

Ch Japan (column 3) ina (column 1) Denmark (column 2) 

Germany (column 4) United Kingdom (column 5) United States (column 6) 

Figure 5-7. Congener group profiles for cigarette tobacco from various 

countries. 

Source: Matsueda et al. (1994). 

reported by Matsueda et al. (1994)—6.3 pg TEQDF-WHO98/pack (5.5 pg I-TEQDF/pack)—is 

representative of cigarettes smoked in the United States, (b) CDDs/CDFs are not formed and the 

congener profile reported by Matsueda et al. (1994) is not altered during combustion of 

cigarettes, and (c) all CDDs/CDFs contributing to the TEQ are released from the tobacco during 

smoking. 

The second method of estimating is based on the assumption that the TEQ emission rates 

for a common Swedish brand of cigarette reported by Löfroth and Zebühr (1992) for mainstream 

smoke (0.96 pg TEQDF-WHO98/cigarette [0.9 pg I-TEQDF/cigarette]) and sidestream smoke (2.08 

pg TEQDF-WHO98/cigarette [1.96 pg I-TEQDF/cigarette]) are representative of the emission rates 

for U.S. cigarettes. For 2000, the two methods yield estimates of 0.11 g TEQDF-WHO98 (0.1 g I

TEQ) and 0.67 g TEQDF-WHO98 (0.63 g I-TEQDF). For 1995, the two methods yield estimates of 

0.2 g (on a TEQDF-WHO98 or I-TEQDF basis) and 1.48 g TEQDF-WHO98 (1.41 g I-TEQDF). For 

5-38
 

­
1987, the two methods yield estimates of 0.14 g TEQDF-WHO98 (0.12 g I-TEQDF basis) and 1.75 

g TEQDF-WHO98 (1.67 g I-TEQDF). 

For purposes of this report, the best estimates of annual emissions are assumed to be the 

average of the annual emissions estimated by the two methods for 2000, 1995, and 1987 (0.4 g, 

0.8 g, and 1 g TEQDF-WHO98 or I-TEQDF, respectively). These emissions are assigned a low 

confidence rating because the emission factor has a low confidence rating. Although these 

emission quantities are relatively small when compared with the emission quantities estimated 

for various industrial combustion source categories, they are significant because humans are 

directly exposed to cigarette smoke. 

5.6. PYROLYSIS OF BROMINATED FLAME RETARDANTS 

The pyrolysis and photolysis of brominated phenolic derivatives and polybrominated 

biphenyl ethers used as flame retardants in plastics (especially those used in electronic devices), 

textiles, and paints can generate considerable amounts of polybrominated dibenzo-p-dioxins 

(BDDs) and dibenzofurans (BDFs) (Watanabe and Tatsukawa, 1987; Thoma and Hutzinger, 

1989; Luijk et al., 1992). Watanabe and Tatsukawa (1987) observed the formation of BDFs from 

the photolysis of decabromobiphenyl ether. Approximately 20% of the decabromobiphenyl ether 

was converted to BDFs in samples that were irradiated with ultraviolet light for 16 hr. 

Thoma and Hutzinger (1989) observed the formation of BDFs during combustion 

experiments with polybutylene-terephthalate polymers containing 9 to 11% decabromodiphenyl 

ether. Maximum formation of BDFs occurred at 400 to 600°C, with a BDF yield of 16%. 

Although the authors 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. 

Luijk et al. (1992) studied the formation of BDDs/BDFs during the compounding and 

extrusion of decabromodiphenyl ether into high-impact polystyrene polymer at 275°C. Heptaand octa-BDF were formed during repeated extrusion cycles, and the yield of BDFs increased as 

a function of the number of extrusion cycles. 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. 

Insufficient data are available at this time from which to derive annual BDD/BDF 

emission estimates for this source. 

5.7. CARBON REACTIVATION FURNACES 

Granular activated carbon (GAC) is an adsorbent that is widely used to remove organic 

pollutants from wastewater and to treat finished drinking water at water treatment plants. 

Activated carbon is manufactured from the pyrolytic treatment of nut shells and coal (Buonicore, 

5-39
 
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 

(600:1,600 m2/g), combined with the extremely small pore diameter of the particles (15 to 25 

angstroms), increases the adsorption characteristics (Buonicore, 1992a). GAC eventually 

becomes saturated, and the adsorption properties significantly degrade. When saturation occurs, 

GAC usually must be discarded and replaced, 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). Large-scale 

regeneration operations, such as those used in industrial water treatment operations, typically use 

multiple-hearth furnaces. For smaller-scale operations, such as those used in municipal water 

treatment operations, fluidized-bed and infrared furnaces are used. Emissions are typically 

controlled by afterburners followed by water scrubbers (U.S. EPA, 1997a). 

The used GAC can contain compounds that are precursors to the formation of 

CDDs/CDFs during the thermal treatment process. A study by EPA (U.S. EPA, 1987a) 

measured precursor compounds in spent GAC that was used as a feed material to a carbon 

reactivation furnace tested during the National Dioxin Study. The total chlorobenzene content of 

the GAC ranged from 150 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. 

EPA has stack-tested two GAC reactivation furnaces for the emission of dioxin (U.S. 

EPA, 1987a; 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. EPA (U.S. 

EPA, 1997a) reported results of other testing performed at a county water facility in California 

during 1990. 

The industrial carbon reactivation plant processed 36,000 kg/day of spent GAC used in 

the treatment of industrial wastewater effluents. This facility was chosen for testing because it 

was considered to be representative of other facilities in the source category (U.S. EPA, 1987a). 

Spent carbon was reactivated in a multiple-hearth furnace, cooled in a water quench, and shipped 

back to primary chemical manufacturing facilities for reuse. The furnace was fired by natural gas 

and consisted of seven hearths arranged vertically in series. The hearth temperatures ranged from 

480 to 1,000°C. Air pollutant emissions were controlled by an afterburner, a sodium spray 

cooler, and an FF. Temperatures in the afterburner were about 930°C. The estimated I-TEQDF 

emission factor (treating nondetect values as zero) was 0.76 ng TEQDF-WHO98/kg (0.64 ng 

5-40
 
I-TEQDF/kg) carbon processed. The emission factor for total CDDs/CDFs was 58.6 ng/kg. 

Because analyses were performed only for 2,3,7,8-TCDD; 2,3,7,8-TCDF; OCDD; and OCDF 

and the congener groups, equivalent concentrations were assumed for all toxic and nontoxic 

congeners in each of the penta, hexa, and hepta congener groups. 

The second GAC reactivation facility tested by 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 chamber was fired by natural gas and consisted of a stoichiometrically balanced 

stream of fuel and oxygen. Combustion temperatures were about 1,038°C. Gases from the 

reactivation section and combustion chamber were directed through an acid gas scrubber and 

high-temperature afterburner prior to discharge from a stack. Although measurable 

concentrations of dioxin-like compounds were detected in the stack emissions, measurements of 

the individual CDD/CDF congeners were not performed; therefore, it was not possible to derive 

TEQ emission factors for this facility. 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/volume basis (ppt/v) and 0.006 to 0.28 ppt/v, respectively. All CDF 

congener groups were detected in the stack emissions even with the afterburner operating. Total 

CDFs emitted from the stack averaged 0.023 ppt/v. 

EPA (U.S. EPA, 1997a) reported a TEQ emission factor of 1.73 ng I-TEQDF/kg of carbon 

processed for the reactivation unit at a county water facility in California in 1990. The emission 

factor for total CDDs/CDFs was reported to be 47 ng/kg (i.e., similar to the total CDD/CDF 

emission factor of 58.6 ng/kg at the industrial GAC facility). Because congener-specific results 

were not reported, it was not possible to calculate the TEQDF-WHO98 emission factor. The report 

also did not provide the configuration and type of furnace tested; however, it did state that the 

emissions from the furnace were controlled by an afterburner and a scrubber. 

The industrial GAC reaction furnace test data indicate that an average of 0.64 ng 

I-TEQDF/kg of GAC may be released. The I-TEQDF emission rate for the reactivation unit at the 

county water treatment facility was 1.73 ng I-TEQDF/kg carbon. Low confidence ratings are 

given to these emission factors because only two GAC reactivation furnaces were stack tested, 

and not all congeners were analyzed at the industrial GAC facility. 

The mass of GAC that is reactivated annually in carbon reactivation furnaces is not 

known. However, a rough estimate, to which a low confidence rating is assigned, is the mass of 

virgin GAC shipped each year by GAC manufacturers. According to the U.S. Department of 

Commerce (U.S. DOC, 1990c), 48,000 metric tons of GAC were shipped in 1987. EPA reported 

that in 1990, water and wastewater treatment operations consumed 65,000 metric tons of GAC 

5-41
 
(U.S. EPA, 1995b, 1997a). The 1990 activity level was used in this document as a surrogate for 

the 1995 and 2000 activity levels. 

Applying the average TEQ emission factor of 1.2 ng (TEQDF-WHO98 or I-TEQDF) per kg 

of reactivated carbon for the two tested facilities to the estimates of potential GAC reactivation 

volumes yields annual release estimates of 0.06 g (TEQDF-WHO98 or I-TEQDF) in 1987 and 

0.08 g (TEQDF-WHO98 or I-TEQDF) in 1995 and 2000 (assuming that the activity level for 1990 is 

representative of the 1995 and 2000 activity levels). These emission estimates are assigned a low 

confidence rating because both the activity and emission factor estimates had low confidence 

ratings. 

5.8. KRAFT BLACK LIQUOR RECOVERY BOILERS 

Kraft black liquor recovery 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. Wood is broken down into two phases: a soluble phase 

containing primarily lignin, and an insoluble phase containing the pulp. The spent liquor (called 

black liquor) from the digester contains sodium sulfate and sodium sulfide, which the industry 

recovers 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% solids. The concentrated black liquor also contains 0.5 

to 4% chlorides by weight, which are recovered through combustion. The concentrated black 

liquor is sprayed into a Kraft black liquor recovery 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. The combustion gas is usually passed 

through an ESP that collects PM prior to being vented out the stack. The PM can be processed to 

further recover and recycle sodium sulfate (Someshwar and Pinkerton, 1992). 

In 1987, EPA stack-tested three Kraft black liquor recovery boilers for the emission of 

dioxin in conjunction with the National Dioxin Study (U.S. EPA, 1987a). The three sites tested 

by EPA were judged to be typical of Kraft black liquor recovery boilers at that time. During 

pretest surveys, two facilities were judged to have average potential for CDD/CDF emissions and 

one was judged to have high potential, based on the amount of chlorine found in the feed to these 

units. Dry-bottom ESPs controlled emissions from two of the boilers; a wet-bottom ESP 

controlled emissions from the third. The results of these tests included congener group 

concentrations but lacked measurement results for specific congeners other than 2,3,7,8-TCDD 

and 2,3,7,8-TCDF. 

5-42
 
NCASI (1995) provided congener-specific emission test results for six additional boilers 

tested during 1990 to 1993. Three boilers were of the direct contact type, and three were 

noncontact type. All were equipped with ESPs. The average congener and congener group 

emission factors are presented in Table 5-13 for the three facilities reported by EPA (U.S. EPA, 

1987a) and the six facilities reported by NCASI (1995). Figure 5-8 presents the average 

congener and congener group profiles based on the test results presented by NCASI (1995). 

The average TEQ emission factor, based on the data for the six NCASI facilities with 

complete congener data, was 0.028 ng TEQDF-WHO98/kg (0.029 ng I-TEQDF/kg) of black liquor 

solids, assuming nondetect values were zero, and 0.078 ng TEQ ­ DF-WHO98/kg (0.068 ng I

TEQDF/kg), assuming nondetect values were present at one-half the DL. This value is assumed to 

apply to all three reference years (1987, 1995, and 2000). The results for the three facilities 

reported by EPA were not used in the derivation of the TEQ emission factor because congenerspecific measurements for most 2,3,7,8-substituted congeners were not made in the study (U.S. 

EPA, 1987a). A medium confidence rating is assigned to those emission factors because they 

were derived from the stack testing of six Kraft black liquor recovery boilers that were judged to 

be fairly representative of technologies used at Kraft pulp mills in the United States. 

A 1995 survey of the industry indicated that 215 black liquor recovery boilers were in 

operation at U.S. pulp and paper mills. All but one of these boilers used ESPs for control of 

particulate emissions; the one unique facility used dual scrubbers. In addition, ESPs were 

reported to have been the predominant means of particulate control at recovery boilers for the 

past 20 years (letter dated October 8, 1998, from W. Gillespie, National Council of the Paper 

Industry for Air and Stream Improvement, Inc., to G. Schweer, Versar, Inc.). 

The amounts of black liquor solids burned in Kraft black liquor recovery boilers in the 

United States during 1987 and 1995 were 69.8 million metric tons and 80.8 million metric tons, 

respectively (American Paper Institute, 1992; American Forest and Paper Association, 1997). 

These activity level estimates are assigned a high confidence rating because they are based on 

comprehensive industry survey data. Combining the emission factors derived above with the 

activity level estimates for 1987 and 1995 yields estimated annual emissions from this source of 

approximately 2 g (TEQDF-WHO98 or I-TEQDF) in 1987 and 2.3 g (TEQDF-WHO98 or I-TEQDF) in 

1995. These emission estimates are assigned a medium confidence rating because the emission 

factor have a medium confidence rating. 

For 2000, NCASI provided estimates of activity levels for Kraft recovery furnaces and 

Kraft lime kilns and CDD/CDF releases, including emissions from 11 Kraft recovery furnaces 

and four Kraft lime kilns (Gillespie, 2002). The activity levels were reported to be 90.7 million 

5-43
 
Table 5-13. CDD/CDF mean emission factors (ng/kg feed) for black liquor 

recovery boilers 

Congener 

U.S. EPA (1987a) 

(3 facilities) 

NCASI (1995) 

(6 facilities) 

Nondetect set to 

zero 

Nondetect set to 1⁄2 

detection limit 

Nondetect set to 

zero 

Nondetect set to 1⁄2 

detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0 

NR 

NR 

NR 

NR 

NR 

4.24 

0.04 

NR 

NR 

NR 

NR 

NR 

4.24 

0 

0 

0.001 

0.003 

0.006 

0.108 

1.033 

0.016 

0.016 

0.018 

0.015 

0.019 

0.135 

1.054 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.04 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

0.35 

0.06 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

0.35 

0.04 

0.03 

0.033 

0.007 

0.012 

0.005 

0.01 

0.024 

0 

0.113 

0.049 

0.036 

0.037 

0.022 

0.021 

0.016 

0.021 

0.035 

0.014 

0.13 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

0.21 

0.27 

0.8 

2.05 

4.24 

0.95 

0.64 

1.16 

1.05 

0.35 

0.36 

0.35 

1.02 

2.05 

4.24 

1 

0.77 

1.2 

1.05 

0.35 

0.106 

0.013 

0.104 

0.252 

1.033 

1.27 

0.37 

0.102 

0.024 

0.113 

0.123 

0.059 

0.122 

0.279 

1.054 

1.275 

0.376 

0.109 

0.038 

0.13 

Total I-TEQDF 

Total TEQDF-WHO98 

0.10a 

0.10a 

0.15a 

0.16a 

0.029 

0.028 

0.065 

0.072 

Total CDD/CDF 11.72 12.39 3.39 3.57 

a

Estimate based on the measured data for 2,3,7,8-TCDD; 2,3,7,8-TCDF; OCDD; and OCDF and congener group

 emissions (i.e., for the penta-, hexa-, and hepta-CDD and CDFs, it was assumed that the measured emission factor

 within a congener group was the sum of equal emission factors for all congeners in that group, including non­

 2,3,7,8-substituted congeners). 

NR = Not reported 

metric tons for Kraft recovery furnaces and 13 million metric tons for Kraft lime kilns. These 

activity level estimates are assigned a high confidence rating because they are based on 

comprehensive industry survey data. Emission factors were taken from “NCASI Handbook of 

Chemical Specific Information for SARA Section 313 Form R Reporting.” The factors provided 

5-44
 
Ratio (me an cong e ne r g roup emission factor / total CDD/CDF emission factor)
 

0a 0.1 0.2 0.3 0.4 0.5
 

TCDD 

Pe CDD 

HxC DD 

Hp C DD 

OC DD 

TCDF 

Pe CDF 

HxC DF 

Hp C DF 

OC DF 

Figure 5-8. Congener and congener group profiles for air emissions from 

Kraft black liquor recovery boilers (nondetect set equal to zero). 

Source: NCASI (1995). 

5-45
 
in this handbook were compiled from valid test data supplied to NCASI by a variety of sources, 

including NCASI member companies who had performed the tests in response to a regulatory 

program. They are assigned a high confidence rating because they are based on a comprehensive 

survey of stack emissions. Congener-specific CDD/CDF TEQ emission factors were provided 

for both source categories (Table 5-14). Using the congener-specific emission factors and the 

activity levels provided above, NCASI estimated CDD/CDF TEQDF-WHO98 emissions for each 

congener (Table 5-14) and reported total emissions of 0.75 g TEQDF-WHO98/yr and 6.9e!5 g 

TEQDF-WHO98/yr for Kraft recovery furnaces and Kraft lime kilns, respectively. This 2000 

emissions estimate has a high confidence rating because both the emission factor and activity 

level are rated as high confidence. 

5.9. OTHER IDENTIFIED SOURCES 

Several manufacturing processes are identified as potential sources of CDD/CDF 

formation because the processes use chlorine-containing components or involve application of 

high temperatures. However, no testing of emissions from these processes has been performed in 

the United States, and only minimal emission rate information has been reported for these 

processes in other countries. Therefore, these sources are rated as Category E sources, meaning 

their emissions cannot be quantified. 

Burning of candles. Schwind et al. (1995) analyzed the wicks and waxes of uncolored 

candles as well as the fumes of burning candles for CDDs/CDFs, total chlorophenol, and total 

chlorobenzene content. The results, presented in Table 5-15, show that beeswax contained the 

highest levels of CDDs/CDFs and total chlorophenols. In contrast, the concentration of total 

chlorobenzenes in stearin wax was higher than that in paraffin or beeswax by a factor of 2 to 3. 

The concentrations of the three analyte groups were significantly lower in the wicks than in the 

waxes. Emissions of CDDs/CDFs from all three types of candles were very low during burning. 

In fact, comparison of the emission factor with the original CDD/CDF concentrations in the wax 

indicates a net destruction of the CDDs/CDFs originally present in the wax. Information on the 

activity level is lacking, therefore, no estimate of environmental release can be made at this time. 

Glass manufacturing. Annual emissions of less than 1 g I-TEQDF/yr have been 

estimated for glass manufacturing facilities in the Netherlands (Bremmer et al., 1994) and the 

United Kingdom (Douben et al., 1995). Glass is manufactured by heating a mixture of sand and, 

depending on the type of glass, lime, sodium carbonate, dolomite, clay, or feldspar to a 

temperature of 1,400 to 1,650°C. In addition, various coloring and clarifying agents may be 

added. Chlorine enters the process as a contaminant (NaCl) in sodium carbonate (Bremmer et al. 

1994). However, the emission factors used by Bremmer et al. and Douben et al. were not 

reported. Umweltbundesamt (1996) reported relatively low emission factors (approximately 

0.002 and 0.007 ng I-TEQDF/kg) for two glass manufacturing facilities in Germany. 

5-46
 
Table 5-14. CDD/CDF TEQ emission factors and emission estimates from Kraft 

recovery furnaces and Kraft lime kilns 

Congener 

Kraft recovery furnaces Kraft lime kilns 

TEQDF-WHO98 

(ng/lb BLS) 

Emissions 

(ng/yr) 

TEQDF-WHO98 

(ng/lb CaO) 

Emissions 

(ng/yr) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

1,2,3,4,6,7,8,9-OCDD 

0.00e+00 

0.00e+00 

0.00e+00 

2.00e!04 

5.00e!04 

4.90e!04 

1.42e!04 

0.00e+00 

0.00e+00 

0.00e+00 

3.33e+07 

8.31e+07 

8.15e+07 

2.36e+07 

0.00e+00 

0.00e+00 

0.00e+00 

1.00e–04 

0.00e+00 

2.80e!04 

2.56e!04 

0.00e+00 

0.00e+00 

0.00e+00 

2.60e+03 

0.00e+00 

7.27e+03 

6.65e+03 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

1,2,3,4,6,7,8,9-OCDF 

5.00e!04 

1.00e!04 

1.50e!03 

4.00e!04 

2.00e!04 

0.00e+00 

4.00e!04 

6.00e!05 

0.00e+00 

2.60e!05 

8.31e+07 

1.66e+07 

2.49e+08 

6.65e+07 

3.33e+07 

0.00e+00 

6.65e+07 

9.98e+06 

0.00e+00 

4.32e+06 

8.00e!04 

1.00e!04 

0.00e+00 

9.00e!04 

2.00e!04 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

2.08e+04 

2.60e+03 

0.00e+00 

2.34e+04 

5.20e+03 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

BLS = Black liquor solids 

CaO = Calcium oxide 

Table 5-15. CDD/CDF concentrations in candle materials and emissions 

Wax 

material 

Candle 

component 

Concentration Emission factor 

CDD/CDF 

(ng I-TEQDF/kg) 

Total 

chlorophenols 

(:g/kg) 

Total 

chlorobenzenes 

(:g/kg) 

CDD/CDF 

(ng I-TEQDF/kg 

burnt wax) 

Paraffin Wax 0.59 14.8 130 0.015 

Stearin Wax 1.62 32.3 330 0.027 

Beeswax Wax 10.99 256 120 0.004 

Paraffin Wick 0.18 1.23 0.67 NR 

Stearin Wick 0.12 0.94 0.34 NR 

Beeswax Wick 0.08 0.74 0.35 NR 

Source: Schwind et al. (1995). 

5-47
 
Lime kilns. Annual emissions from lime kilns in Belgium and the United Kingdom have 

been reported by Wevers and De Fre (1995) and Douben et al. (1995), respectively. However, 

the emission factors used to generate those estimates were not provided. Umweltbundesamt 

(1996) reported low emissions (0.016 to 0.028 ng I-TEQDF/kg) during tests at two lime kilns in 

Germany. 

Ceramics and rubber manufacturers. Douben et al. (1995) estimated annual emissions 

from ceramic manufacturers and rubber manufacturers in the United Kingdom. Lexen et al. 

(1993) had previously detected high concentrations of CDDs/CDFs in emissions from a ceramic 

manufacturer in Sweden that occasionally glazed ceramics by volatilization of sodium chloride in 

a coal-fired oven. Lexen et al. (1993) also detected high pg/L levels of I-TEQDF in the scrubber 

water from the vulcanization process at a Swedish rubber manufacturer. 

5-48
 
6. COMBUSTION SOURCES OF CDDs/CDFs: MINIMALLY CONTROLLED 

AND UNCONTROLLED COMBUSTION SOURCES 

This chapter discusses combustion sources of CDDs/CDFs that have some (in the case of 

combustion of landfill gas) or no post-combustion pollution control equipment for conventional 

pollutant emissions. 

6.1. COMBUSTION OF LANDFILL GAS 

6.1.1. Emissions Data 

Although no data could be located on levels in untreated landfill gas, several studies have 

reported detecting CDDs/CDFs in the emissions resulting from the combustion of landfill gas. 

Only one study of CDD/CDF emissions from a landfill flare has been reported for a U.S. landfill 

(CARB, 1990c). The TEQDF-WHO98 and I-TEQDF emission factor calculated from the results of 

this study is approximately 2.4 ng TEQ/m3 landfill gas combusted. The congener-specific results 

of this study are presented in Table 6-1. Figure 6-1 presents the CDD/CDF congener emission 

profile based on these emission factors. Bremmer et al. (1994) reported a lower emission factor, 

0.4 ng I-TEQDF/m3, from the incineration of untreated landfill gas in a flare at a facility located in 

the Netherlands. No congener-specific emission factors were provided. The average TEQ 

emission factor for the CARB and Bremmer et al. studies is 1.4 ng I-TEQDF/m3 landfill gas 

combusted. 

Umweltbundesamt (1996) reported even lower TEQ emission factors for landfill gas 

burned in engines or boiler mufflers rather than in a flare. The reported results for 30 engines 

and mufflers tested in Germany ranged from 0.001 to 0.28 ng I-TEQDF/m3, with most values 

below 0.1 ng I-TEQDF/m3. However, Bremmer et al. (1994) reported an emission factor of 0.5 ng 

I-TEQDF/m3

 from a landfill gas-fired engine in the Netherlands. 

6.1.2. Activity Level Information 

In 1996 EPA promulgated emission standards and guidelines to control emissions of 

landfill gas from existing and future landfills under the Clean Air Act (Federal Register, 1996b). 

Those regulations require the largest landfills in the United States (on the basis of design 

capacity) to periodically measure and determine their annual emissions of landfill gas. Landfills 

that emit more than 50 metric tons of nonmethane organic compounds (NMOC) annually must 

collect landfill gas and reduce its NMOC content by 98% weight through the use of a control 

device. 

6-1
 

Table 6-1. CDD/CDF emission factors for a landfill flare 

Congener/congener group 

Mean facility emission factora 

(ng/m3 gas combusted) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.02 

0.09 

0.07 

0.07 

0.26 

0.76 

4.41 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

14.07 

0.39 

1.14 

1.46 

0.42 

0.11 

0.68 

1.22 

0.07 

0.64 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

5.68 

20.20 

2.39 

2.43 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

Total CDD/CDF NR a

Assumes heat content of 1.86e+07 J/m3 for landfill gas (Federal Register, 1996a). 

NR = Not reported 

Source: CARB (1990c). 

6-2
 

Ratio (congener emission factor/total 2378-CDD/CDF emission factor) 

0 0.1 0.2 0.3 0.4 0.5

Figure 6-1. Congener profile for landfill flare air emissions. 

Source: CARB (1990c). 

6.1.2.1. Activity Levels for 1987 and 1995 

EPA estimated that when the 1996 regulations were implemented, the controls would 

reduce annual NMOC emissions from existing landfills by 77,600 metric tons. The cost analysis 

supporting this rulemaking based control device costs on open flares because flares are 

applicable to all the regulated facilities. Assuming that the mass reduction would be achieved by 

the use of flares, the corresponding volume of landfill gas burned would be approximately 14 

billion m3/yr. The calculation was based on an assumed default NMOC concentration in landfill 

gas of 1,532 ppmv and a conversion factor of 3.545 mg/m3 NMOC per 1 ppmv NMOC (Federal 

Register, 1993b). 

6-3
 

0.6
Of the approximately 312 landfills that were affected by the promulgation of the emission 

standards and guidelines in 1996, EPA estimated that more than 100 had some form of collection 

or control system (or both) in place in 1991 (Federal Register, 1991c). Thus, a rough 

approximation of the volume of landfill gas combusted in 1995 was 4.7 billion m3/yr (or 33% of 

the future expected 14 billion m3/yr reduction). This estimate is similar to the 2 to 4 billion m3 

landfill gas estimated by the Energy Information Administration (EIA, 1994) as collected and 

consumed for energy recovery purposes in 1992. 

EIA (1992) estimated that between 0.9 and 1.8 billion m3 of landfill gas were collected 

and burned in 1990 for energy recovery purposes. Because there were no specific data available 

for 1987, EPA assumed that the mean of this range, 1.35 billion m3, would serve as an 

approximate estimate of the volume of landfill gas combusted in 1987. 

6.1.2.2. Activity Level for 2000 

According to the EPA 2001 inventory of greenhouse gas emissions, approximately 7.7 

billion m3 of landfill gas were combusted in 2000 through 477 landfill flares (average of 16.5 

million m3 of landfill gas per flare). As of 2003, there were more than 1,000 landfill flares in the 

United States (e-mails dated February 28, 2003, and March 7, 2003, from B. Guzzone, U.S. EPA, 

to K. Riley, Versar). Assuming that the amount of landfill gas combusted through the 477 

landfill flares inventoried is representative of the landfill gas combusted through these 1,000+ 

flares, approximately 16 billion m3 of landfill gas were combusted in the United States through 

flares in 2000. 

6.1.3. Emission Estimates 

The limited emission factor data that are available were judged inadequate for developing 

national emission estimates that could be included in the national inventory. However, a 

preliminary estimate of the potential annual TEQ releases from landfills can be obtained using 

the estimated volume of combusted gas and the available emission factors. Combining the 

estimates of landfill gas volume that was combusted (1.35 billion m3

 in 1987, 4.7 billion m3 in 

1995, and 16 billion m3 in 2000) with the emission factor of 1.4 ng I-TEQDF/m3 of flarecombusted gas yields annual emission estimates of 1.9, 6.6, and 22 g I-TEQDF/m3

 for 1987, 1995, 

and 2000, respectively. These estimates should be regarded as preliminary indications of 

possible emissions from this source; further testing is needed to confirm the true magnitude of 

those emissions. 

6-4
 

6.2. ACCIDENTAL FIRES 

Accidental fires in buildings and vehicles are uncontrolled combustion processes that, 

because of poor combustion conditions, typically result in relatively high emissions of 

incomplete combustion products (Bremmer et al., 1994), which can include CDDs and CDFs. 

Polyvinyl chloride (PVC) building materials and furnishings, chloroparaffin-containing textiles 

and paints, and other chlorinated organic compound-containing materials appear to be the 

primary sources of the chlorine (Rotard, 1993). Although the results of several studies have 

demonstrated the presence of CDD/CDF concentrations in soot deposits and residual ash from 

such fires, few direct measurements of CDDs/CDFs in the fumes or smoke of fires have been 

reported. The results of some of those studies are described below, and an evaluation of the 

available data follows. 

6.2.1. Soot and Ash Studies 

Christmann et al. (1989a) analyzed the soot formed during combustion and pyrolysis of 

pure PVC and PVC cable sheathings in simple laboratory experiments designed to mimic the 

conditions of fires. For the combustion experiments, 2 g of a PVC sample were incinerated with 

a laboratory gas burner. The combustion products were collected on the inner walls of a cooled 

gas funnel placed above the sample. For the pyrolysis experiments, about 50 mg of the sample 

were placed in a quartz tube and heated to about 950°C for 10 min in either an air atmosphere or 

a nitrogen atmosphere. The combustion experiments yielded CDD/CDF concentrations in soot 

of 110 μg I-TEQDF/kg for a low-molecular-weight PVC, 450 μg I-TEQDF/kg for a highmolecular-weight PVC, and 270 μg I-TEQDF/kg for PVC cable. The pyrolysis experiments in the 

air atmosphere yielded lower CDD/CDF concentrations in soot: 24.4 μg I-TEQDF/kg for a lowmolecular-weight PVC, 18.7 μg I-TEQDF/kg for a high-molecular-weight PVC, and up to 41 μg ITEQDF/kg for PVC cable. 

In general, more CDFs than CDDs were formed. The lower-chlorinated CDF congeners 

were dominant in the combustion experiments; however, the HpCDF and OCDF congeners were 

dominant in the pyrolysis experiments. No CDDs/CDFs were detected in pyrolysis experiments 

under a nitrogen atmosphere. Also, no CDDs/CDFs were detected when chlorine-free 

polyethylene samples were subjected to the same combustion and pyrolysis conditions. 

Deutsch and Goldfarb (1988) reported finding CDD/CDF concentrations ranging from 

0.04 to 6.6 μg/kg in soot samples collected after a 1986 fire in a State University of New York 

lecture hall. The fire consumed or melted plastic furnishings and cleaning products containing 

chlorine, wood, and paper. 

In a study that analyzed 200 ash and soot samples from sites of accidental fires in which 

PVC was involved (Funcke et al., 1988, as reported in Bremmer et al., 1994; Rotard, 1993), 

6-5
 

CDDs/CDFs were detected in more than 90% of the samples at concentrations in the ng ITEQDF/kg to μg I-TEQDF/kg range. Fires involving the combustion of materials containing 

relatively large amounts of PVC and other chlorinated organic substances resulted in the highest 

levels of CDDs/CDFs, with concentrations ranging from 0.2 to 110 μg I-TEQDF/kg residue. 

Thiesen et al. (1989) analyzed residues from surfaces of PVC-containing materials that 

were partially burned during accidental fires at sites in Germany that manufactured or stored 

plastics. CDD/CDF concentrations in residues were reported as 0.5 μg I-TEQDF/kg for soft PVC, 

4.6 μg I-TEQDF/kg for PVC fibers, and 28.3 μg I-TEQDF/kg for a hard PVC. The ratio of total 

CDFs to total CDDs in the three samples ranged from 4:1 to 7:1. The dominant 2,3,7,8­

substituted CDF and CDD congeners in all three samples were 1,2,3,4,6,7,8-HpCDF and 

1,2,3,4,6,7,8-HpCDD. 

In an accidental fire at a Swedish carpet factory in 1987, 200 metric tons of PVC and 500 

metric tons of PVC-containing carpet were burned. Marklund et al. (1989) analyzed snow 

samples up to 1,500 m downwind from the fire site and found CDD/CDF concentrations in the 

top 2 cm ranging from 0.32 μg I-TEQDF/m2 at 10 m from the site to 0.01 μg I-TEQDF/m2 at 1,500 

m. Because of an atmospheric inversion and very light wind at the time of the fire, the smoke 

from the fire remained close to the ground. The soot deposited onto the snow was thus assumed 

to be representative of the soot generated and released from the fire. Wipe samples of soot from 

interior posts of the plant that were 5 and 20 m from the fire contained Eadon TEQ 

concentrations of 0.18 and 0.05 μg/m2, respectively. On the basis of these deposition 

measurements, the investigators estimated total CDD/CDF emissions from the fire to be less than 

3 mg I-TEQDF. 

Carroll (1996) estimated a soot-associated CDD/CDF emission factor for the Swedish 

carpet factory fire (i.e., not including volatile emissions) of 28 to 138 ng I-TEQDF/kg PVC burned 

using the following assumptions: the PVC carpet backing was one-half the weight of the carpet, 

the carpet backing contained 30% by weight PVC resin, and 20 to 100% of the PVC and PVC 

carpet backing present in the warehouse actually burned. Using the results of wipe samples 

collected at downwind distances of up to 6,300 m, Carroll (1996) also estimated a similar sootassociated emission factor (48 to 240 ng I-TEQDF/kg of PVC burned) for a fire at a plastics 

recycling facility in Lengerich, Germany. 

Fiedler et al. (1993) presented a case study of CDD/CDF contamination and associated 

remedial actions taken at a kindergarten in Germany following a fire that destroyed parts of the 

roof, windows, and furnishings. Soot collected from the building contained CDDs/CDFs at a 

concentration of 45 μg I-TEQDF/kg (15 μg I-TEQDF/m2

). The study authors attributed the 

CDDs/CDFs detected to the combustion of plastic and wooden toys, floors, and furnishings; 

however, no information was provided on the quantities of those materials. 

6-6
 
Fiedler and Lindert (1998) presented results of soot sampling following a serious fire at 

the airport in Düsseldorf, Germany. Polystyrene sheets and PVC-coated cables were involved in 

the fire, together with PCB-containing condensers (bulbs). Surface wipe samples contained up to 

0.33 μg I-TEQDF/m2. Concentrations in soot ranged from 7 to 130 μg I-TEQDF/kg. 

Concentrations of polybrominated dibenzo-p-dioxins and dibenzofurans were detected in soot at 

concentrations as high as 0.9 mg/kg soot. 

Wichmann et al. (1993, 1995) measured the CDD/CDF content of ash and debris and 

deposited surface residues that resulted from experimental test burns of two cars (a 1974 Ford 

Taurus [old car] and a 1988 Renault Espace [new car]), one subway car, and one railway coach in 

a tunnel in Germany. On the basis of measurements obtained from the sampled ash and debris 

and from soot collectors placed at regular intervals up to 420 m downwind of the burn site, the 

total estimated amount of CDDs/CDFs in the ash/debris and tunnel surface residues from each 

vehicle burn experiment was 0.044 mg I-TEQDF for the 1974 model car, 0.052 mg I-TEQDF for 

the 1988 model car, 2.6 mg I-TEQDF for the subway car, and 10.3 mg I-TEQDF for the railway 

coach. For each vehicle burn experiment, the mass of TEQ in tunnel surface residue exceeded 

the mass in ash and debris: 73 to 89% were accounted for by the tunnel surface residues and 11 

to 27% by ash and debris. The average CDD/CDF content of the ash and debris from each 

experimental burn was: 0.14 μg I-TEQDF/kg for the new car, 0.3 μg I-TEQDF/kg for the old car, 

3.1 μg I-TEQDF/kg for the subway car, and 5.1 μg I-TEQDF/kg for the railway coach. 

6.2.2. Fume and Smoke Studies 

Merk et al. (1995) collected fume and smoke generated during the burning of 400 kg of 

wood and 40 kg of PVC in a building (4,500 m3 volume) over a 45-min period. The sampling 

device consisted of dual glass fiber filters to collect particles greater than 0.5 μm followed by a 

polyurethane foam filter to collect vapor-phase CDDs/CDFs. The particulate phase (particles 

greater than 0.5 μm diameter) and the gas phase showed the same congener pattern: decreasing 

concentration with increasing degree of chlorination, thus indicating no preferential sorption of 

higher-chlorinated congeners to smoke particulates. However, the CDDs/CDFs found in the gas 

phase (about 5 ng I-TEQDF/m3) accounted for more than 90% of the detected CDDs/CDFs. The 

authors also reported that the soot deposited from this fire onto a 1 m2 aluminum sheet resulted in 

surface contamination of 0.05 μg I-TEQDF/m2. 

Although it was stated in Merk et al. (1995) that the building was “closed,” subsequent 

communication with one of the coauthors (telephone conversation on September 29, 1998, 

between Karl-Werner Schramm, GSF-Institute of Ecological Chemistry, and Greg Schweer, 

Versar, Inc.) clarified that a “gas cleaning” system was in operation. Because a ventilation 

system was in operation, there was likely some loss of vapor-phase CDDs/CDFs from the hall; 

6-7
 

therefore, the deposits (from particulate deposition and vapor-phase condensation) on the test 

aluminum plate may not have reflected total CDD/CDF formation during the fire. 

Dyke and Coleman (1995) reported a fourfold increase in CDD/CDF TEQ concentrations 

in the ambient air during “bonfire night” (an annual event held on November 5 during which it is 

customary to set off fireworks and have bonfires) in Oxford, England. Air concentrations before 

and after bonfire night ranged from 0.15 to 0.17 pg I-TEQDF/m3. The air concentration during 

bonfire night was 0.65 pg I-TEQDF/m3

. The dominant congeners in all samples were the heptaand octa-CDDs. The study was not designed to collect data that would enable calculation of an 

emission rate or to differentiate the relative importance of the various materials combusted. 

However, the results do indicate that open burning of materials likely to be combusted in 

accidental fires (with the exception of fireworks) results in the release of CDDs and CDFs. 

6.2.3. Data Evaluation 

6.2.3.1. Structural Fires 

6.2.3.1.1. Emissions data. Only limited emissions data for structural fires were located. Most 

of the studies obtained involved situations (field and laboratory) where relatively high loadings 

of PVC or plastics were combusted. The effects of different mixes of combusted materials, 

oxygen supplies, building configurations, and durations of burn, for example, that are likely to 

occur or be found in accidental fires cannot be accounted for by the factors that can be derived 

from these studies. Also, most of the studies addressed only soot or ash residues and did not 

address potential volatile emissions of CDDs/CDFs, which, according to Merk et al. (1995), may 

represent 90% of the CDDs/CDFs generated during the burning of PVC. 

Two reports (Carroll, 1996; Thomas and Spiro, 1995) attempted to quantify CDD/CDF 

emissions from U.S. structural fires, and Lorenz et al. (1996) estimated emissions from structural 

fires in Germany. 

Carroll (1996) estimated the total CDD/CDF content of soot and ash generated from the 

358,000 residential fires in the United States in 1993 (as reported in U.S. DOC, 1995b). Detailed 

estimates were developed of the PVC content of items in typical homes, including plumbing, 

wiring, siding and windows, wallpaper, blinds and shades, and upholstery. The typical 

percentage of PVC burned in household fires was assumed to be 9.5%. Extrapolating to all 

358,000 one- to two-family unit fires yielded an annual mass of 2,470 metric tons PVC burned. 

Carroll then developed TEQ emission factors from the results of Thiesen et al. (1989) and 

Marklund et al. (1989). The estimated CDD/CDF content ranged from 0.47 to 22.8 g I-TEQDF, 

with 0.07 to 8.6 g I-TEQDF in soot and 0.4 to 14.2 g I-TEQDF in ash. A soot emission factor (i.e., 

grams of soot produced per gram of PVC combusted) was derived from the investigator’s 

assumptions regarding the surface area of the soot collection funnel used by Christmann et al. 

6-8
 

(1989b) and the soot deposition rate on that funnel. These I-TEQDF emission factors were then 

applied to the estimated 2,470 metric tons of PVC burned annually in one- to two-family unit 

residential fires to obtain estimates of the annual mass of TEQ that would be found in the soot 

and ash of residential fires (0.48 to 22.8 g I-TEQDF/yr). The average emission per fire is thus 1.3 

to 64 μg I-TEQDF. 

Thomas and Spiro (1995) estimated that 20 g I-TEQDF may be released annually to air 

from structural fires. This estimate assumes an emission factor of 4 ng I-TEQDF/kg material 

combusted (i.e., the emission rate for “poorly” controlled wood combustion), a material 

combustion factor of 6,800 kg per fire, and 688,000 structural fires per year. The average 

emission per fire is thus 29 μg I-TEQDF. 

Lorenz et al. (1996) estimated annual generation of CDDs/CDFs in Germany using data 

on the number of residential and industrial/commercial structural fires coupled with data on 

CDD/CDF content in soot and ash residues remaining after fires. The potential annual I-TEQDF 

generation was estimated to be 78 to 212 g. 

Using the emissions data estimated by Carroll (1996) and Thomas and Spiro (1995) 

provides an average emission factor of 32 μg I-TEQ/fire. 

6.2.3.1.2. Activity level information. In 1987, there were approximately 2,330,000 fires in the 

Unites States, of which approximately 745,600 (32%) were structural fires (FEMA, 1997). In 

1995, approximately 574,000 structural fires were reported in the United States. Of these, 

426,000 were reported to be in residential structures, including 320,000 in one- to two-family 

units, 94,000 in apartments, and 12,000 in other residential settings. The types of structures for 

the remaining 148,000 fires were public assembly, 15,000; educational, 9,000; institutional, 

9,000; stores and offices, 29,000; special structures, 29,000; storage, 39,000; and industry, utility, 

and defense, 18,000. The latter two categories may be underreported, as some incidents were not 

recorded because they were handled by private fire brigades or fixed suppression systems (U.S. 

DOC, 1997). For 2000, the National Fire Data Center estimated that approximately 1,708,000 

fires occurred in the United States, of which approximately 512,400 (30%) were structural fires 

(FEMA, 2001). 

6.2.3.1.3. Emission estimates. The limited data available on structural fires were judged 

inadequate for developing national emission estimates. This conclusion was also reached for 

national emission inventories developed for the Netherlands (Bremmer et al., 1994) and the 

United Kingdom (U.K. Department of the Environment, 1995). However, preliminary estimates 

were calculated by combining the average emission factor of 32 μg I-TEQ/fire and the number of 

structural fires in the Unites States (745,600 in 1987; 426,000 in 1995; and 512,400 in 2000), 

6-9
 

yielding an annual release of 24 g I-TEQDF in 1987, 14 g I-TEQDF in 1995, and 16 g I-TEQDF in 

2000. Confidence in these estimated emissions is very low because of the numerous assumptions 

employed in their derivation. If the conclusion by Merk et al. (1995) is assumed to be correct, 

that 90% of the CDDs/CDFs formed in fires are in the gaseous phase rather than particulate 

phase, and it is also assumed that the estimates by Carroll (1996) and Thomas and Spiro (1995) 

did not totally account for volatile emissions, then the total CDD/CDF emissions estimated by 

Carroll and Thomas and Spiro may be underestimates. Further testing is needed to confirm the 

true magnitude of these releases. 

6.2.3.2. Vehicle Fires 

As with structural fires, the limited data available on vehicle fires were judged inadequate 

for developing national emission estimates that could be included in the national inventory. 

However, a preliminary estimate of the range of potential CDD/CDF emissions that may result 

from vehicle fires can be calculated using the results reported by Wichmann et al. (1993, 1995) 

for controlled vehicle fires in a tunnel (0.044 mg I-TEQDF for an old car to 2.6 mg I-TEQDF for a 

subway car). Although Wichmann et al. did not measure volatile CDDs/CDFs (which were 

reported by Merk et al., 1995, to account for the majority of CDDs/CDFs formed during a fire), 

the study was conducted in a tunnel, and it is likely that a significant fraction of the volatile 

CDDs/CDFs sorbed to tunnel and collector surfaces and were thus measured as surface residues. 

The number of vehicle fires reported in the United States was approximately 561,530 in 

1987 (FEMA, 1997), 406,000 in 1995 (U.S. DOC, 1997), and 341,600 in 2000 (FEMA, 2001). 

If it is assumed that 99% of those fires involved cars and trucks (the approximate percentage of 

all U.S. motor vehicles that are in-service cars and trucks; U.S. DOC, 1995b) and that the 

applicable emission rate is 0.044 mg I-TEQDF per incident, then the annual TEQ formation was 

24.4 g I-TEQDF for 1987, 17.7 g I-TEQDF for 1995, and 14.9 g I-TEQDF for 2000. The emission 

factor of 2.6 mg I-TEQDF/fire is assumed to be applicable to the remaining 1% of vehicle fires, 

thus yielding emissions of 14.6 g I-TEQDF/yr for 1987, 10.6 g I-TEQDF/yr for 1995, and 8.9 g ITEQDF/yr for 2000. Total TEQ annual emissions for 1987, 1995, and 2000 were roughly 

estimated to have been 39, 28.3, and 23.8 g I-TEQDF/yr, respectively. These estimates should be 

regarded as preliminary indications of possible emissions from this source category; further 

testing is needed to confirm the true magnitude of these emissions. 

6-10
 

6.3. LANDFILL FIRES 

6.3.1. Emissions Data 

In the late 1980s, two serious fires occurred in landfills near Stockholm, Sweden. The 

first fire was in a large pile of refuse-derived fuel. Using measurements of chlorobenzenes in the 

air emissions, it was estimated that 50 to 100 kg of chlorobenzenes were released. CDD/CDF 

emissions were estimated to be several tens of grams, on the assumption that the ratio of 

CDDs/CDFs to chlorobenzenes in landfill fire emissions is similar to the ratio observed in stack 

gases of municipal waste combustors. To measure releases in connection with the second fire, 

which occurred at a large conventional landfill, birch leaves were collected from trees close to 

the fire and at distances up to 2 km downwind of the fire, as well as from nearby areas not 

affected by smoke from the fire. The discharge of CDDs/CDFs necessary to cause the 

concentrations measured on the leaves was estimated to be several tens of grams (Persson and 

Bergström, 1991). 

In response to these incidents, Persson and Bergström (1991) also measured CDD/CDF 

emissions from experimental fires designed to simulate surface landfill fires and deep landfill 

fires. The experiments used 9-month-old domestic waste. The tests showed no significant 

difference in CDD/CDF content of the fire gases produced by the simulated surface and that of 

the deep fires. The average CDD/CDF emission rate was reported to be 1 μg Nordic TEQ/kg 

waste burned. 

Persson and Bergström (1991) and Bergström and Björner (1992) estimated annual 

CDD/CDF Nordic TEQ emissions in Sweden from landfill fires to be 35 g. The estimate was 

based on the emission rate of 1 μg Nordic TEQ/kg waste burned, an assumed average density of 

landfill waste of 700 kg/m3, an assumed waste burn of 150 m3 for each surface landfill fire (167 

fires in Sweden per year), and an assumed waste burn of 500 m3 for each deep landfill fire (50 

fires in Sweden per year). The estimates of waste burn mass for each type of fire were the 

average values obtained from a survey of 62 surface fires and 25 deep fires. The estimated 

number of fires per year was based on the results of a survey of all Swedish municipalities for 

fires reported during 1988 and 1989. In 1991, Sweden had an estimated 400 municipal landfills 

(Persson and Bergström, 1991). 

Ruokojärvi et al. (1995) measured ambient air concentrations of CDDs/CDFs in the 

vicinity of real and experimental landfill fires in Finland. The most abundant toxic congeners 

were the hepta- and octa-CDDs and the penta-, hepta-, and octa-CDFs. The highest contributors 

to the measured TEQ were 1,2,3,7,8-PeCDD and 2,3,4,7,8-PeCDF. In Finland, annual 

CDD/CDF emissions from landfill fires are estimated to be 50 to 70 g Nordic TEQ (Aittola, 

1993, as reported in Ruokojärvi et al., 1995). 

6-11
 
6.3.2. Activity Level Information and Emission Estimates 

Although no U.S. monitoring studies are available, an emission factor similar to the 

Swedish emission factor would be expected for the United States because the content of 

municipal waste in the United States and Sweden are expected to be similar. Because no data 

could be located on characterization of landfill fires in the United States (i.e., number, type, mass 

of waste involved), the limited data available were judged inadequate for developing national 

emission estimates that could be included in the national inventory. However, a preliminary 

estimate of the potential magnitude of TEQ emissions associated with landfill fires in the United 

States can be obtained by assuming a direct correlation of emissions to population size for the 

United States and Sweden or by assuming a direct correlation between emissions and the number 

of landfills in each country. 

Both the United States and Sweden are industrialized countries. Although the per capita 

waste generation rate in the United States is nearly 1.5 times that of Sweden, the composition of 

municipal waste and the fraction of municipal waste disposed of in landfills in the two countries 

are nearly identical (U.S. EPA, 1996d). The population of Sweden was 8,825,417 in 1995 (U.S. 

DOC, 1995b) and 8,873,052 in 2000 (U.S. DOC, 2002). Based on these population estimates 

and the estimated annual Nordic TEQ emission factor of 35 g, the per capita landfill 

fire–associated Nordic TEQ emission factor was 4 μg TEQ per person per year for both 1995 and 

2000. Because congener-specific results were not provided in Persson and Bergström (1991) or 

Bergström and Björner (1992), it was not possible to derive emission factors in units of TEQDFWHO98 or I-TEQDF. Applying this factor to the U.S. population of 263,814,000 in 1995 (U.S. 

DOC, 1995b) and 281,421,906 in 2000 (U.S. DOC, 2003) results in an estimated annual 

emission of 1,050 g TEQ for 1995 and 1,126 g TEQ for 2000. These estimates should be 

regarded as preliminary indications of possible emissions from this source category; further 

testing is needed to confirm the true magnitude of these emissions. 

6.4. FOREST AND BRUSH FIRES 

6.4.1. Emissions Data 

Because CDDs/CDFs have been detected both in the soot from residential wood burning 

(Bumb et al., 1980; Nestrick and Lamparski, 1982, 1983; Bacher et al., 1992) and in the flue 

gases from residential wood burning (Schatowitz et al., 1993; Vickelsoe et al., 1993; Launhardt 

and Thoma, 2000; Environment Canada, 2000), it is reasonable to assume that wood burned in 

forest and brush fires may also be a source of CDDs/CDFs (Section 4.2 contains details on these 

studies). 

Only one study (Tashiro et al., 1990) could be found that reported direct measurements of 

CDDs/CDFs in emissions from forest fires. This study reported detection of total CDDs/CDFs in 

6-12
 

air at levels ranging from about 15 to 400 pg/m3

. The samples were taken from fixed collectors 

located 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 a preliminary report; however, no firm conclusions were drawn about whether 

forest fires are a CDD/CDF source. The final report on this study, Clement and Tashiro (1991), 

showed total CDD/CDF levels in the smoke of about 20 pg/m3. The authors concluded that 

CDDs/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 

CDDs/CDFs. 

Although not designed to directly assess whether CDDs/CDFs are formed during brush 

fires, Buckland et al. (1994) measured CDD/CDF levels in soil samples from both burnt and 

unburnt areas in national parks in New Zealand 6 weeks after large-scale brush fires. Four 

surface soil cores (2 cm depth) were collected and composited from each of three burnt and three 

unburnt areas. Survey results indicated that brush fires did not have a major impact on 

CDD/CDF levels in soil. The I-TEQDF content in the soil sample composites was 3, 8.7, and 10 

ng/kg for the three unburnt areas and 2.2, 3.1, and 36.8 ng/kg for the three burnt areas. Total 

CDD/CDF content ranged from 1,050 to 7,700 ng/kg in the unburnt area soil samples and from 

1,310 to 27,800 ng/kg in the burnt area soil samples. OCDD accounted for 94 to 97% of total 

CDD/CDF content in all samples. 

Similarly, a survey of controlled straw-field burning in the United Kingdom (Walsh et al., 

1994) indicated that the straw burning did not increase the CDD/CDF burden in the soil; 

however, a change in congener distribution was observed. Soils from three fields were sampled 

immediately before and after burning, along with ash from the fire. The mean I-TEQDF 

concentrations in the preburn soil, postburn soil, and ash were 1.79, 1.72, and 1.81 ng/kg, 

respectively. Concentrations of 2,3,7,8-TCDF were lower in the postburn soils than in the 

preburn soils. Conversely, concentrations of OCDD were higher in the postburn soils, indicating 

possible formation of OCDD during the combustion process. 

Van Oostdam and Ward (1995) reported finding no detectable levels of 2,3,7,8­

substituted CDDs/CDFs in three soil samples and four ash samples following a forest fire in 

British Columbia. The detection limits (DLs), on a congener-specific basis (unweighted for 

TEQ), ranged from 1 to 2 ng/kg. Nondetect values were also reported for ashes at a slash and 

burn site: the soil contained about 0.05 ng I-TEQDF/kg, whereas background soil contained about 

0.02 ng I-TEQDF/kg. 

The concentrations presented by Clement and Tashiro (1991) cannot accurately be 

converted to an emission factor because the corresponding rates of combustion gas production 

6-13
 

and wood consumption are not known. As a result, the following four alternative approaches 

were considered for developing an emission factor. 

Soot-based approach. This approach assumes that the levels of CDDs/CDFs in chimney 

soot are representative of the CDDs/CDFs in emissions. The CDD/CDF emission factor is 

calculated as the product of the CDD/CDF concentration in soot and the total particulate 

emission factor. This calculation involves first assuming that the CDD/CDF levels measured in 

chimney soot by Bacher et al. (1992) (720 ng I-TEQDF/kg) are representative of the CDD/CDF 

concentrations of particles emitted during forest fires. Second, the total particulate generation 

factor must be estimated. Using primarily data for head fires, Ward et al. (1976) estimated the 

national average particulate emission factor for wildfires as 150 lb/ton biomass dry weight. 

Ward et al. (1993) estimated the national average particulate emission factor for prescribed 

burning to be 50 lb/ton biomass dry weight. Combining the total particulate generation rates 

with the I-TEQDF level in soot results in emission factor estimates of 54 ng of I-TEQDF and 18 ng 

of I-TEQDF/kg of biomass burned in wildfires and prescribed burns, respectively. These 

estimated factors are likely to be overestimates because the levels of CDDs/CDFs measured in 

chimney soot by Bacher et al. (1992) may represent the accumulation and enrichment of 

CDDs/CDFs measured in chimney soot over time, leading to much higher assumed levels than 

the actual levels on emitted particles. 

Carbon monoxide (CO) approach. CO is a general indicator of the efficiency of 

combustion, and the emission factors of many emission products can be correlated with the CO 

emission factor. Data from Schatowitz et al. (1993) for emissions during natural wood burning 

in open stoves suggest an emission factor of 10 μg I-TEQDF/kg CO. Combining this factor with 

the CO emission factor during forest fires (roughly 0.1 kg CO/kg of biomass [Ward et al., 1993]) 

yields an emission factor of 1,000 ng I-TEQDF/kg biomass. This factor is higher than the sootbased factor discussed above, which is itself considered to be an overestimate. In addition, 

although the formation kinetics of CDDs/CDFs during combustion are not well understood, 

CDD/CDF emissions have not been shown to correlate well with CO emissions from other 

combustion sources. 

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 4.2.1, this approach suggests an emission factor of about 0.5 ng I

TEQ/kg wood combusted. This value appears more reasonable than the factors suggested by the 

soot and CO approaches because it is based on direct measurement of CDDs/CDFs from 

combustion of wood rather than on indirect techniques. However, forest fire conditions differ 

significantly from combustion conditions in wood stoves. For example, forest fire combustion 

6-14
 

­
does not occur in an enclosed chamber, and the biomass consumed in forest fires is usually green 

and includes underbrush, leaves, and grass. 

Forest fire simulation approach. This approach quantifies CDD/CDF emissions 

through the combustion of forest biomass in a controlled-burn facility. Using this approach, 

Gullet and Touati (2003) estimated CDD/CDF emissions through the testing of three biomass 

samples collected from the Oregon coast near Seal Rock and from four biomass samples 

collected from the North Carolina Piedmont region, approximately 200 km from the Atlantic 

coast. The samples generally consisted of equal portions of live shoots (needles cut from tree 

branches) and needle litter gathered from the forest floor. The Oregon samples were composed 

of pine needles (Pinus contorta and Pinus monticola) and hemlock needles (Tyuga heterophylla); 

the North Carolina samples were composed entirely of lobolly pine (Pinus taeda). The 

combustion of these seven samples, piled approximately 10 cm high, took place on top of an 

open, flat combustion platform. CDD/CDF emissions were measured using a Graseby PS-1 

sampler and EPA’s ambient TO-9 method. 

As shown in Table 6-2, the overall average total TEQ emission factor for the seven 

samples was 19.9 ng TEQDF-WHO98/kg (18.6 ng I-TEQDF), assuming nondetects were zero. 

Separately, the average total TEQ emission factors for the three Oregon samples and the four 

North Carolina samples were 15 ng and 25 ng TEQDF-WHO98/kg, respectively. Even though the 

average TEQ emission factors for the Oregon and North Carolina runs were similar, CDF 

congeners were dominant in the Oregon samples, whereas CDD congeners were dominant in the 

North Carolina samples. Figure 6-2 shows the congener profile for the Oregon and North 

Carolina samples combined. 

To test an alternative CDD/CDF sampling method, CDD/CDF emissions from one of the 

Oregon samples were also measured using a “Nomad” (a prototype portable sampler designed for 

mobile, in-field sampling). The results from both sampling methods showed very similar 

CDD/CDF TEQ values, total values, and ratio values. An additional Oregon sample was also 

combusted to test influences of fuel configuration on emissions. In this experiment, the biomass 

was placed in a metal barrel with air holes cut into the bottom. The results showed the highest 

total TEQ emission value calculated in this study (47 ng TEQDF-WHO98/kg); however, this value 

is similar to the next highest total TEQ value (46 ng TEQDF-WHO98/kg). 

Because the waxy cuticle layer on pine needles has been demonstrated to absorb 

lipophilic compounds from the atmosphere, Gullet and Touati (2003) also extracted a raw, asreceived Oregon biomass sample to determine whether the observed emissions were due to 

simple vaporization of existing CDDs/CDFs or the formation of new CDDs/CDFs in the 

combustion process. The CDD/CDF concentration in the sample measured 1.3 ng TEQDFWHO98/kg, which is approximately 20 times lower than the Oregon CDD/CDF emission 

6-15
 

Table 6-2. CDD/CDF mean emission factors (ng/kg) for forest firesa 

Congener Nondetect set to zero 

Nondetect set to 1⁄2 detection 

limit

 2,3,7,8-TCDD 1.15 1.28

 1,2,3,7,8-PeCDD 3.83 3.83

 1,2,3,4,7,8-HxCDD 5.68 5.68

 1,2,3,6,7,8-HxCDD 10.70 10.70

 1,2,3,7,8,9-HxCDD 17.34 17.34

 1,2,3,4,6,7,8-HpCDD 166.27 166.27

 OCDD 663.67 663.67

 2,3,7,8-TCDF 6.98 6.98

 1,2,3,7,8-PeCDF 6.34 6.35

 2,3,4,7,8-PeCDF 10.09 10.11

 1,2,3,4,7,8-HxCDF 16.72 16.74

 1,2,3,6,7,8-HxCDF 7.14 7.16

 1,2,3,7,8,9-HxCDF 1.11 1.20

 2,3,4,6,7,8-HxCDF 9.81 9.85

 1,2,3,4,6,7,8-HpCDF 25.39 25.39

 1,2,3,4,7,8,9-HpCDF 3.06 3.12

 OCDF 10.27 10.32

 Total CDD/CDF 965.55 965.99

 Total TEQDF-WHO98 19.90 20.06

 Total I-TEQDF 

18.60 18.75 

a

Values were derived from a total of seven biomass samples from Oregon and North Carolina. 

Source: Gullet and Touati (2003). 

concentrations (average, 25 ng TEQDF-WHO98/kg; range, 14 to 46 ng TEQDF-WHO98/kg). The 

CDD/CDF isomer patterns of the extracted biomass samples and the emission samples were 

similar. Therefore, this preliminary evidence suggests CDD/CDF emissions are not due solely to 

vaporization of cuticle-bound CDDs/CDFs but are formed anew during forest fires. 

Additionally, the new CDDs/CDFs formed may be adsorbed to the waxy cuticle layer in such a 

manner that the isomer pattern reflects the ambient CDD/CDF concentrations. 

6-16
 
8F 

Ratio (congener emission factor / 2378-CDD/CDF emission factor) 

0.000 0.100 0.200 0.300 0.400 0.500 0.600 0.700 0.800 

2,3,7,8,-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

1,2,3,4,6,7,8,9-OCDD 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-HxCDF 

1,2,3,4,7,8-PeCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

1,2,3,4,6,7,8,9-OCDF 

Figure 6-2. Congener profile for forest fire simulation approach emissions. 

Source: Gullet and Touati (2003). 

Many factors may affect forest fire CDD/CDF emissions, such as the type of fire (crown 

vs. understory and duff), types of species combusted, and location of the fire (near coastal vs. 

inland). Additionally, combustion conditions such as wind speed and fuel moisture content may 

also result in variations in emissions. These variables, therefore, yield uncertainties in the 

calculation of a representative emission factor through forest fire simulations. However, the 

emission factor of 19.9 ng TEQDF-WHO98/kg (18.6 ng I-TEQDF) calculated through this approach 

appears to be more reasonable than the factors suggested by the soot, CO, and wood stove 

approaches because the forest fire simulation approach directly measures CDD/CDF emissions 

from forest biomass combusted in an open pile. Additionally, the forest biomass samples 

consisted of both live shoots and needle litter of representative species from two distinct 

locations. 

6-17
 

6.4.2. Activity Level Information 

6.4.2.1. Approach for Reference Year 2000 (Office of Air Quality Planning and Standards 

[OAQPS]) 

As part of the 2000 National Emissions Inventory, OAQPS developed activity levels of 

wildfires and prescribed burning on a county-level basis for reference year 2000. The number of 

acres burned by wildfires and prescribed burning was obtained from the U.S. Forest Service 

(USFS) and four U.S. Department of Interior agencies: Bureau of Land Management, National 

Park Service, U.S. Fish and Wildlife Service, and Bureau of Indian Affairs (BIA). USFS 

provided data for federal, state, and private lands. All data were provided on a state level except 

for the BIA wildfire data and the USFS prescribed burning data, which were provided on a 

regional level. 

Prior to allocating the forest fire activity to the county level, the BIA and USFS regional 

data were first allocated to the state level. The BIA data were allocated using the number of 

acres of tribal land in each state. The USFS data were allocated using factors developed from 

landcover data in the Biogenic Emissions Landcover Database (BELD2) within EPA’s Biogenic 

Emissions Inventory System; however, the BELD2 data for California were replaced with data 

from the 1996 National Toxics Inventory because USFS’s Region 5 contains both Hawaii and 

California. 

For each of the forest fire categories, the activity from all the agencies were then totaled 

by state and allocated to the county level using state-to-county land cover factors developed from 

BELD2. These BELD2 factors were based on the acreage of rural forest, brush, and grass in each 

county. This procedure was used for all states except Alaska and Hawaii, for which BELD2 does 

not contain land cover data. For Alaska and Hawaii, state-to-county factors were derived from 

data contained in the allocation factor file used for the 1996 National Emissions Inventory. 

Using this approach, OAQPS estimated that approximately 8,357,958 acres were burned 

by wildfires in 2000 and approximately 1,261,607 acres were burned by prescribed fires in 2000. 

To obtain the amount of biomass consumed by wildfires and prescribed burning, the acres of 

forest burned were combined with region-specific fuel loading factors, as shown in Table 6-3. 

Nationally, approximately 228 million tons of biomass were consumed by wildfires and 15.8 

million metric tons of biomass were consumed by prescribed burning in 2000. 

6.4.2.2. Approach for Reference Years 1987 and 1995 

According to the Council on Environmental Quality’s 25th annual report (CEQ, 1997), 5 

million acres of forest were lost to wildfires in 1987 and 7 million acres were lost in 1995. 

Estimates of the acreage consumed annually during prescribed burns are not readily available for 

6-18
 

Table 6-3. Forest fire fuel loading factors (tons/acre) 

Region Wildfires Prescribed burning

 Alaska 6 12.6

 California 18 14.2

 Intermountain 8 6.3

 North Central 11 8.7

 Northern 60 47.3

 Pacific Northwest 60 47.3

 Rocky Mountain 30 23.7

 Southern 9 7.1

 Southwestern 10 7.9 

Source: U.S. EPA (2002b). 

1995 and 1997. An estimated 5.1 million acres of biomass were burned in 1989 during 

prescribed burns (Ward et al., 1993). This value was assumed to be appropriate for use for 

reference years 1987 and 1995. 

To obtain the amount of biomass consumed by wildfires and prescribed burning, the acres 

of forest burned were combined with biomass consumption rates of 9.43 metric tons/acre in areas 

consumed by wildfires (Ward et al., 1976) and 7.44 metric tons/acre in areas consumed in 

prescribed burns. For 1987 and 1995, approximately 38 million tons were consumed by 

prescribed burns. Approximately 47 million metric tons of biomass were consumed by wild fires 

in 1987 and approximately 66 million metric tons of biomass were consumed in 1995. 

6.4.3. Emission Estimates 

Combining the emission factor developed using the forest fire simulation approach (19.9 

ng TEQDF-WHO98/kg biomass [18.6 ng I-TEQDF/kg biomass]) with the amount of biomass 

consumed annually in wildfires and prescribed fires (total of 85 million metric tons in 1987, 104 

million metric tons in 1995, and 244 million metric tons in 2000) yields annual emission 

estimates of 1,700 g TEQDF-WHO98 (1,581 g I-TEQDF) for 1987; 2,080 g TEQDF-WHO98 (1,934 g 

I-TEQDF) for 1995; and 4,880 g TEQDF-WHO98 (4,538 g I-TEQDF) for 2000. For wildfires 

specifically, annual reference year emission estimates are 940 g TEQDF-WHO98 (874.2 g I

TEQDF) for 1987; 1,320 g TEQDF-WHO98 (1,228 g I-TEQDF) for 1995; and 4,560 g TEQDFWHO98 (4,241 g I-TEQDF) for 2000. For prescribed fires specifically, annual emission estimates 

6-19
 

­
are 760 g TEQDF-WHO98 (706.8 g I-TEQDF) for reference years 1987 and 1995 and 320 g TEQDFWHO98 (297 g I-TEQDF) for reference year 2000. 

These estimates should be regarded as preliminary indications of possible emissions from 

this source; further testing is needed to confirm the true magnitude of emissions. The activity 

level for both forest fires and biomass combustion is given a low confidence rating because these 

values were estimated and may not be representative. The emission factor is highly variable and 

dependent on type of biomass burned, therefore, it is judged to be clearly nonrepresentative. 

6.5. BACKYARD BARREL BURNING 

6.5.1. Emissions Data 

In many rural and nonurban areas of the United States, residences may dispose of 

household refuse through open backyard burning. This practice usually consists of burning 

refuse in a 208-L capacity steel drum. Holes are punched near the bottom of the drum to allow 

combustion air to enter. The fire is ignited with a petroleum fuel, e.g., kerosene. The low 

combustion temperatures and oxygen-starved conditions associated with household refuse 

burning in these “burn barrels” results in poor and uncontrolled combustion conditions. Under 

such conditions, products of incomplete combustion are formed and visible smoke is emitted into 

the air. 

The practice of open burning in burn barrels causes CDDs and CDFs to be formed and 

released as toxic air contaminants. In 1997, EPA’s Control Technology Center, in cooperation 

with the New York State’s Department of Health and Department of Environmental 

Conservation, conducted an initial study that examined, characterized, and quantified emissions 

from simulated open burnings of household waste materials in barrels (Lemieux, 1997). The 

representative waste was prepared on the basis of the typical percentages of various waste 

materials disposed of by New York State residents; hazardous wastes (e.g., chemicals, paints, 

oils) were not included in the test waste. A variety of compounds, including CDDs/CDFs, were 

measured in the emissions from two simulated open burnings of this “baseline” waste. 

Combustion studies were subsequently performed by EPA to provide additional baseline 

waste tests and an initial indication of the impact of limited variation in waste composition and 

combustion conditions on CDD/CDF emissions from a simulated domestic backyard barrel burn 

of 6.8 kg of unshredded household waste (Gullet et al., 1999, 2000a, b; Lemieux et al., 2000; email dated September 7, 2000, from P. Lemieux, U.S. EPA, to D. Cleverly, U.S. EPA). The 

results of seven baseline open burning waste tests were reported in these EPA studies. These 

tests exhibited variation in the emissions of CDDs/CDFs, with a one to two order-of-magnitude 

spread between the lowest and highest values for individual congeners, congener groups, total 

CDDs/CDFs, and TEQ values. The average TEQ emission factor for the seven baseline tests was 

6-20
 

72.8 ng I-TEQDF/kg waste burned (setting nondetect values equal to zero) and 73.7 ng ITEQDF/kg waste burned (setting nondetect values equal to one-half the DL). The corresponding 

TEQDF-WHO98 values were 76.8 and 77.7 ng/kg. Table 6-4 presents the average congener and 

congener group results for these tests. 

In addition to the baseline tests, the combustion experiments included testing at three 

different PVC levels: 0, 1, and 7.5% by weight PVC. The average emissions were 14, 201, and 

4,916 ng I-TEQDF/kg waste burned, respectively. Two tests using waste impregnated with 

inorganic chloride (CaCl2) at a concentration of 7.5% by weight (and no PVC) averaged 734 ng ITEQDF/kg. Qualitative comparisons suggest that the tests conducted with higher chlorine, via 

PVC or CaCl2, resulted in substantial increases in TEQ emissions. 

Other variations in baseline waste composition included conducting one test with 

compressed waste, one test with a double load of waste, and one test in which some of the waste 

paper was wetted to simulate high-moisture burns. These tests resulted in a higher mean TEQ 

emission factor (534 ng I-TEQDF/kg) than that of the baseline runs. 

Several waste combustion variables were evaluated, such as average temperatures at 

prescribed barrel heights, length of time temperatures (favorable temperature ranges) for 

CDD/CDF formation, and measurement of CO, CO2, O2, particulate matter, and HCl. Statistical 

analyses of the results indicated that CO emissions and temperature measured in the uppermost 

portion of the barrel were the best predictors of TEQ variation. However, the wide variability in 

test results (from less than 10 to more than 6,000 ng I-TEQDF/kg) also indicates that a high degree 

of CDD/CDF emission variation can be expected due to factors that are not wholly related to 

waste composition or burning practice, such as waste orientation. A mean emission factor of the 

baseline tests (without PVC added) was developed from the data. This mean emission factor was 

78.6 mg TEQDF-WHO/kg, and it was used to estimate releases from barrel burning. The 

emission factor is given a low confidence rating because it is possibly nonrepresentative of 

barrel-burning emissions. 

6.5.2. Activity Level Information 

The amount of refuse that is combusted annually in the United States in residential 

backyard burn barrels is largely unknown. Although no national statistics are available, a limited 

number of telephone surveys have attempted to measure the prevalence of backyard barrel 

burning in a few geographical areas. This limited number of surveys, combined with census data 

on the rural and nonurban population of the United States, were used to estimate annual activity 

level in terms of the quantity of refuse combusted in burn barrels per reference year. The 

following is a summary of this estimation procedure. 

6-21
 
Table 6-4. CDD/CDF average air emission factors (ng/kg waste burned) 

from barrel burning of household wastea 

Congener/congener group Nondetect set to 1⁄2 detection limit Nondetect set to zero 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

3.4 

8.2 

6.6 

9.9 

19.1 

39.8 

49.7 

2.7 

8.1 

6.4 

9.7 

19 

39.8 

49.7 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

45.6 

37.2 

65.2 

113.8 

38.5 

61.9 

3 

128.6 

14.6 

37.5 

45.6 

37.2 

65.2 

113.8 

38.5 

61.9 

2.5 

124.4 

15 

36.4 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

136.7 

545.9 

73.7 

77.7 

135.4 

540.5 

72.8 

76.8 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

413 

281 

221 

105 

43 

1,880 

1,021 

492 

169 

32 

413 

281 

221 

105 

43 

1,880 

1,021 

492 

169 

30 

Total CDD/CDF 4,657 4,655 a

Listed values are the arithmetic averages of seven tests for the congeners and the averages of five tests for the 

congener groups. 

Sources: E-mail dated September 7, 2000, from P. Lemieux, U.S. EPA, to D. Cleverly, U.S. EPA; Gullett et al. 

(1999, 2000, 2001). 

6-22
 

6.5.2.1. Summary of Barrel Burn Surveys 

A total of seven surveys of the prevalence of backyard combustion of domestic refuse in 

burn barrels were identified in the literature. For the most part, these surveys were an attempt to 

estimate the barrel-burning activity in a specific state, county, or region in support of regulatory 

determinations on barrel burning. In general, the results of the surveys showed a prevalence of 

barrel burning within the rural population to range from 12 to 40%, with a mean of 28%. The 

following is a review of the surveys. 

The Two Rivers Region Council of Public Officials (TRRCPO) and Patrick Engineering 

conducted a telephone survey in the early 1990s of residents of five central Illinois counties. 

They found that about 40% of the residents in a typical rural county burned household waste. 

The survey also indicated that, on average, those households that burned waste disposed of 

approximately 63% of their household waste by burning it in barrels (TRRCPO, 1994). 

Similar results were obtained in a survey conducted by Zenith Research Group, Inc., 

(2000) for the Western Lake Superior Sanitary District of Minnesota. This survey of 760 

residents of selected portions of northwestern Wisconsin and northeastern Minnesota addressed, 

in part, the use of burn barrels or other devices to burn household garbage or other materials. 

Among all survey respondents, 27.5% said they used burn barrels or other devices to burn 

household garbage or other materials. 

Environics Research Group conducted a household garbage disposal and burning survey 

of 1,516 residents of Ontario, Canada. All respondents resided in detached single-family homes. 

Approximately 24% of all respondents reported burning their household refuse in burn barrels 

(Environics Research Group, 2001). 

E.H. Pechan and Associates conducted a residential municipal solid waste survey for the 

Mid-Atlantic/Northeast Visibility Union (MANE-VU) states and tribes (Pechan and Associates, 

2002). The MANE-VU entities include Connecticut, Delaware, the District of Columbia, Maine, 

Maryland, Massachusetts, New Hampshire, New Jersey, New York, Pennsylvania, the Penobscot 

Indian Nation, Rhode Island, the St. Regis Mohawk Tribe, and Vermont. Household waste 

burning surveys were conducted by telephone for 72 residents of rural, suburban, and urban 

jurisdictions, as classified by the 1990 census. The residents were asked to estimate the number 

of households in their jurisdiction that burned household waste or trash. In general, the survey 

estimated that 11.9% of the rural population burned refuse in backyard burn barrels. 

The State of California Air Resources Board (CARB) undertook a study of the prevalence 

of backyard refuse burning in rural areas of 21 air management districts in California (CARB, 

2002). From this study, CARB estimated that approximately 18% of the rural population in 

California combusted their household refuse in backyard burn barrels. 

6-23
 
In 1993, the St. Lawrence County Planning Office in Canton, NY, conducted a survey of 

open burning of domestic refuse (St. Lawrence County, 1993). From the survey, it was 

concluded that 48.2% of 9,926 households in rural areas of the county burned household refuse 

in burn barrels. 

In 1997, the State of Maine Department of Conservation Forestry Bureau surveyed rural 

town fire wardens and state fire rangers regarding the prevalence of backyard burning of 

household waste. It was revealed through the survey that each day approximately 19,147 kg of 

domestic refuse was being combusted state wide in approximately 8,510 burn barrels. In relation 

to the total population, it was noted that one burn barrel existed for every 144 individuals. 

6.5.2.2. Estimates of Activity Level 

The following chart summarizes the steps taken to estimate the quantity of household 

refuse combusted in backyard burn barrels in1987, 1995, and 2000. 

Step Assumption 2000 1995 1987 

1 U.S. population 281,400,000 260,600,000 242,300,000 

2 Population in rural and 

nonurban areas 59,000,000 52,700,000 50,700,000 

3 Percent nonurban 

population burning 

household refuse 28 40 40 

4 Adjusted population 

burning household refuse in 

barrels 16,726,500 21,080,000 20,280,000 

5 Per capita household refuse 

generation rate (kg/yr) 616 616 616 

6 Percent of household refuse 

generated burned at homes 63 63 63 

7 Adjusted per capita 

household refuse burned 

(kg/yr) 388 388 388 

8 Total refuse generated by 

rural and nonurban 

population burning 

household refuse (kg/yr) 6,491,220,120 8,180,726,400 7,870,262,400 

For steps 1 and 2, the U.S. Census Bureau Statistical Abstract of the United States was 

used to determine population size. For step 3, the assumption of the percent of rural population 

6-24
 

combusting refuse in burn barrels in 2000 was derived from the overall mean prevalence rate 

from six surveys (CARB, 2002; Zenith Research Group, 2001; Environics Research Group, 

2001: Pechan and Associates, 2004; St. Lawrence County, 1993; and TRRCPO, 1994). This 

produced a mean prevalence of 28%, which should reflect the impact of state bans or restrictions 

on the practice of open burning of refuse. The 1995 and 1987 population was estimated using 

the survey by TRRCPO (1994). This produced a prevalence of 40%, which should reflect the 

fact that the practice of open burning of refuse was not banned or restricted by the majority of the 

states. 

The above mean prevalence rates were used in step 4 to calculate the number of people 

residing in rural areas assumed to have burned household refuse in burn barrels in each reference 

year (assumption in step 2 multiplied by the percentage in step 3). The annual per capita 

household refuse generation rate in step 5 is from Municipal Solid Waste Fact Book (2000). The 

figure of 616 kg/person/yr is the result of subtracting out weight of yard waste from the per capita 

generation rate. The assumption in step 6 is that 63% of the household refuse generated in rural 

areas is burned in backyard burn barrels. This percentage was derived from the survey conducted 

in rural counties of Illinois by TRRCPO (1994). 

These activity levels are adopted and assigned a confidence rating of low because they are 

derived from limited surveys that are possibly nonrepresentative of the national activity level. 

6.5.2.3. Alternative Approach to Estimating Activity Level 

The actual amount of refuse combusted in burn barrels in the United States is unknown 

and must be estimated. OAQPS developed activity levels of residential MSW combusted in 

backyard barrels for reference year 2000. The activity levels were determined by first estimating 

the amount of waste generated for each county in the United States. The amount of waste 

generated was estimated by using a national average per capita waste generation factor, which is 

1.5 kg/person/day. This value was calculated using population data from the 2000 census and 

2000 waste generation data (U.S. EPA, 2002c). To better reflect the actual amount of household 

residential waste subject to being burned, noncombustibles (glass and metals) and yard waste 

were excluded. This factor was then applied to the portion of the county’s total population that is 

considered rural, since open burning is generally not practiced in urban areas. 

Using data from TRRCPO (1994), it was estimated that for rural populations, 25 to 32% 

of generated MSW is burned. A median value of 28% was assumed for the nation, and this 

correction factor was applied to the total amount of waste generated. Controls (or burning bans) 

were accounted for by assuming that no burning takes place in counties where the urban 

population is at least 80% of the total population (i.e., urban plus rural). Zero emissions from 

open burning were attributed to these counties. This technique produced an estimated annual 

6-25
 

activity level of 7.79 billion kg of residential household waste combusted in burn barrels in 2000. 

This estimate is approximately 16.5% greater than the estimate used in this report. 

6.5.3. Emission Estimates 

CDD/CDF emissions from burn barrels for reference years 1987, 1995, and 2000 were 

calculated by multiplying the estimated annual total weight of household refuse combusted in 

burn barrels (see Section 6.5.2.2) by the dioxin emission factor. The emission factor was 76.8 ng 

TEQDF-WHO98/kg (72.8 ng I-TEQDF/kg) waste burned. 

Annual nationwide TEQ emissions for 1987, 1995, and 2000 were calculated using 

the following equation: 

ETEQ = EFTEQ H AL 

where: 

ETEQ = Annual TEQDF emissions (g/yr) 

EFTEQ = TEQDF emission factor (76.8 ng TEQDF-WHO98/kg waste burned) 

AL = Annual activity level (7,870,262,400 kg in 1987; 8,180,726,400 kg in 1995; and 

 6,491,220,120 kg in 2000). 

Using this equation, estimated nationwide TEQDF emissions were 604 g TEQDF-WHO98 

(573 g I-TEQDF) in 1987, 628 g TEQDF-WHO98 (595 g I-TEQDF) in 1995, and 498.53 g TEQDFWHO98 (472.56 g I-TEQDF) in 2000. A low confidence rating is given to both the emission factor 

and the estimate of activity level, therefore, the confidence rating is low for the estimate of TEQ 

emissions from backyard barrel burning of refuse. 

6.5.4. Composition of Ash from Barrel Burning 

Ash samples were collected from open barrel burning (Lemieux, 1997) and analyzed for 

CDDs/CDFs and PCBs. Ash samples from the experiments were combined, resulting in two 

composite samples, one for recyclers and one for nonrecyclers. The results for PCBs depict only 

the data for specific PCB congeners. The remaining PCB data reported in Lemieux (1997) could 

not be related to a particular congener. The results are presented in Tables 6-5 and 6-6. 

6.6. RESIDENTIAL YARD WASTE BURNING 

6.6.1. Emissions Data 

It is reasonable to assume that residential yard waste burning may be a source of 

CDDs/CDFs, as they have been detected in forest and brush fires. No direct measurements of 

CDD/CDF emissions from residential yard waste burning have been performed; however, Gullet 

6-26
 

Table 6-5. CDD/CDF analysis for composite ash samples from barrel 

burning (ng/kg ash) 

Congener/congener 

group 

Average 

concentration in 

composite ash sample I-TEQDF TEQDF-WHO98 

Avid 

recycler 

Nonrecycler Avid 

recycler 

Nonrecycler Avid 

recycler 

Nonrecycler 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

31 

230 

270 

420 

300 

4,000 

9,600 

9 

53 

44 

74 

56 

630 

690 

31 

115 

27 

42 

30 

40 

9.6 

9 

26.5 

4.4 

7.4 

5.6 

6.3 

0.69 

31 

230 

27 

42 

30 

40 

0.96 

9 

53 

4.4 

7.4 

5.6 

6.3 

0.069 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

830 

1,000 

2,500 

2,300 

2,100 

2,900 

810 

12,000 

1,400 

8,200 

220 

270 

690 

480 

490 

670 

150 

2,100 

170 

560 

83 

50 

1,250 

230 

210 

290 

81 

120 

14 

8.2 

22 

13.5 

345 

48 

49 

67 

15 

21 

1.7 

0.56 

83 

50 

1,250 

230 

210 

290 

81 

120 

14 

0.82 

22 

13.5 

345 

48 

67 

15 

21 

1.7 

0.056 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

2,500 

4,100 

5,600 

7,600 

9,600 

490 

740 

1,300 

1,300 

690 

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

25,000 

21,000 

19,000 

17,000 

8,200 

8,200 

6,600 

4,600 

2,900 

560 

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

Total CDD 

Total CDF 

29,400 

90,200 

4,520 

22,860 

-

-

-

-

-

-

-

-

Total CDD/CDF 119,600 27,380 - - - -

Source: Lemieux (1997). 

6-27
 

Table 6-6. PCB analysis for composite ash samples from barrel burning 

(ng/kg ash) 

Compound Avid recycler Nonrecycler 

2-Chlorobiphenyl <2,500 4,900 

2,3'-Dichlorobiphenyl 3,700 4,700 

2,2',6-Trichlorobiphenyl <500 5,600 

2,2',5-Trichlorobiphenyl 32,000 6,300 

2,3',5-Trichlorobiphenyl 800 800 

2,3',4-Trichlorobiphenyl <500 700 

2,4',5-Trichlorobiphenyl 1,500 900 

2,4,4'-Trichlorobiphenyl <500 500 

2,2',4,6'-Tetrachlorobiphenyl <500 1,500 

2,2',3,6'-Tetrachlorobiphenyl 5,300 1,300 

2,2',5,5'-Tetrachlorobiphenyl 3,100 1,800 

2,2',3,5'-Tetrachlorobiphenyl 2,600 1,200 

2,2',4,4',5-Pentachlorobiphenyl 3,400 1,300 

2,2',3,3',5-Pentachlorobiphenyl 400 <500 

2,2',3,4,5,5'-Hexachlorobiphenyl 1,200 <500 

Source: Lemieux (1997). 

and Touati (2003) measured an average CDD/CDF emission factor of 20 ng TEQDF-WHO98/kg 

during forest fire simulations where biomass samples from Oregon and North Carolina were 

burned on an open platform (see Section 6.4). Therefore, the emission factor of 20 ng TEQDFWHO98/kg developed by Gullet and Touati (2003) was used for residential yard waste burning. 

6.6.2. Activity Level Information 

Pechan and Associates (2002) estimated that approximately 233 lb of yard waste per 

household per year (based on a four-person household) were burned in 2000. This is similar to 

the estimate of 106 to 319 lb yard waste per household per year (based on a four-person 

household) that Pechan calculated using the results of a household yard waste burning survey. 

Pechan’s telephone survey was completed by 181 rural, suburban, and urban jurisdictions of the 

MANE-VU entities. The results indicated that approximately 28% of the population in rural 

areas burned household yard waste and that households typically conducted two to three burns 

6-28
 

per year. Additionally, as indicated by information provided by three respondents, 1 to 3 cubic 

yards of yard waste was typically burned at a time. 

As part of the 2000 National Emissions Inventory, OAQPS determined on a county-level 

basis the amount of yard waste burned in 2000. The activity level estimates were based on the 

assumption that yard waste was generated at a rate of 0.54 lb/person/day in 2000, which in turn 

was derived using population data for 2000 and the assumption that 27.7 million tons of yard 

waste were generated in 2000 (U.S. EPA, 2002c). Of the total amount of yard waste generated, 

the composition was assumed to be 25% leaves, 25% brush, and 50% grass by weight (U.S. EPA, 

2001b). Because open burning of grass clippings is not typically practiced by homeowners, only 

50% of the yard waste generated was assumed to be burnable. Additionally, OAQPS assumed 

that burning primarily occurs in rural areas (i.e., the per capita yard waste generation factor was 

applied to only the rural population in each county) and that only 28% of the total yard waste 

generated was actually burned (see Section 6.5). 

The amount of yard waste assumed to be generated in each county was then adjusted for 

variation in vegetation using BELD2. For counties with 10 to 50% forested land, the amount of 

yard waste generated was reduced to 50% and for counties with less than 10% forested land, to 

zero (i.e., no yard waste was generated). Adjustments for variation in vegetation were not made 

to counties where the percentage of forested acres was greater than or equal to 50%. Before 

calculating the percentage of forested acres per county, the acreage of agricultural lands was 

subtracted from the acreage of forested lands to better account for the native vegetation that 

would likely be occurring in the residential yards of farming states. Controls (or burning bans) 

were accounted for by assuming that no burning took place in counties where the urban 

population exceeded 80% of the total population (i.e., urban plus rural). Using this method, 

OAQPS estimated that 255,000 metric tons of leaf and 255,000 metric tons of brush (total of 

510,000 metric tons of yard waste) were burned in 2000. 

6.6.3. Emission Estimates 

Using the emission factor of 20 ng TEQDF-WHO98/kg (18.6 ng I-TEQDF/kg) and the 

activity level of 510,000 metric tons yard waste burned in 2000, CDD/CDF emissions from open 

burning of yard waste were 10.2 g TEQDF-WHO98 (9.5 g I-TEQDF) in 2000. Assuming 772 and 

754 million kg of yard waste were burned in 1987 and 1995, respectively, then 15.4 g TEQDFWHO98 (14.4 g I-TEQDF) and 15.19 g TEQDF-WHO98 (14 g I-TEQDF) were emitted in 1987 and 

1995, respectively. These numbers should be regarded as preliminary estimates of possible 

emissions from this source; further testing is needed to confirm the true magnitude of emissions 

because both the emission factor and activity levels are judged to be clearly nonrepresentative of 

the source category. 

6-29
 
6.7. LAND-CLEARING DEBRIS BURNING 

6.7.1. Emissions Data 

During the clearing of land for the construction of new buildings (residential and 

nonresidential) and highways, trees, shrubs, and brush are often torn out, collected in piles, and 

burned. As with residential yard waste burning, it is assumed that the burning of land-clearing 

debris may generate CDDs/CDFs because emissions have been detected from forest and brush 

fires. No direct measurements of CDD/CDF emissions from the burning of land-clearing debris 

have been performed, so the average emission factor of 20 ng TEQDF-WHO98/kg, which was used 

for both forest fires and residential yard waste burning, is also used for burning of land clearing 

debris (see Sections 6.4 and 6.6). 

6.7.2. Activity Level Information 

Activity levels associated with land-clearing debris were calculated by OAQPS on a 

county-level basis using the number of acres disturbed through residential, nonresidential, and 

roadway construction. 

6.7.2.1. Residential Construction 

In 2000, approximately 330,551 acres were disturbed by residential construction. This 

number is based on county-level housing permit data and regional housing start data obtained 

from the U.S. Census Bureau for single-family units, two-family units, and apartment buildings. 

The county permit data were first adjusted to equal regional housing-start data, and then the 

number of buildings in each housing category was estimated. The total number of acres 

disturbed by residential construction was then determined by applying the following conversion 

factors to the housing-start data for each category:

 Unit type Acres per building 

Single-family unit 1/4 

Two-family unit 1/3 

Apartments 1/2 

6.7.2.2. Nonresidential Construction 

In 2000, approximately 336,224 acres were disturbed by nonresidential construction. 

This number is based on the national value of construction put in place, as reported by the U.S. 

Census Bureau. The national value was allocated to counties using construction employment 

data from the Bureau of Labor Statistics and Dun & Bradstreet. A conversion factor of 1.6 acres 

disturbed per $100,000 spent was applied to the county-level estimates of the value of 

6-30
 
construction put in place to obtain the acres disturbed by nonresidential construction per county. 

The conversion factor was developed using the Price and Cost Indices for Construction by 

adjusting the 1992 value of 2 acres per $100,000 for 2000. 

6.7.2.3. Roadway Construction 

In 2000, approximately 190,367 acres were disturbed by roadway construction. This 

number is based on 1999 Federal Highway Administration state expenditure data for capital 

outlay within the following six road classifications: interstate (urban and rural), other arterial 

(urban and rural), and collectors (urban and rural). The expenditure data were converted to miles 

of road constructed based on data from the North Carolina Department of Transportation 

(NCDOT). According to NCDOT, approximately $4 million per mile is spent for freeway and 

interstate projects, and approximately $1.9 million per mile is spent for arterial and collector 

projects. The number of miles was then converted to acres disturbed using the following 

conversion factors for each road classification category:

 Road type Acres per mile 

Interstate, urban 15.2 

Interstate, rural 15.2 

Other arterial, urban 15.2 

Other arterial, rural 12.7 

Collectors, urban 9.8 

Collectors, rural 7.9 

For 1995, state expenditure for capital outlay was assumed to be 74% of total funding. 

This percentage was derived using 2000 data (U.S. DOT, 2002). For 1987, 74% of the total 

capital outlay of the average of 1985 and 1989 was used (capital outlays for 1985 and 1989 are 

reported in U.S. DOT, 2002). Therefore, approximately 83,110 and 123,140 acres were 

disturbed as the result of roadway construction in 1987 and 1995, respectively. 

6.7.2.4. Fuel Loading Factors 

To obtain the amount of biomass consumed by the burning of land-clearing debris, the 

total acreage of land disturbed in each county by residential, nonresidential, and roadway 

construction was distributed according to vegetation type (hardwood, softwood, and grass) and 

then combined with vegetation-specific fuel loading factors. The percentage of vegetation type 

within each county was determined using BELD2. The average loading factors used for each 

fuel type were 99 tons/acre for hardwood, 57 tons/acre for softwood, and 4.5 tons/acre for grass. 

6-31
 

Using this method, OAQPS estimated that 28.4 million metric tons of biomass were 

burned through land clearing activities in 2000. EPA developed a national average biomass 

loading factor of 33 metric tons burned per acre in 2000. Using this loading factor combined 

with total acreage disturbed, EPA estimated that approximately 27.7 and 26.4 million metric tons 

of biomass were burned by land clearing in 1987 and 1995, respectively. 

6.7.3. Emission Estimates 

Using the emission factor of 20 ng TEQDF-WHO98/kg (18.6 ng I-TEQDF/kg) and the 

activity level estimates in Section 6.7.2.4, CDD/CDF emissions from land clearing burning were 

568 g TEQDF-WHO98 (528 g I-TEQDF) in 2000, 528 g TEQDF-WHO98 (491 g I-TEQDF) in 1995, 

and 553 g TEQDF-WHO98 (515 g I-TEQDF) in 1987. These should be regarded as preliminary 

estimates of possible emissions from this source because the emission factor is clearly 

nonrepresentative; further testing is needed to confirm the true magnitude of emissions. 

6.8. UNCONTROLLED COMBUSTION OF POLYCHLORINATED BIPHENYLS 

The accidental combustion of PCB-containing electrical equipment or intentional 

combustion of PCBs in incinerators and boilers not approved for PCB burning (40 CFR 761) 

may produce CDDs/CDFs. At elevated temperatures, such as in transformer fires, PCBs can 

undergo reactions to form CDFs and other by-products. More than 30 accidental fires and 

explosions involving PCB transformers and capacitors in the United States and Scandinavia that 

involved the combustion of PCBs and the generation of CDDs/CDFs have been documented 

(Hutzinger and Fiedler, 1991a; O’Keefe and Smith, 1989; Williams et al., 1985). For example, 

analyses of soot samples from a Binghamton, NY, office building fire detected 20 μg/g total 

CDDs (0.6 to 2.8 μg/g 2,3,7,8-TCDD) and 765 to 2,160 μg/g total CDFs (12 to 270 μg/g 2,3,7,8­

TCDF). At that site, the fire involved a mixture containing PCBs (65%) and chlorobenzene 

(35%). Laboratory analyses of soot samples from a PCB transformer fire that 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. 

Using a bench-scale thermal destruction system, Erickson et al. (1984) determined the 

optimum conditions for CDF formation to be a temperature of 675°C, an excess oxygen 

concentration of 8%, and a residence time of 0.8 sec (or longer). Combusting mineral oil and 

silicone oil containing 5, 50, and 500 ppm of Aroclor 1254 at these conditions yielded PCB to 

CDF conversion efficiencies as high as 4%. Up to 3% conversion efficiency was observed when 

an Askarel (70% Aroclor 1260) was combusted under the same conditions. 

6-32
 
The use of PCBs in new transformers in the United States is banned, and their use in 

existing transformers and capacitors is being phased out under regulations promulgated under the 

Toxic Substances Control Act. 

Because of the accidental nature of these incidents, the variation in duration and intensity 

of elevated temperatures, the variation in CDD/CDF content of residues, and uncertainty 

regarding the amount of PCBs still in service in electrical equipment, EPA judged the available 

data inadequate for developing any quantifiable emission estimates. However, Thomas and 

Spiro (1995) conservatively estimated that about 15 g of TEQ may be generated annually from 

fires in commercial and residential buildings each year. This estimate is based on the following 

assumptions: (a) the I-TEQDF emission rate is 20 μg/kg of PCB burned, (b) 74,000 metric tons of 

PCB are still in use in various electrical equipment, and (c) 1% of the in-use PCBs are burned 

during the course of structural fires annually. 

6.9. VOLCANOES 

To date, no studies demonstrating the formation of CDDs/CDFs by volcanoes have been 

published. Given the available information from the studies discussed below, volcanoes do not 

appear to be sources of CDD/CDF release to the environment. 

Gribble (1994) summarized some of the existing information on the formation of 

chlorinated compounds by natural sources, including volcanoes. Gribble reported that several 

studies had demonstrated the presence of chlorofluorocarbons and simple halogenated aliphatic 

compounds (one and two carbon chain length) in volcanic gases. In addition, several chlorinated 

monoaromatic compounds as well as three PeCB congeners were detected in the ash from the 

1980 eruption of Mount St. Helens. Gribble hypothesized that the formation of these PCB 

compounds was the result of rapid, incomplete, high-temperature combustion of chloridecontaining plant material in the eruption zone. However, no information was presented to 

indicate the formation of CDDs/CDFs by volcanoes. 

Lamparski et al. (1990) analyzed groundfall ash samples collected at various distances 

and locations from Mount St. Helens following the eruption in 1980. The findings of this study 

indicate that volcanic particulate emissions were free of detectable PCBs and nearly free of 

detectable CDDs (0.8 ng/kg HpCDD detected) upon exiting the volcano and remained so 

throughout their period of deposition in the blast zone. However, upon transport through the 

atmosphere, measurable and increasing levels of CDDs and PCBs were detected in deposited ash 

as it passed from rural to urban environments. The authors hypothesized that CDDs and PCBs in 

the atmosphere became associated with the volcanic ash particulates through gas-phase sorption 

or particulate agglomeration. 

6-33
 
Takizawa et al. (1994) investigated the CDD/CDF content of volcanic dust fall from two 

active volcanoes in Japan (Mt. Fugendake and Sakurajima). The study was not designed to 

determine whether the CDDs/CDFs observed were formed by the volcanoes or were scavenged 

from the atmosphere by the falling dust and ash. The dust fall was collected for one-month 

periods during July and October 1992; two samples of the volcanic ash were collected in 1992. 

The results of the sample analyses for 2,3,7,8-substituted CDDs and CDFs, presented in Table 

6-7, show that no 2,3,7,8-substituted congeners with less than seven chlorines were detected; 

however, the authors reported that non-2,3,7,8-substituted congeners in the lower-chlorinated 

congener groups were detected. 

Table 6-7. CDDs/CDFs in dust fall and ashes from volcanoes 

2,3,7,8-substituted 

congener group 

Dust fall (mg/km2/month)a Volcanic ash (ng/kg)b 

July 1992 Oct. 1992 Ash no. 1 Ash no. 2 

TCDD 

PeCDD 

HxCDD 

HpCDD 

OCDD 

TCDF 

PeCDF 

HxCDF 

HpCDF 

OCDF

 <0.5

 <0.5

 <0.5

 9.2

 14 

<0.5

 <0.5

 <0.5

 1.9

 4.2 

<0.5 

<0.5 

<0.5 

5.2 

11 

<0.5 

<0.5 

<0.5 

2.8 

1.8 

<0.1 

<0.1 

<0.1 

2.5 

1.7 

<0.1 

<0.1 

<0.1 

1.2 

<0.5 

<0.1 

<0.1 

<0.1 

1.8 

2.2 

<0.1 

<0.1 

<0.1 

1.2 

<0.5 a

Dust fall measured from the active volcano Fugendake. b

Volcanic ash measured from the active volcano Sakurajima. 

Source: Takizawa et al. (1994). 

6.10. FIREWORKS 

In order to produce various effects and illuminations, modern fireworks contain black 

powder and substances such as chlorine-based oxidizers, flame-coloring copper salts, and 

pulverized polyvinylchloride, which are known to be involved in dioxin-forming processes. 

During deflagration of pyrotechnics, core temperatures reach as high as 2,500°C, which would 

most likely inhibit the formation of organic pollutants. However, CDDs/CDFs may be generated 

in the areas adjacent to the combustion zone, where temperatures are lower and dwell times are 

longer. Therefore, CDDs/CDFs may be generated during the cooling period of the deflagration 

6-34
 

products because the temperatures of the smoke and ash are within the possible temperature 

range of dioxin formation (Fleischer et al., 1999). 

During a celebration in Oxford, England, that was accompanied by fireworks and bonfires, 

Dyke and Coleman (1995) reported a fourfold increase in CDD/CDF TEQ concentrations in the 

ambient air (see Section 6.2.2). Fleischer et al. (1999) conducted an experiment to measure the 

air emissions resulting specifically from the following seven types of fireworks: firecracker, 

cone fountain, jumping jack, whistler, sparkling rocket, roman candle, and four-color fountain. 

The paper cartridges and charges were separated from each firework and deflagrated separately in 

a steel chamber. CDD/CDF concentrations were measured both in air samples and in paper and 

ash samples. The results indicated that dioxins were not present in significant quantities in the 

air samples collected. Therefore, Fleisher et al. suspected that the increased background 

concentrations of CDDs/CDFs detected by Dyke and Coleman (1995) were due mainly to the 

bonfires and not the fireworks. However, concentrations of HpCDD and OCDD/OCDF were 

present in the paper and ash collected after the fireworks were detonated at concentrations 

ranging from less than the DL (10 ng/kg) to 1,200 ng/kg. Table 6-8 depicts the results of 

Fleischer’s tests. 

Table 6-8. Residue of HpCDD/HpCDF and OCDD/OCDF (ng/kg) in paper 

cartridges and charges of select pyrotechnic products 

Product 

Paper Cartridges Charge 

HpCDD OCDD OCDF HpCDD OCDD OCDF 

Firecracker 16 322 79 <10 535 26 

Cone fountain 111 384 22 <10 <10 <10 

Jumping Jack <10 33 24 <10 28 <10 

Whistler 22 353 121 <10 35 1200 

Sparkling rocket 30 129 12 <10 13 <10 

Roman candle <10 426 39 <10 <10 22 

Four-color fountain <10 18 <10 <10 <10 <10

 Source: Fleischer et al. (1999). 

Given the lack of information on the potential for CDD/CDF emissions from fireworks, 

emissions cannot be quantified. 

6-35 
6.11. OPEN BURNING AND OPEN DETONATION OF ENERGETIC MATERIALS 

Open burning and open detonation (OB/OD) practices are routinely used to destroy surplus 

or unserviceable energetic materials. Mitchell and Suggs (1998) conducted a study to determine 

emission factors from OB/OD. Air samples were collected for CDD/CDF analysis during four 

burns and after three detonations. The results of the study indicated that emission levels of 

CDDs/CDFs as a result of disposal of energetic materials by OB/OD were nondetectable. 

6-36
 

7. METAL SMELTING AND REFINING SOURCES OF CDDs/CDFs 

7.1. PRIMARY NONFERROUS METAL SMELTING/REFINING 

Little information has been published on the potential for the formation and 

environmental release of CDDs/CDFs from primary nonferrous metal smelting facilities. 

CDD/CDF releases from these facilities were first reported in the wastewater of a magnesium 

refining facility and in the receiving water sediments downstream of a nickel refining facility in 

Norway (Oehme et al., 1989). This study resulted in the evaluation of the potential for 

CDD/CDF releases from primary nonferrous metal smelting operations in the United States. Air 

emissions from several U.S. smelting operation facilities have been sampled. The findings of 

these studies are reviewed in the following sections. 

7.1.1. Primary Copper Smelting and Refining 

Environmental Risk Sciences (1995) prepared an analysis for the National Mining 

Association on the potential for CDD/CDF emissions from operations in the primary copper 

smelting industry. The analysis included reviewing the process chemistry and technology of 

primary copper smelting, identifying operating conditions, and comparing process stream 

compositions from seven of the eight U.S. primary copper smelters that are members of the 

National Mining Association. The analysis also included stack testing for CDDs/CDFs at two 

facilities. The stack testing involved the principal off-gas streams for copper smelters: the main 

stack stream, the plant tail gas stack stream, and vent fume exhaust (Secor International, Inc., 

1995b). The two facilities that were tested (Phelps Dodge Mining Co. in Playas, NM, and 

Cyprus Miami Mining Co. in Claypool, AZ) were selected as representative of the industry 

because of their similarity to other facilities in terms of process chemistry, process stream 

composition, and process stream temperatures. CDDs/CDFs were not detected in the air 

emissions from either facility. 

The results of the analysis indicate that although there is some potential for CDD/CDF 

formation in this industry’s operations, several factors lessen the probability, including the 

following: (a) most of the energy used to melt copper is derived from oxidation of copper sulfide 

ore minerals (CuFeS2) rather than carbon (fossil fuels), (b) low concentrations of organic carbon 

and chloride are present in raw materials and reagents, (c) high concentrations of SO2 are present 

in process gases (6 to 40% by volume), (d) high temperatures are maintained in the furnaces and 

converters (1,100 to 1,500°C), and (e) copper (II) chloride is apparently absent in process 

emissions. 

Although CDDs/CDFs have not been detected in U.S. facilities, CDD/CDF emissions 

have been measured in the stack of a primary copper smelter in Canada. In 2001, emission 

7-1
 

measurements for various persistent, bioaccumulative, and toxic substances, including 

CDDs/CDFs, were collected in Canada as a voluntary initiative under the Great Lakes Binational 

Toxics Strategy (Cianciarelli, 2001). One of the facilities tested was the Falconbridge Kidd 

Metallurgical plant in Timmins, Ontario, a copper smelting plant. Emission summaries are 

provided in Table 7-1 as TEQ concentrations corrected for 11% oxygen for the average of three 

runs. The total concentrations for the three runs were 3.8, 1.7, and 0.7 pg TEQ/m3. Annual 

CDD/CDF emission rates were estimated to be 0.002 g I-TEQ/yr. 

In 2002, Environment Canada began developing a generic dioxin/furan emissions testing 

protocol for use by the base metals smelting sector (Charles E. Napier Company, Ltd., 2002). 

Several base metals smelting and refining complexes were identified, and a summary of readily 

available published information on dioxin/furan emissions from the base metals smelter 

processes was compiled. A summary of this information is provided in Table 7-2. Four facilities 

were identified as primary copper smelters and had CDD/CDF emission concentrations ranging 

from less than 1 to 559 pg I-TEQ/dscm. 

Table 7-1. CDD/CDF emission concentrations (pg TEQ/m3 @ 11% oxygen) for 

primary copper smelters 

Congener/congener group Run 1 Run 2 Run 3 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0 

0 

0 

0.1 

0.1 

0.1 

0 

0 

0 

0 

0 

0 

0.1 

0 

0 

0.1 

0 

0 

0 

0 

0 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.7 

0.1 

2.2 

0 

0.1 

0.3 

0 

0.1 

0 

0 

0.2 

0 

0.9 

0.2 

0.1 

0.2 

0 

0.1 

0 

0 

0.1 

0 

0 

0.1 

0 

0.1 

0 

0 

0 

0 

TOTAL 3.8 1.8 0.4 

Source: Cianciarelli (2001). 

7-2 
Table 7-2. CDD/CDF emissions data from primary and secondary copper 

and secondary lead smelters 

Company, location Process units 

Emission control 

technology 

CDD/CDF emission 

concentration 

(pg I-TEQ/dscm) 

Primary copper smelters 

Norddeutsche Affinerie, 

Germany 

Outokumpu flash smelting 

furnace 

Waste heat boiler, ESP < 20 

Peirce-Smith converter ESP NA 

Falconbridge, Sudbury, 

Ontario (nickel and copper) 

Roasting 

Electric smelting 

Peirce-Smith converters 

Cyclone/ESP 559 

Noranda, Horne smelter, 

Noranda, Quebec 

Noranda reactor 

Noranda continuous converter 

ESP <1 

Noranda, Gaspe smelter, 

Murdockville, Quebec 

Reverberatory furnace ESP NA 

Peirce-Smith converter 82 

Secondary copper smelters 

Norddeutsche Affinerie, 

Germany 

Peirce-Smith converter FF <500 

Huttenwerke Kayser, 

Germany 

Blast furnace Post-combustion, waste 

heat boiler, FF 

<500 

Peirce-Smith converters 

Hearth furnace (for tin/lead) 

FF <100 

Reverberatory anode furnace Waste heat boiler, FF 

Mansfelder Kupfer und 

Messing, Germany 

Blast furnace Post-combustion, waste 

heat boiler, cooler, FF, FF 

with lime/coke injection 

<500 

Unknown company, 

Germany 

Shaft furnace Post-combustion, dry 

quench with secondary 

off-gas, FF 

<100 

Unknown company, 

Germany 

Rotary furnace FF <100–1,000 

Unknown company, 

Germany 

Rotary furnace Gas cooling, FF <100 

Unknown company, 

Germany 

Rotary furnace Gas cooling, FF, activated 

carbonized lignite 

adsorbent boxes 

<100 

ESP = Electrostatic precipitator 

FF = Fabric filter 

NA = Not available 

Source: Charles E. Napier Company, Ltd. (2002). 

7-3
 

In 1995, eight primary smelters were in operation in the United States, one of which 

closed at the end of that year (Edelstein, 1995). Total refinery production was 1.60 million 

metric tons in 1995, including 0.36 million metric tons from scrap material (Edelstein, 1995), 

and 1.13 million metric tons in 1987 (USGS, 1997a). In 2000, four primary smelters of copper 

were in operation in the United States, producing 1.61 million metric tons of copper (USGS, 

2002). 

CDD/CDF emission estimates for primary copper smelters were developed using the 

stack test data from the two tested facilities in the United States. Conservatively assuming that 

all nondetect values were present at one-half the detection limit, Environmental Risk Sciences, 

Inc. (1995) calculated the annual TEQ emission to air to be less than 0.5 g I-TEQDF in 1995 for 

the seven facilities (out of a total of eight) belonging to the National Mining Association. 

Assuming that 1987 feed and processing materials were similar, 1987 releases can be estimated 

at less than 0.5 g I-TEQDF as well. Because the number of facilities was reduced in 2000, the 

national emissions estimate was reduced proportionally to 0.29 g I-TEQ. The activity level 

estimates are assigned a high confidence rating and the emission factor estimates a medium 

rating; therefore, the emission estimates are assigned a medium confidence rating. The activity 

levels are based on comprehensive surveys. The emission factors are reasonably representative 

of emissions from the source category. 

7.1.2. Primary Magnesium Smelting and Refining 

Oehme et al. (1989) reported that the production of magnesium can lead to the formation 

of CDDs and CDFs. They estimated that 500 g of I-TEQDF were released to the environment in 

wastewater and 6 g I-TEQDF were released to air annually from a magnesium production facility 

in Norway; CDFs predominated, with a CDF-to-CDD concentration ratio of 10:1. At the time of 

sampling, the magnesium production process involved formation of magnesium oxide (MgO) 

from calcinated dolomite followed by a step in which magnesium chloride (MgCl2) was produced 

by heating MgO/coke pellets in a shaft furnace in a pure chlorine atmosphere to about 700 to 

800°C. The MgCl2 was then electrolyzed to form metallic magnesium and chloride. The 

chloride excess from the MgCl2 process and the chloride formed during electrolysis were 

collected by water scrubbers and directly discharged to the environment. The discharged 

wastewater contained 200 to 500 ppm of suspended particulate matter. All but trace quantities of 

the hexa through octa congeners were associated with the particulates; up to 10% of the tetra and 

penta congeners were present in the water phase. 

A study by the firm operating the facility (Musdalslien et al., 1998) indicated that 

installation of a water treatment system had reduced annual emissions to water to less than 1 g 

Nordic TEQ, and emissions to air had been reduced to less than 2 g Nordic TEQ. This study also 

7-4
 

presented results demonstrating that the carbon-reducing agent used in the MgCl2 production step 

and the operating conditions of the shaft furnace greatly affected the formation of CDDs/CDFs. 

Gases from the furnace were measured nine times over sampling periods of 6 to 8 hr. The 

calculated emission factor to air (i.e., before any air pollution control device [APCD] controls) 

ranged from 468 to 3,860 ng Nordic TEQ per kg of MgCl2 produced. The APCD controls 

consisted of three water scrubbers, a wet electrostatic precipitator (ESP), and an incinerator. 

From 1950 to 2000, the United States was the world’s largest producer of metallic 

magnesium (Kramer, 1995). In 1995, three magnesium production facilities were operating in 

the United States. As in the Norwegian plant, an electrolytic process (electrolysis of MgCl2) was 

used at the plants in Texas (capacity of 65,000 metric tons/yr) and Utah (capacity of 40,000 

metric tons/yr) to recover metallic magnesium from MgCl2. However, these two facilities 

reportedly used seawater and lake brines as the source of magnesium, and the procedures to 

obtain and purify MgCl2 did not involve chlorinating furnaces and carbonized pellets (Lockwood 

et al., 1981). A thermic process was used to recover magnesium from dolomite at the facility in 

Washington (capacity of 40,000 metric tons/yr) (Kramer, 1995). In thermic processes, MgO, a 

component of calcinated dolomite, is reacted with a metal such as silicon (usually alloyed with 

iron) to produce metallic magnesium. In 2000, the Magnesium Corporation of America facility 

near Rowley, UT, was the only operational magnesium smelting facility in the United States. 

Monitoring of wastewater discharges from U.S. magnesium production facilities for 

CDD/CDF content has not been reported. Wastewater discharges of CDDs/CDFs reported for 

the Norwegian facility (Oehme et al., 1989), discussed in the previous paragraphs, are not 

adequate to support development of wastewater emission factors for U.S. facilities because of 

possible differences in the processes used to manufacture MgCl2 and pollution control 

equipment. 

Monitoring of air emissions for CDD/CDF content has been reported for the Magnesium 

Corporation of America facility near Rowley, UT (Western Environmental Services and Testing, 

Inc., 2000). The average emission rates (for three tests) reported for the melt reactor stack and 

the cathode stack were 0.31 mg I-TEQDF/hr and 0.16 mg I-TEQDF/hr, respectively. 

Emissions data were judged inadequate for developing national emission estimates for 

1987 that could be included in the national inventory. The confidence in the degree to which the 

one tested facility represents emissions from the other two U.S. facilities is low. However, an 

estimate of the potential TEQ annual emissions for 1995 from U.S. primary magnesium 

production facilities can be made by assuming that the average total emission factor for the Utah 

facility measured in May 2000 (0.47 mg I-TEQDF/hr ) is representative of the other two facilities 

for magnesium production. Specifically, if it is assumed that this facility operated for 24 hr/day 

for 365 days in 1995, then the annual release in 1995 would have been 4.1 g I-TEQDF. If it is 

7-5
 

further assumed that this facility operated at 98% of its rated capacity of 40,000 metric tons/yr, 

then the production-based emission factor would be 105 ng I-TEQDF/kg of magnesium produced. 

Applying this emission factor to 98% of the industry’s production capacity in 1995 (142,000 

metric tons) yields a preliminary annual emissions estimate of 14.6 g I-TEQDF in 1995. 

In 2000, the Magnesium Corporation of America facility near Rowley, UT, was the only 

magnesium smelting facility operating in the United States. Production of primary magnesium at 

this facility was 41,000 metric tons in 2000. Using the emission factor of 105 ng I-TEQDF/kg of 

magnesium produced, the national estimate for dioxin releases in 2000 due to primary 

magnesium smelting is 4.3 g I-TEQ/yr. The emission factor has a high confidence rating because 

it was developed using data from this facility; therefore, the emissions estimate is assigned a high 

confidence level. 

7.1.3. Primary Nickel Smelting and Refining 

Oehme et al. (1989) reported that certain primary nickel refining processes generate 

CDDs and CDFs, primarily CDFs. Although the current low-temperature process used at a 

Norwegian facility was estimated to result in releases to water of only 1 g I-TEQDF/yr, a hightemperature (800°C) process to convert nickel chloride to nickel oxide that had been used for 

17 yr at the facility is believed to have resulted in significant releases in earlier years, based on 

the parts-per-billion levels of CDFs detected in aquatic sediments downstream of the facility. 

According to Kuck (1995), the only nickel mining and smelting complex in the United 

States (located in Oregon) had a capacity of 16,000 metric tons/yr. The facility had been on 

standby since August 1993 and had no production in 1994. The facility restarted operations in 

April 1995 and produced 8,290 metric tons of nickel that year. In 1998, the smelter closed 

because of low nickel prices (USGS, 2002). Monitoring for discharges of CDDs/CDFs at this 

facility has not been reported. Emissions of CDDs/CDFs were reported for a Norwegian facility 

in the late 1980s, as discussed above (Oehme et al., 1989). The emissions information contained 

in the Norwegian study is not adequate to support development of emission factors for the U.S. 

facility for 1987 and 1995. Because the facility closed in 1998, emission estimates for 2000 for 

primary nickel smelting are zero. 

7.1.4. Primary Aluminum Smelting and Refining 

No sampling of air emissions for the presence of CDDs/CDFs has been reported for this 

industry. 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 CDDs/CDFs. 

Because the primary smelting process does not use chlorine, there is widespread belief that 

7-6
 

dioxin emissions from primary aluminum smelting facilities do not exist; therefore, no sampling 

has been done. 

In the primary aluminum smelting process, bauxite ore, a hydrated oxide of aluminum 

consisting of 30 to 56% alumina (Al2O3), is refined into alumina by the Bayer process. The 

alumina is then shipped to a primary aluminum smelter for electrolytic reduction to aluminum. 

Electrolytic reduction of alumina occurs in shallow rectangular cells, or pots, which are steel 

shells lined with carbon. Carbon electrodes (petroleum coke mixed with a pitch binder) 

extending into the pot serve as the anodes, and the carbon lining serves as the cathode. Three 

types of pots are used: prebaked anode cell, horizontal stud Soderberg anode cell, and vertical 

stud Soderberg anode cell. Most of the aluminum produced in the United States is produced 

using the prebaked cells. Molten cryolite (Na3AlF6) functions as both the electrolyte and the 

solvent for the aluminum. Aluminum is deposited on the cathode as molten metal (U.S. EPA, 

1998a). 

Prior to casting, the molten aluminum may be batch treated in reverberatory furnaces 

(such as those used in secondary aluminum smelting) to remove oxides, gaseous impurities, and 

active metals such as sodium and magnesium. One process consists of adding a flux of chloride 

and fluoride salts and then bubbling chlorine gas through the molten mixture (U.S. EPA, 1998a). 

U.S. production of primary aluminum was 3.343 million metric tons in 1987 and 3.375 

million metric tons in 1995. In 1995, 13 companies operated 22 primary aluminum reduction 

plants (USGS, 1997b, c). In 2000, 12 companies operated 23 primary aluminum reduction 

plants, and primary aluminum smelters produced 3.7 million metric tons of aluminum (USGS, 

2002). Because emission factors have not been developed for this sector, there are no emission 

estimates for this category. 

7.1.5. Primary Titanium Smelting and Refining 

It has been suggested that carbochlorination processes used in this industry may be a 

source of CDDs/CDFs (Bramley,1998; ERG, 1998). As discussed below, CDDs/CDFs have 

been measured in titanium dioxide production sludges. A brief summary of the processes used in 

this industry is presented in the following paragraphs. 

In primary titanium smelting, titanium oxide ores and concentrates are chlorinated in 

fluidized-bed reactors in the presence of coke at 925 to 1,010°C to form titanium tetrachloride 

(TiCl4). The TiCl4 is separated from other chlorides by double distillation. The TiCl4 is then 

either oxidized at 985°C to form pigment-grade titanium dioxide or reduced using sodium or 

magnesium to form titanium sponge (i.e., metallic titanium) (Knittel, 1983). Titanium ingot is 

produced by melting titanium sponge or scrap or a combination of both using electron beam, 

7-7
 

plasma, and vacuum arc methods. Scrap currently supplies about 50% of ingot feedstock 

(Gambogi, 1996). 

Titanium sponge is produced at two facilities in the United States, one in Albany, OR, 

and the other in Henderson, NV. In 1995, the U.S. production volume of titanium sponge was 

withheld to avoid disclosing proprietary data; domestic sponge capacity was 29,500 metric 

tons/yr. In 1987, U.S. production of titanium sponge was 17,849 metric tons. 

More than 90% of titanium dioxide is produced using the process described above. 

Titanium dioxide pigment is used in paints, plastics, and paper products. In 1995, titanium 

dioxide was produced at nine facilities in the United States. Production volumes in 1987 and 

1995 were 821,000 and 1.8 million metric tons, respectively (Gambogi, 1996; USGS, 1997d). In 

2000, four companies at eight facilities in seven states produced 1.44 million metric tons of 

titanium dioxide (USGS, 2002). 

Titanium dioxide production creates a sludge waste, and CDDs/CDFs have been 

measured in these sludges (U.S. EPA, 2001c). For the most part, these sludges have been 

disposed of in either on-site or off-site RCRA Subtitle D solid waste disposal facilities. 

However, given the potential for leaching of the heavy metals from the sludge in the Subtitle D 

landfill, EPA has listed this waste as hazardous waste under Subtitle C. These sludges are now 

considered a hazardous waste under RCRA and must be disposed of in permitted landfills (U.S. 

EPA, 2001c). Therefore, they are not considered to cause environmental releases under the 

definition in this document and are not included in the inventory. 

7.2. SECONDARY NONFERROUS METAL SMELTING 

Secondary smelters primarily engage 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 and refining processes for some metals (e.g., aluminum, 

copper, and magnesium) use chemicals such as sodium chloride, potassium chloride, 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/CDFs, as evidenced by 

their detection in the stack emissions of secondary aluminum, copper, and lead smelters (Aittola 

et al., 1992; U.S. EPA, 1987a, 1997a). 

7.2.1. Secondary Aluminum Smelters 

Secondary aluminum smelters reclaim aluminum from scrap using two processes: 

precleaning and smelting. Both processes may produce CDD/CDF emissions. 

7-8
 
Precleaning processes involve sorting and cleaning scrap to prepare it for smelting. Cleaning 

processes that may produce CDD/CDF emissions use heat to separate aluminum from 

contaminants and other metals. These techniques are “roasting” and “sweating.” Roasting uses 

rotary dryers with a temperature high enough to vaporize organic contaminants but not high 

enough to melt aluminum. An example of roasting is the delacquering and processing of used 

beverage cans. Sweating involves heating aluminum-containing scrap metal to a temperature 

above the melting point of aluminum but below the melting temperature of other metals such as 

iron and brass. The melted aluminum trickles down and accumulates in the bottom of the sweat 

furnace and is periodically removed (U.S. EPA, 1997a). 

After precleaning, the treated aluminum scrap is smelted and refined. This usually takes 

place in a reverberatory furnace. Once smelted, flux is added to remove impurities. The melt is 

demagged to reduce the magnesium content of the molten aluminum by adding chlorine gas. The 

molten aluminum is then transferred to a holding furnace and alloyed to final specifications (U.S. 

EPA, 1997a). 

CDD/CDF emissions to air have been measured at seven U.S. secondary aluminum 

operations. Five facilities were tested in 1995 and 1996 and two facilities were tested in 1992. 

Four of the 1995 tests were conducted by EPA in conjunction with The Aluminum Association, 

Inc., to identify emission rates from facilities with potentially maximum-achievable-controltechnology-grade operations and APCD equipment. The fourth test was performed by EPA (U.S. 

EPA, 1995c). Results from two facilities tested by the California Air Resources Board (CARB) 

in 1992 were presented in two confidential reports. 

The first facility tested in 1995 was a top-charge melt furnace (Advanced Technology 

Systems, Inc., 1995). During testing, the charge material to the furnace was specially formulated 

to contain no oil, paint, coatings, rubber, or plastics other than incidental amounts. The TEQ 

emission factor from such a clean charge, 0.27 ng TEQDF-WHO98/kg (0.26 ng I-TEQDF/kg) charge 

material, would be expected to represent the low end of the normal industry range. 

The second facility operated a sweat furnace to preclean the scrap and a reverberatory 

furnace to smelt the precleaned aluminum (U.S. EPA, 1995c). Stack emissions were controlled 

by an afterburner operated at 788°C. The TEQ emission factor for this facility was 3.37 ng 

TEQDF-WHO98/kg (3.22 ng I-TEQDF/kg) aluminum produced. 

The third facility employed a crusher/roasting dryer as a precleaning step followed by a 

reverberatory furnace (Galson Corporation, 1995). The emissions from the two units were 

vented separately. The exhaust from the crusher/dryer was treated with an afterburner and a 

fabric filter (FF). The exhaust from the furnace passed through an FF with lime injection. Both 

stack exhausts were tested, and the combined TEQ emission factor was 13.55 ng TEQDFWHO98/kg (12.95 ng I-TEQDF/kg) aluminum produced. Because the activity level of the facility 

7-9
 

at the time of sampling was treated as confidential business information, the calculated emission 

factor was based on the reported typical production rates of the two operations: 26,000 lb/hr for 

the crusher/dryer and 6,700 lb/hr for the furnace. 

The fourth facility operated a scrap roasting dryer followed by a sidewell reverberatory 

furnace (Weston,1996). The emissions from the two units were vented separately. Exhaust from 

the dryer passed through an afterburner and a lime-coated FF. The exhaust from the furnace 

passed through a lime-coated FF. Both stack exhausts were tested, and the combined TEQ 

emission factor was 8.52 ng TEQDF-WHO98/kg (7.93 ng I-TEQDF/kg) of charge material. 

The two facilities tested by CARB in 1992, which were reported in two confidential 

reports (CARB, 1992a, b, as reported in U.S. EPA, 1997a) had TEQ emission factors of 55.68 

and 23.44 ng TEQDF-WHO98/kg (52.21 and 21.67 ng I-TEQDF/kg) of scrap aluminum consumed. 

One facility was equipped with a venturi scrubber; the other was assumed to be uncontrolled 

(U.S. EPA, 1997a). 

The seventh facility was tested in 1995 by Commonwealth Aluminum Corp. (1995). The 

test involved sampling air emissions from a delaquering kiln. The facility was equipped with an 

FF as the primary APCD. 

The CDD/CDF congener and congener group emission factors derived from these stack 

tests were used to represent emissions from secondary aluminum facilities operating in 2000, 

1995, and 1987. Table 7-3 shows the TEQ emission factors for facilities operating in 2000, and 

Table 7-4 shows those for 1995 and 1987. The 2000 emission factors do not include the results 

from the two facilities tested by CARB (1992a, b) because it was assumed that all facilities 

operating in 2000 were equipped with APCDs on all vents and stacks. The average congener and 

congener group profiles are presented in Figure 7-1. The average of the TEQ emission factors 

measured at the five tested facilities representative of the year 2000 is 5.17 ng TEQDF-WHO98/kg 

(4.90 ng I-TEQDF/kg) of scrap feed. A low confidence rating is assigned to the average emission 

factor for 2000 because it is based on the results of testing at only five facilities, and they may 

not be representative of all facilities operating in the United States in that year. 

The emission factors for facilities operating in 1995 and 1987 do include the results from 

the two facilities tested by CARB in 1992. The average of the TEQ emission factor measured at 

the seven tested facilities representative of the years 1995 and 1987 is 15.0 ng TEQDF-WHO98/kg 

(14.05 ng I-TEQDF/kg) of scrap feed. A low confidence rating is assigned to the average 

emission factor for 1995 and 1987 because it is based on the results of testing at only seven 

facilities, and that may not be representative of all facilities operating in the United States in 

those reference years. 

7-10
 
Table 7-3. CDD/CDF emission factors (ng/kg scrap feed) for secondary 

aluminum smelters for 2000 

Congener 

Mean facility emission factor Overall 

mean 

emission 

a b c d e factor 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND (0.01) 

0.02 

0.05 

0.13 

0.15 

0.51 

0.42 

0.13 

0.39 

0.24 

0.86 

1.26 

7.67 

14.97 

0.51 

1.19 

1.35 

1.52 

2.51 

2.6 

1.01 

0.40 

1.19 

0.72 

0.94 

1.62 

3.49 

NR 

0.01 

0.02 

0.02 

0.03 

0.05 

0.10 

NR 

0.21 

0.56 

0.48 

0.70 

1.12 

2.87 

5.47 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.44 

0.06 

0.17 

0.32 

0.11 

0.02 

0.3 

0.07 

0.03 

0.3 

0.74 

1.51 

2.44 

2.44 

2.69 

1.02 

3.82 

11.39 

5.5 

30.4 

14.20 

10.47 

11.06 

21.84 

7.1 

0.47 

7.09 

14.61 

1.21 

3.15 

10.46 

12.52 

6.98 

8.14 

2.74 

0.14 

4.30 

3.18 

0.45 

1.23 

0.07 

0.08 

0.12 

0.16 

0.06 

0.01 

0.08 

0.17 

0.04 

0.06 

5.18 

4.93 

4.15 

6.58 

2.54 

0.33 

3.12 

5.88 

1.45 

7.03 

Total I-TEQDF 

f 0.26 3.22 12.95 8.09 0.14 4.93 

Total TEQDF-WHO98 

f 0.27 3.37 13.55 8.68 0.15 5.20 

a

Source: Advanced Technology Systems, Inc. (1995).
 
 b

Source: U.S. EPA (1995c).
 
 c

Source: Galson Corporation (1995).
 
 d

Source: Weston (1996).
 
 e

Source: Commonwealth Aluminum Corp. (1995).
 
 f

TEQ calculations assume nondetect values were zero. 
 

NR = Not reported
 


ND = Not detected (value in parenthesis is the detection limit)
 


For comparison purposes, the European Commission uses 22 ng I-TEQDF/kg scrap 

aluminum as the typical emission factor for the European Dioxin Inventory (Quab and Fermann, 

1997). Umweltbundesamt (1996) reported stack testing results for 25 aluminum smelters and 

foundries in Germany. This study provided sufficient data to enable calculation of TEQ emission 

factors for 11 of the tested facilities. The calculated emission factors ranged from 0.01 to 167 ng 

7-11
 

Table 7-4. CDD/CDF emission factors (ng/kg scrap feed) for secondary 

aluminum smelters for 1995 and 1987 

Congener 

Mean facility emission factor Overall 

mean 

emission 

a b c d e f f factor 

2,3,7,8-TCDD ND (0.01) 0.13 0.51 0.40 0.01 1.97 0.85 0.54 

1,2,3,7,8-PeCDD 0.02 0.39 1.19 1.19 0.02 7.1 3.64 1.94 

1,2,3,4,7,8-HxCDD 0.05 0.24 1.35 0.72 0.02 4.26 2.82 1.35 

1,2,3,6,7,8-HxCDD 0.13 0.86 1.52 0.94 0.03 5.3 4.12 1.84 

1,2,3,7,8,9-HxCDD 0.15 1.26 2.51 1.62 0.05 5.3 2.02 1.84 

1,2,3,4,6,7,8-HpCDD 0.51 7.67 2.6 3.49 0.10 28.9 19.3 8.94 

OCDD 0.42 14.97 1.01 NR NR 33.2 24.3 14.78 

2,3,7,8-TCDF 0.44 0.74 14.20 10.46 0.07 23.2 4.84 7.71 

1,2,3,7,8-PeCDF 0.06 1.51 10.47 12.52 0.08 33.8 1.18 8.52 

2,3,4,7,8-PeCDF 0.17 2.44 11.06 6.98 0.12 48 23.3 13.15 

1,2,3,4,7,8-HxCDF 0.32 2.44 21.84 8.14 0.16 46.1 17.6 13.80 

1,2,3,6,7,8-HxCDF 0.11 2.69 7.1 2.74 0.06 46.1 16.9 10.81 

1,2,3,7,8,9-HxCDF 0.02 1.02 0.47 0.14 0.01 22 1.35 3.57 

2,3,4,6,7,8-HxCDF 0.3 3.82 7.09 4.30 0.08 39 16 10.08 

1,2,3,4,6,7,8-HpCDF 0.07 11.39 14.61 3.18 0.17 122 42.6 27.71 

1,2,3,4,7,8,9-HpCDF 0.03 5.5 1.21 0.45 0.04 27.1 6.2 5.79 

OCDF 0.3 30.4 3.15 1.23 0.06 60.5 29.5 17.88 

Total I-TEQDF

g 0.26 3.22 12.95 8.09 0.14 52.21 21.68 14.08 

Total TEQDF-WHO98g 0.27 3.37 13.55 8.68 0.15 55.68 23.45 15.02 

a

Source: Advanced Technology Systems, Inc. (1995).
 
 b

Source: U.S. EPA (1995c).
 
 c

Source: Galson Corporation (1995).
 
 d

Source: Weston (1996).
 
 e

Source: Commonwealth Aluminum Corp. (1995).
 
 f

Source: CARB (1992a, b), as reported in U.S. EPA (1997a).
 
 g

TEQ calculations assume nondetects were zero.
 


NR = Not reported
 


ND = Not detected (value in parenthesis is the detection limit)
 


I-TEQDF/kg of scrap feed. Three facilities had emission factors exceeding 100 ng I-TEQDF/kg, 

and two facilities had emission factors of less than 1 ng I-TEQDF/kg. The mean emission factor 

for the 11 facilities was 42 ng I-TEQDF/kg. 

Approximately 727,000 metric tons of scrap aluminum were consumed by 67 secondary 

aluminum smelters in the United States in 1987 (U.S. DOC, 1995c). In 1995, consumption of 

7-12
 

Figure 7-1. Congener and congener group profiles for air emissions from 

secondary aluminum smelters. 

Sources: U.S. EPA (1995c); Galson Corporation (1995). 

scrap aluminum by the 76 facilities that composed the secondary aluminum smelting industry 

had nearly doubled, to 1.3 million metric tons (USGS, 1997e; The Aluminum Association, Inc., 

1997). In 2000, secondary aluminum smelters consumed 1.6 million metric tons of scrap 

aluminum (USGS, 2002). A high confidence rating is assigned to these production estimates 

7-13
 

because they are based on government survey data. Applying the 15.0 ng TEQDF-WHO98/kg 

(14.05 ng I-TEQDF/kg) emission factor to the metric tons of scrap aluminum processed in 1987 

and 1995 yields estimated annual emissions of 10.9 g TEQDF-WHO98 (10.21 g I-TEQ) in 1987 

and 19.5 TEQDF-WHO98 (18.27 g I-TEQDF) in reference year 1995. Applying the emission factor 

of 5.17 ng TEQDF-WHO98/kg (4.9 ng I-TEQDF/kg) to the metric tons of scrap aluminum processed 

in 2000 yields estimated annual emissions of 8.27 g TEQDF-WHO98 (7.84 g I-TEQ) in reference 

year 2000. These emission estimates are assigned a low confidence rating because the rating 

given to the emission factor was low. 

It should be noted that a significant amount of scrap aluminum is also consumed by other 

segments of the aluminum industry. However, this scrap is generally from metal manufacturing 

processes, including metal and alloy production (e.g., borings, turnings, and dross), rather than 

old scrap that results from recycling of consumer products (e.g., cans, radiators, auto shredders). 

In 1995, integrated aluminum companies consumed 1.4 million metric tons of scrap aluminum, 

and independent mill fabricators consumed 0.68 million metric tons (USGS, 1997e). 

7.2.2. Secondary Copper Smelters 

Secondary copper smelting is part of the scrap copper, brass, and bronze reprocessing 

industry. Brass is an alloy of copper and zinc; bronze is an alloy of copper and tin. Facilities in 

this industry fall into three general classifications: secondary smelting, ingot making, and 

remelting. Similar processing equipment may be used at all three types of facilities, so the 

distinguishing features are not immediately apparent (U.S. EPA, 1994b). 

The feature that distinguishes secondary smelters from ingot makers and remelters is the 

extent to which pyrometallurgical purification is performed. A typical charge at a secondary 

smelter may contain from 30 to 98% copper. The secondary smelter upgrades the material by 

reducing the quantity of impurities and alloying materials, thereby increasing the relative 

concentration of copper. This degree of purification and separation of the alloy constituents does 

not occur at ingot makers and remelters. Feed material to a secondary copper smelter is a 

mixture of copper-bearing scrap such as tubing, valves, motors, windings, wire, radiators, 

turnings, mill scrap, printed circuit boards, telephone switching gear, and ammunition casings. 

Nonscrap items such as blast furnace slags and drosses from ingot makers or remelters may 

represent a portion of the charge. Secondary smelter operators use a variety of processes to 

separate the alloy constituents. Some purify the scrap in the reductive atmosphere of a blast 

furnace and then purify the charge in the oxidizing atmosphere of a converter; others perform all 

purification by oxidation in top-blown rotary converters or in reverberatory furnaces (U.S. EPA, 

1994b). 

7-14
 
Ingot maker operators blend and melt scrap copper, brass, and bronze of various 

compositions to produce a specification brass or bronze ingot. When necessary, they add ingots 

of other metals (e.g., zinc or tin) to adjust the metallurgy of the final product. The feed materials 

for ingot makers contain relatively high amounts of copper. Examples of feed materials include 

copper tubing, valves, brass and bronze castings, ammunition shell casings, and automobile 

radiators. “Fire-refined” anode copper or cathode copper may also be charged. Items such as 

motors, telephone switchboard scrap, circuit board scrap, and purchased slags are not used by 

ingot makers. The reductive step (melting in a reducing atmosphere, as in a blast furnace) that 

some secondary smelters employ is not used by ingot makers. Ingot makers do, however, use 

some of the other types of furnaces used by secondary smelters, including direct-fired converters, 

reverberatory furnaces, and electric induction furnaces (U.S. EPA, 1994b). 

Remelting facilities do not conduct substantial purification of the incoming feeds. These 

facilities typically melt the charge and then cast or extrude a product. The feeds to a remelter are 

generally alloy material of approximately the desired composition of the product (U.S. EPA, 

1994b). 

7.2.2.1. Emissions Data 

Stack emissions of CDDs/CDFs from a secondary copper smelter were measured by EPA 

during 1984 and 1985 as part of the National Dioxin Tier 4 Study (U.S. EPA, 1987a). The 

facility chosen for testing was estimated to have a high potential for CDD/CDF emissions 

because of the abundance of chlorinated plastics in the feed. This facility ceased operations in 

1986. The facility was chosen for testing by EPA because the process technology and APCD 

equipment in place were considered typical for the source category. 

During operations, copper and iron-bearing scrap were fed in batches to a cupola blast 

furnace, which produced a mixture of slag and black copper. Approximately 4 to 5 tons of 

metal-bearing scrap were fed to the furnace per charge, with materials typically being charged 10 

to 12 times per hour. Coke fueled the furnace and represented approximately 14% (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 floorcleaning material. The telephone scrap made up 22% (by weight) of the feed and was the only 

scrap component that contained plastic materials. Oxygen-enriched combustion air for 

combustion of the coke was blown through tuyeres (nozzles) 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 FFs, and the flue gas was then 

discharged into a common stack. The estimated emission factors derived for this site are 

7-15
 

presented in Table 7-5. The emission factors are based on the total weight of scrap fed to the 

furnace. Based on the measured congener and congener group emission factors, the TEQ 

emission factor is 779 ng I-TEQDF/kg (810 ng TEQDFWHO98/kg) of scrap metal smelted. Figure 

7-2a presents the congener group profile based on these emission factors. 

In 1992, stack testing of the blast furnace emissions of a secondary smelter located in 

Philadelphia, PA (Franklin Smelting and Refining Co.), was conducted by Applied Geotechnical 

& Environmental Services Corporation (AGES, 1992). Like the facility tested by EPA during 

1984 and 1985, this facility processed low-purity copper-bearing scrap, telephone switch gear, 

and slags, as well as higher-copper-content materials (U.S. EPA, 1994b). The facility used a 

blast (cupola-type) furnace coupled with a pair of rotary converters to produce blister copper. 

The blast furnace used coke as both the fuel and the agent to maintain a reducing atmosphere. 

The black copper-slag mixture from the blast furnace was charged to the rotary converters for 

further refining with the aid of oxygen, sand, and oak logs (AGES, 1992; U.S. EPA, 1994b). The 

APCD equipment installed on the blast furnace included an afterburner, a cooling tower, and an 

FF. During testing, the afterburner was reported to be operating erratically and was particularly 

low during one of the two sampling episodes. Stack gas flow was also low during both sampling 

episodes because one or more FF compartments were inoperable (AGES, 1992). The estimated 

emission factors derived for this site from the AGES results are presented in Table 7-5. The 

emission factors were based on the total weight of scrap fed to the blast furnace. The TEQ 

emission factor was 16,618 ng I-TEQDF/kg (16,917 ng TEQDF-WHO98/kg) of scrap. Figure 7-2b 

presents the congener and congener group profiles based on these emission factors. 

In 1991, stack testing of the rotary furnace stack emissions of a secondary smelter 

(Chemetco, Inc.) located in Alton, IL, was conducted by Sverdrup Corp. (1991). The Chemetco 

facility used four tap-down rotary (i.e., oxidizing) furnaces. Furnace-processed gas emissions 

were controlled by a primary quencher and a venturi scrubber. The feed was relatively highpurity copper scrap containing minimal, if any, plastics. The same manufacturing process and 

APCD equipment were in place in 1987 and 1995 (U.S. EPA, 1994b). Because this facility 

operated under oxidizing rather than reducing conditions and processed relatively high-purity 

scrap, the potential for CDD/CDF formation and release was expected to be dramatically 

different from that of the two tested facilities reported above. The estimated emission factors 

derived for this site from the results of Sverdrup Corp. (1991) are presented in Table 7-5. The 

emission factors were based on the total weight of scrap feed going to the furnace. The TEQ 

emission factor was 3.60 ng I-TEQDF/kg (3.66 ng TEQDF-WHO98/kg) of scrap. 

7-16
 

Table 7-5. CDD/CDF mean emission factors (ng/kg scrap feed) for secondary 

copper smelters 

Congener/ 

congener group 

EPA Tier 4 

facilitya,b 

Franklin smelting 

facilityc 

Chemetco smelting 

facilityd 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

127 

NR 

NR 

NR 

NR 

NR 

1,350 

227 

846 

1,476 

1,746 

2,132 

17,065 

55,668 

ND (0.05) 

0.21 

0.39 

0.70 

1.26 

8.95 

22.45 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

2,720 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

2,520 

4,457 

9,455 

5,773 

70,742 

20,524 

5,362 

12,082 

37,251 

7,570 

82,192 

2.11 

1.47 

2.63 

7.30 

2.15 

4.06 

0.27 

11.48 

2.74 

21.61 

Total I-TEQDF e 779f 16,618 3.59 

Total TEQDF-WHO98 e 810f 16,917 3.66 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

736 

970 

1,260 

2,080 

1,350 

13,720 

8,640 

4,240 

3,420 

2,520 

14,503 

30,248 

55,765 

38,994 

55,668 

108,546 

71,136 

164,834 

66,253 

82,192 

3.05 

5.19 

9.62 

16.71 

22.45 

46.42 

27.99 

27.96 

23.38 

21.61 

Total CDD/CDF 38,936 688,139 204.38 a

No nondetect values were reported for 2,3,7,8-TCDD, 2,3,7,8-TCDF, or any congener group in the three test runs.
 
 b

Source: U.S. EPA (1987a).
 
 c

Source: AGES (1992).
 
 d

Source: Sverdrup Corp. (1991).
 
 e

TEQ calculations assume nondetect values were zero.
 
 f Estimated using the measured data for 2,3,7,8-TCDD, 2,3,7,8-TCDF, OCDD, and OCDF and congener group 
 

emissions (i.e., for the penta-, hexa-, and hepta-CDDs and CDFs, it was assumed that the measured emission factor
 


within a congener group was the sum of equal emission factors for all congeners in that group, including non-
 

2,3,7,8-substituted congeners).
 


NR = Not reported
 


ND = Not detected (value in parenthesis is the detection limit)
 


7-17
 
Figure 7-2a. Congener group profile for air emissions from a secondary 

copper smelter. 

Source: U.S. EPA (1987d). 

Only limited data on emissions from secondary copper smelters are reported in the 

European Dioxin Inventory (LUA, 1997). TEQ emission factors reported for German shaft 

furnaces/converters and reverberatory furnaces range from 5.6 to 110 ng I-TEQDF/kg and from 

0.005 to 1.56 ng I-TEQDF/kg, respectively. Emission factors reported for two smelter and casting 

furnaces in Sweden in which relatively clean scrap is used as input are 0.024 and 0.04 ng I

TEQDF/kg. A smelter in Austria is reported to have a TEQ emission factor of 4 ng I-TEQDF/kg. 

The minimum, typical, and maximum default emission factors selected in LUA (1997) are 5, 50, 

and 400 ng I-TEQDF/kg, respectively. 

In the 2002 Environment Canada report on CDD/CDF emissions from the base metals 

smelting sector (Charles E. Napier Company, Ltd., 2002), three secondary copper smelters were 

identified (see Table 7-2). CDD/CDF emission concentrations were reported as ranging from 

less than 100 to less than 500 pg I-TEQ/dscm. 

7-18
 

­
Figure 7-2b. Congener and congener group profiles for a closed secondary 

copper smelter. 

Source: AGES (1992). 

7-19
 

7.2.2.2. Activity Level Information 

In 1987, four secondary copper smelters were in operation: Franklin Smelting and 

Refining Co. (Philadelphia, PA), Chemetco, Inc. (Alton, IL), Southwire Co. (Carrollton, GA), 

and a facility located in Gaston, SC, that was owned by American Telephone and Telegraph 

(AT&T) until 1990, when it was purchased by Southwire Co. In 1987, estimated smelter 

capacities were 13,600 metric tons for the Franklin facility, 120,000 metric tons for the 

Chemetco facility, 48,000 metric tons for the Southwire facility, and 85,000 metric tons for the 

AT&T facility (telephone conversation on August 17, 1999, between D. Edelstein, U.S. 

Geological Survey, and G. Schweer, Versar, Inc.). In 1995, only three of these four facilities 

were still in operation. The Southwire facility in Gaston was closed in January 1995. The 

Franklin facility ceased operations in August 1997. Estimated smelter capacities in 1995 were 

16,000 metric tons for the Franklin facility, 135,000 metric tons for the Chemetco facility, and 

92,000 metric tons for the Southwire Georgia facility (Edelstein, 1999). In May 2000, the 

Southwire Co. closed its Georgia facility and ceased operations (Edelstein, 2000). In November 

2001, Chemetco closed its facility and ceased operations (Edelstein, 2001). 

According to Edelstein (2001), smelters and refineries consumed 255,000 metric tons of 

purchased copper-based scrap in 2000 and 196,000 metric tons in 2001. Assuming Chemetco 

was the sole smelter facility operating in 2001, and that it operated for 10 of 12 months in 2001, 

its estimated annual consumption of copper-based scrap would be 235,000 metric tons per year. 

Assuming Chemetco’s annual consumption rate did not change from 2000 to 2001, the estimated 

consumption of copper-based scrap for the Southwire Co. in 2000 was 20,000 metric tons. 

7.2.2.3. Emission Estimates 

Although little research has been done to define the CDD/CDF formation mechanisms in 

secondary copper smelting operations, two general observations have been made (Buekens et al., 

1997). First, the presence of chlorinated plastics in copper scraps used as feed for smelters is 

believed to increase CDD/CDF formation. Second, the reducing or pyrolytic conditions in blast 

furnaces can lead to high CDD/CDF concentrations in the furnace process gases. As noted in 

Section 7.2.2.1, two of the U.S. facilities that have been tested (U.S. EPA, 1987a; AGES, 1992) 

had the following characteristics: both processed low-purity copper-bearing scrap containing 

significant quantities of plastics and telephone switch gears, and both used blast furnaces. The 

APCD equipment at both facilities consisted of an afterburner, a cooling tower (Franklin facility 

only), and an FF (U.S. EPA, 1994b). The other tested U.S. facility used oxidizing rather than 

reducing conditions and processed relatively high-purity scrap (Sverdrup Corp., 1991). 

7-20
 

Annual TEQ emissions for 1987, 1995, and 2000 were derived as the sum of the TEQ 

emissions for each secondary copper facility in operation during the reference years. The 

following discussion summarizes the procedure used to estimate annual TEQ air emissions. 

The Franklin Smelting facility operated in 1987 and 1995 but not in 2000. The TEQ 

emission factor measured at this facility in 1992 is assumed to be representative of the TEQ 

emission factors in 1987 and 1995. Combining this emission factor (16,618 ng I-TEQDF/kg 

[16,917 ng TEQDF-WHO98/kg] of scrap feed) with the estimated smelter capacities (data are not 

available on the amount of scrap processed) for this facility in 1987 (13,600,000 kg) and 1995 

(16,000,000 kg) yields TEQ emission estimates of 226 g I-TEQ DF (230 g TEQDF-WHO98) in 1987 

and 266 g I-TEQDF (271 g TEQDF-WHO98) in 1995. This facility ceased operations in 1997. 

The Chemetco facility operated in 1987, 1995, and 2000. Similarly, for purposes of this 

report, the TEQ emission factor for the Chemetco facility is considered to be representative of the 

TEQ emission factor for this facility for 1987, 1995, and 2000. Combining this emission factor 

(3.60 ng I-TEQDF/kg [3.66 ng TEQDF-WHO98/kg] of scrap feed) with the estimated smelter 

capacities of 120,000,000 kg in 1987 and 135,000,000 kg in 1995 yields TEQ estimates of 0.43 g 

I-TEQDF (0.44 g TEQDF-WHO98) in 1987 and 0.49 g I-TEQDF (0.49 g TEQDF-WHO98) in 1995. 

Combining the same emission factor with the scrap consumption for this facility in 2000 

(235,000,000 kg) yields a TEQ estimate of 0.85 g I-TEQDF (0.86 g TEQDF-WHO98) for 2000. 

The facility in Gaston, SC, was in operation during 1987 but ceased operations in 1995. 

Prior to 1990, when this facility was owned by AT&T, the plant processed a great deal of high

plastics-content scrap (such as whole telephones). This scrap was fed to a pyrolysis unit prior to 

entering the blast furnace. In addition to a blast furnace, the facility also had an oxidizing 

reverberatory furnace for processing higher-purity scrap. The facility had separate FFs for the 

blast furnace, the converters, and the reverberatory furnace (U.S. EPA, 1994b). Because this 

facility processed low-purity, high-plastics-content scrap in 1987, and presumably processed 

much of this in the reducing atmosphere of a pyrolysis unit and blast furnace, the average of the 

TEQ emission factors for the Tier 4 EPA-tested facility (U.S. EPA, 1987a) and the Franklin 

facility (8,700 ng I-TEQDF/kg [8,860 ng TEQDF-WHO98/kg]) was used to estimate potential 

emissions in 1987 of 740 g I-TEQDF (753 g TEQDF-WHO98) (assuming an activity level of 

85,000,000 kg). This activity level is the estimated capacity of the facility; data were not 

available on the amount of scrap processed. 

The Southwire facility had both a blast furnace and a reverberatory furnace. In 1992, 

approximately 50% of incoming scrap was processed in each furnace (U.S. EPA, 1994b). Unlike 

the Franklin, Chemetco, and Gaston secondary copper smelters, the Southwire facility stopped 

processing plastic-coated scrap in the 1970s. In addition, this facility had a more complex APCD 

system, which may have reduced the formation and release of CDDs/CDFs. The blast furnace­

­

7-21
 

processed gases passed through an afterburner (871°C), U-tube coolers, and an evaporative spray 

system before entering the FF at a temperature of 107 to 191°C. For these reasons, EPA has 

determined that the existing emissions data for secondary smelters cannot reliably be used to 

generate a quantitative estimate of potential emissions during 1987, 1995, or 2000 for this 

facility. 

Total secondary copper smelter emissions for 1987 are the sum of the Franklin smelting 

facility emissions (271 g TEQDF-WHO98 [266 g I-TEQDF]), the Chemetco smelter facility (0.44 g 

TEQDF-WHO98 [0.43 g I-TEQDF]) and the Gaston, SC, facility (753 g TEQDF-WHO98 [740 g ITEQDF]). Total secondary copper smelter emissions for 1987 are 983.44 g TEQDF-WHO98 

(966.43 g I-TEQDF). 

Total secondary copper smelter emissions for 1995 are the sum of the Franklin smelting 

facility emissions (271 g TEQDF-WHO98 [266 g I-TEQ DF]) and the Chemetco smelter facility 

(0.49 g TEQDF-WHO98 [0.49 g TEQDF-WHO98]). Total secondary copper smelter emissions for 

1995 are 271.49 g TEQDF-WHO98 (266.49 I-TEQDF). 

The Chemetco smelter provides the TEQ emissions estimate for the year 2000. Total 

secondary copper smelter emissions for 2000 are 0.86 g TEQDF-WHO98 (0.85 g I-TEQDF). 

A high confidence rating is assigned to the production and consumption estimates 

because they are based on government survey data. A low confidence rating is assigned to the 

TEQ emission estimates because they are based on limited measurements made at three smelters, 

one of which was not in operation in 1987 or 1995. 

It should be noted that a significant amount of scrap copper is consumed by other 

segments of the copper industry. In 1995 and 2000, brass mills and wire-rod mills consumed 

886,000 and 1,070,000 metric tons of copper-based scrap, respectively; foundries and 

iscellaneous manufacturers consumed 71,500 and 96,200 metric tons, respectively (USGS, 

1997e; Edelstein, 2001). As noted above, however, these facilities generally do not conduct any 

significant purification of the scrap. Rather, the scrap consumed is already of alloy quality, and 

processes employed typically involve only melting, casting, and extruding. Thus, the potential 

for formation of CDDs/CDFs is expected to be much less than the potential during secondary 

smelting operations. 

7.2.3. Secondary Lead Smelters 

The secondary lead smelting industry produces elemental lead through the chemical 

reduction of lead compounds in a high-temperature furnace (1,200 to 1,260°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 make up the 

current secondary lead smelting industry in the United States. Of the 45 furnaces at these 23 

7-22
 

facilities, 15 are reverberatory furnaces, 24 are blast furnaces, 5 are rotary furnaces, and 1 is an 

electric furnace. The 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, 1994c). 

Furnace charge materials consist of lead-bearing raw materials, lead-bearing slag and 

drosses, fluxing agents (blast and rotary furnaces only), and coke. Scrap motor vehicle lead-acid 

batteries represent about 90% of the lead-bearing raw materials at a typical lead smelter. 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. Organic emissions from 

co-located blast and reverberatory furnaces are more similar to the emissions of a reverberatory 

furnace than to those of a blast furnace (U.S. EPA, 1994c). 

In 1987, the lead smelting industry consisted of 24 facilities producing 0.72 million 

metric tons of lead (U.S. EPA, 1994c). In 1995, there were 23 companies producing 0.97 million 

metric tons (USGS, 1997e), and in 2000 there were 27 secondary lead smelters in operation in 

the United States producing 1.02 million metric tons (USGS, 2002). In 1995, the total annual 

production capacity of the 23 companies that made up the U.S. lead smelting industry was 1.36 

million metric tons. Blast furnaces not co-located with reverberatory furnaces accounted for 21% 

of capacity (0.28 million metric tons). Reverberatory furnaces and blast and electric furnaces colocated with reverberatory furnaces accounted for 74% of capacity (1.01 million metric tons). 

Rotary furnaces accounted for the remaining 5% of capacity (0.07 million metric tons) (U.S. 

EPA, 1994c). 

Actual production volume statistics by furnace type were not available. However, if it is 

assumed that the total actual production volume of the industry reflects 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 were 0.15, 0.53, and 0.04 million metric tons, respectively, in 1987; 0.2, 0.72, and 0.05 

million metric tons, respectively, in 1995; and 0.29, 1, and 0.07 million metric tons, respectively, 

in 2000. 

A report commissioned by Environment Canada (Charles E. Napier Company, Ltd., 

2000) reviewed published literature and other information on the dioxin/furan formation 

mechanisms; dioxin/furan emissions; emission control technology, including cost; and 

dioxin/furan published emission standards pertinent to steel production processes of plants 

in Canada. The report included four facilities identified as primary lead smelters. CDD/CDF 

emission concentrations were reported to range from less than 100 to less than 1,000 pg ITEQ/dscm. 

7-23
 
CDD/CDF emission factors were estimated for secondary lead smelters using the results 

of emission tests performed by EPA at three smelters (a blast furnace [U.S. EPA, 1995d], a colocated blast/reverberatory furnace [U.S. EPA, 1992c], and a rotary kiln furnace [U.S. EPA, 

1995e]). The air pollution control systems at the three tested facilities consisted of both FFs and 

scrubbers. Congener-specific measurements were made at both APCD exit points at each 

facility. Table 7-6 presents the congener and congener group emission factors for the FF and the 

scrubber for each site. Figure 7-3 presents the corresponding profiles for the FF emissions from 

the tested blast furnace and reverberatory furnace. For the facilities in operation in 1995, all 23 

smelters employed FFs, with only 9 employing scrubber technology. Facilities with scrubbers 

accounted for 14% of the blast furnace (not co-located) production capacity, 52% of the 

reverberatory and co-located furnace production capacity, and 57% of the rotary furnace 

production capacity. TEQ emission factors (ng TEQ/kg lead produced when nondetect values 

are set equal to zero) from the reported data for each of the three furnace configurations, 

presented as a range reflecting the presence or absence of a scrubber, are 

•	 Blast furnace: 0.64 to 8.81 ng TEQDF-WHO98/kg (0.63 to 8.31 ng I-TEQDF/kg) 

•	 Reverberatory/co-located furnace: 0.05 to 0.42 ng TEQDF-WHO98/kg (0.05 to 0.41 ng 

I-TEQDF/kg) 

•	 Rotary furnace: 0.24 to 0.66 ng TEQDF-WHO98/kg (0.24 to 0.66 ng I-TEQDF/kg) 

If these ranges of emission rates are assumed to be representative of those at nontested 

facilities with the same basic furnace configuration, with and without scrubbers, then combining 

these emission rates with the estimated production volumes derived above and the percentage of 

each configuration type that have scrubbers yields the estimated air emissions shown in Table 7-7 

for reference years 1987, 1995, and 2000. 

A medium confidence rating is assigned to the emission factors because stack test data 

were available for 3 of the 27 smelters operating in the United States (of which only 16 were in 

operation as of December 1993), and the stack test data used represent the three major furnace 

configurations. The activity level 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. 

Therefore, a medium confidence rating is assigned to the emission estimates. 

7-24
 

Table 7-6. CDD/CDF emission factors (ng/kg lead produced) for secondary 

lead smeltersa 

Congener/congener 

group 

Blast furnaceb 

Blast/reverberatory 

furnacec Rotary kilnd 

Before 

scrubber/ 

FF 

After 

scrubber/ 

FF 

Before 

scrubber/ 

FF 

After 

scrubber/ 

FF 

Before 

scrubber/ 

FF 

After 

scrubber/ 

FF 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

2.11 

0.99 

0.43 

0.99 

1.55 

2.06 

1.4 

0.25 

0.03 

0 

0.03 

0.03 

0.08 

0.39 

0 

0 

0 

0 

0 

0.1 

0.57 

0 

0 

0 

0 

0 

0.06 

0.55 

0.1 

0.01 

0 

0 

0 

0 

0.24 

0.24 

0 

0 

0 

0 

0.22 

2.41 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

8.73 

3.88 

6.65 

5.83 

1.67 

0.11 

2.06 

2.34 

0.63 

1.39 

0.93 

0.43 

0.36 

0.37 

0.11 

0 

0.11 

0.19 

0.06 

0.18 

1.46 

0.24 

0.31 

0.63 

0.19 

0 

0.15 

0.48 

0 

0.29 

0.49 

0.02 

0 

0 

0 

0 

0 

0 

0 

0 

0.40 

0.14 

0.14 

0.11 

0.02 

0.04 

0 

0.03 

0 

0 

1.2 

0.4 

0.46 

0.27 

0.1 

0.13 

0 

0.13 

0 

0 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF

 (nondetect set to zero) 

Total TEQDF-WHO98

 (nondetect set to zero) 

9.53 

33.29 

8.31 

8.81 

0.81 

2.74 

0.63 

0.64 

0.67 

3.75 

0.42 

0.42 

0.61 

0.51 

0.05 

0.05 

0.35 

0.88 

0.24 

0.24 

2.87 

2.69 

0.67 

0.66 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

74.33 

39.29 

20.05 

4.2 

1.39 

145.71 

69.59 

19.73 

4.74 

1.39 

7.39 

1.73 

0.81 

9.72 

0.18 

17.34 

3.45 

1.02 

0.11 

0.18 

0.97 

0.15 

0.14 

0.09 

0.57 

8.21 

3.07 

1.14 

0.72 

0.29 

1.58 

0.16 

0.02 

0.09 

0.55 

4.71 

0.36 

0.19 

0.01 

0.00 

3.4 

0.29 

0.1 

0.01 

0.24 

10.82 

1.69 

0.15 

0.05 

0 

7.9 

0.27 

0.23 

0.29 

2.41 

28.57 

5.04 

0.73 

0.14 

0 

Total CDD/CDF 

(nondetect set to zero) 

380.42 41.93 15.35 7.67 16.75 45.58 

a

Except where noted, emission factors were calculated assuming nondetect values were zero. b

Source: U.S. EPA (1995d). c

Source: U.S. EPA (1992c). d

Source: U.S. EPA (1995e). 

FF = Fabric filter 

7-25
 

Figure 7-3. Congener and congener group profiles for air emissions from 

secondary lead smelters. Profiles are for emissions from fabric filters; nondetect 

values set equal to zero. 

Sources: U.S. EPA (1992c, 1995d, e). 

7-26 
Table 7-7. Estimated annual TEQ emissions (g TEQ)

a 

Configuration 

Ref. year 1987 Ref. year 1995 Ref. year 2000 

TEQDF ­

WHO98 I-TEQDF 

TEQDF ­

WHO98 I-TEQDF 

TEQDF ­

WHO98 I-TEQDF 

Blast furnaces w/scrubbers 0.013 0.013 0.018 0.018 0.026 0.026 

Blast furnaces w/o 

scrubbers 

1.136 1.072 1.515 1.429 2.197 2.073 

Reverberatory furnaces 

w/scrubbers 

0.014 0.014 0.019 0.019 0.026 0.026 

Reverberatory furnaces w/o 

scrubbers 

0.106 0.104 0.145 0.142 0.202 0.197 

Rotary furnaces 

w/scrubbers 

0.015 0.015 0.019 0.019 0.026 0.026 

Rotary furnaces w/o 

scrubbers 

0.004 0.004 0.005 0.005 0.007 0.007 

TOTAL 1.288 1.222 1.721 1.632 2.484 2.355 

a

Calculated using emission factors based on nondetect values set equal to zero. 

7.3. PRIMARY FERROUS METAL SMELTING/REFINING 

Iron is manufactured from its ores (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. The primary production of iron and steel involves two operations identified by 

European researchers as potential emission sources of CDDs/CDFs: iron ore sinter production 

and coke production. Each of these potential sources is discussed in the following subsections. 

7.3.1. Sinter Production 

At some iron manufacturing facilities, iron ores and waste iron-bearing materials undergo 

sintering to convert the materials to usable feed for the blast furnace. In the sintering process, 

iron ore fines and waste materials are mixed with coke fines, and the mixture is placed on a grate 

that is then heated to a temperature of 1,000 to 1,400°C. The heat generated during combustion 

sinters the small particles. Iron-bearing dusts and slags from processes in the steel plant are the 

types of iron-bearing waste materials used as a feed mix for the sintering plant (Knepper, 1981; 

Capes, 1983; U.S. EPA, 1995b). 

7-27
 
Several European investigators have reported that iron ore sintering plants are major 

sources of airborne emissions of CDDs/CDFs (Rappe, 1992a; Lexen et al., 1993; Lahl, 1993, 

1994). Lahl reported that the practice of recycling dusts and scraps from other processes in the 

steel plant for use in the sintering plant introduces traces of chlorine and organic compounds that 

generate the CDDs/CDFs found in these plants. 

Organic compounds that are potential precursors to CDD/CDF formation come primarily 

from oil, which is found in mill scale, as well as from some blast furnace sludges that are used as 

part of the sinter feed mixture. Most U.S. plants limit the amount of oil because it increases 

emissions of volatile organic compounds and may create a fire hazard. In addition, plants with 

FFs must limit the oil content because the oil tends to blind the FFs. Typical oil content of the 

feed at U.S. sintering plants ranges from 0.1 to 0.75% (Calcagni et al., 1998). 

Sintering plants in Sweden have been reported to emit up to 3 ng I-TEQDF/Nm3 stack gas, 

or 2 to 4 g I-TEQDF/yr (Rappe, 1992a; Lexen et al., 1993). Bremmer et al. (1994) reported the 

results of stack testing at three iron ore sintering plants in the Netherlands. One facility equipped 

with wet scrubbers (WSs) had an emission factor of 1.8 ng I-TEQDF/dscm (at 11% oxygen). The 

other two facilities, both equipped with cyclones, had emission factors of 6.3 and 9.6 ng ITEQDF/dscm (at 7% oxygen). Lahl (1993, 1994) reported stack emissions for sintering plants in 

Germany (after passage through mechanical filters and ESPs) ranging from 3 to 10 ng ITEQDF/Nm3

. A compilation of emission measurements by the German Federal Environmental 

Agency indicated stack emission concentrations ranging from 1.2 to 60.6 ng I-TEQDF/m3 (at 7% 

oxygen); the majority of emissions in 1996 were around 3 ng I-TEQDF/m3 (Umweltbundesamt, 

1996). 

The report commissioned by Environment Canada in 2000 to review steel production 

processes in Canadian plants (Charles E. Napier Company, Ltd., 2000) included information on 

emissions from iron sintering. For iron sintering, the CDD/CDF emissions from one facility, the 

Stelco Hilton Works sintering plant, were assumed to be representative of the 1998 sinter 

production. The average emission rate was 19.9 ng I-TEQ/day. Applying a production rate of 

1,143 metric tons/day yields a mass emission factor of 17.4 ng I-TEQ/kg of sinter. 

EPA conducted tests at two of the nine U.S. sintering plants operating in 1997 in order to 

quantify emissions of CDDs/CDFs (Calcagni et al., 1998). In choosing representative plants for 

testing, EPA considered a variety of issues, including the types and quantities of feed materials, 

the types of emission controls, and the oil content of the sinter feed. EPA decided to test a plant 

with an FF and a plant with a venturi (or wet) scrubber. FFs and WSs are the principal APCDs 

used to control emissions from the sintering plant windbox. Four plants used an FF and five 

plants used a WS. The types of feed materials and oil content at the two selected plants were 

7-28
 

determined to be representative of other plants in the industry. Sampling was performed over 3 

days (4 hr/day) at each plant. 

The average CDD/CDF TEQ concentrations measured in the stack emissions were 0.19 

ng I-TEQDF/Nm3 and 0.81 ng I-TEQDF/Nm3 for the WS and the FF, respectively. The 

corresponding TEQ emission factors are 0.62 ng TEQDF-WHO98/kg (0.55 ng I-TEQDF/kg) sinter 

and 4.61 ng TEQDF-WHO98/kg (4.14 ng I-TEQDF/kg) sinter, respectively, for WSs and FFs. 

These emission factors are assigned a high rating because they are based on EPA testing at two 

facilities considered by EPA to be representative of both current and 1995 standard industry 

practices. 

Congener-specific emission factors for these two facilities are presented in Table 7-8. 

Figure 7-4 presents the congener profiles for these facilities. Although concentrations were 

higher from the FF than from the WS, both concentrations were low relative to what had been 

reported from testing at German, Dutch, Swedish, and Canadian sintering plants. This disparity 

may be due to differences in the operation or APCDs of U.S. sintering plants and the tested 

European plants. 

Most of the U.S. integrated iron and steel plants, including those with sintering plants, 

have eliminated the purchase and use of chlorinated organics in their facilities, and their rolling 

mill oils (lubricants and hydraulic fluids) do not contain chlorinated compounds. In addition, 

routine analyses of waste materials going to the sintering plant have not detected any chlorinated 

solvents. 

Finally, none of the U.S. plants use an ESP to control emissions from the sinter windbox 

(Calcagni et al., 1998). 

In 1996 (data were not readily available for 1995), 11 sintering plants were operating in 

the United States, with a total annual production capacity of about 17.6 million metric tons 

(Metal Producing, 1996). Since the 1980s, the size of this industry has decreased dramatically. 

In 1982, 33 facilities were in operation, with a combined total capacity of 48.3 million metric 

tons (U.S. EPA, 1982a). The nine U.S. sintering plants operating in 1995 had a combined 

capacity of 15.6 million metric tons (Calcagni et al., 1998). In 1987, sinter consumption by iron 

and steel plants was 14.5 million metric tons (AISI, 1990); in 1995, consumption was 12.4 

million metric tons (Fenton, 1996), or approximately 70% of production capacity, assuming that 

production capacity in 1995 was the same as in 1996. These activity level estimates are assigned 

a confidence rating of medium. 

Based on the production capabilities shown in Table 7-9, 59% of 1998 sinter production 

capacity was at facilities with WSs and 41% was at facilities with FFs. If it is assumed that these 

proportions of APCD-to-production capacity existed in 1995 and that actual production in 1995 

was equal to sinter consumption at iron and steel plants (12.4 million metric tons), then estimated 

7-29
 

Table 7-8. CDD/CDF emission factors (ng/kg sinter) for sintering plants 

Congener/congener 

group 

Wet scrubber Fabric filter 

Nondetect set 

to zero 

Nondetect set to 1⁄2 

detection limit 

Nondetect set 

to zero 

Nondetect set to 1⁄2 

detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.049 

0.138 

0.03 

0.612 

0.288 

0.696 

0.496 

0.049 

0.138 

0.03 

0.612 

0.288 

0.696 

0.496 

0.406 

0.937 

0.135 

1.469 

0.609 

0.698 

0.695 

0.406 

0.937 

0.135 

1.469 

0.609 

0.698 

0.695 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.602 

0.343 

0.349 

0.421 

0.164 

0.011 

0.142 

0.247 

0.036 

0.103 

0.602 

0.343 

0.349 

0.421 

0.164 

0.014 

0.142 

0.247 

0.036 

0.103 

10.232 

3.518 

3.228 

1.382 

0.495 

0.029 

0.285 

0.316 

0 

0.05 

10.232 

3.518 

3.228 

1.382 

0.495 

0.057 

0.285 

0.316 

0.115 

0.192 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

2.309 

2.418 

0.55 

0.62 

2.309 

2.421 

0.55 

0.62 

4.949 

19.820 

4.14 

4.61 

4.949 

19.82 

4.14 

4.61 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

NR 

NR 

NR 

NR 

0.496 

NR 

NR 

NR 

NR 

0.103 

NR 

NR 

NR 

NR 

0.496 

NR 

NR 

NR 

NR 

0.103 

NR 

NR 

NR 

NR 

0.695 

NR 

NR 

NR 

NR 

0.050 

NR 

NR 

NR 

NR 

0.695 

NR 

NR 

NR 

NR 

0.192 

Total 2,3,7,8 CDD/CDFa 4.73 4.73 24.77 24.77 a

The listed values for total CDD/CDF include only the 17 toxic congeners. 

Source: Calcagni et al. (1998). 

7-30
 

Figure 7-4. Congener profiles for air emissions from U.S. iron ore sintering 

plants. 

Source: Calcagni et al. (1998). 

TEQ emissions from WS-equipped facilities were 4.5 g TEQDF-WHO98 (4 g I-TEQDF) and 

emissions from FF-equipped facilities were 23.4 g TEQDF-WHO98 (21 g I-TEQDF), for a total of 

27.9 g TEQDF-WHO98 (25.1 g I-TEQDF). These emission estimates are assigned an overall 

medium confidence rating on the basis of the medium rating for the activity level estimates. 

If these same assumptions are applied to the 1987 sinter consumption rate of 14.5 million 

metric tons, then estimated TEQ emissions from WS-equipped facilities were 5.3 g TEQDFWHO98 (4.7 g I-TEQDF) and emissions from FF-equipped facilities were 27.4 g TEQDF-WHO98 

(24.6 g I-TEQDF), for a total of 32.7 g TEQDF-WHO98 (29.3 g I-TEQDF). These emission 

estimates are less certain than the estimates for 1995 because of uncertainties concerning actual 

APCDs in place in 1987 and the content of waste feed (i.e., oil content and presence of 

7-31
 
Table 7-9. Operating parameters for U.S. iron ore sintering plants 

Company Location 

1998 capacity 

(1,000 metric 

tons/yr) 

Current air 

pollution control 

device 

AK Steel Middletown, OH 907 Wet scrubber 

AK Steela Ashland, KY 816a NA 

Bethlehem Steel Burns Harbor, IN 2,676 Wet scrubber 

Bethlehem Steel Sparrows Point, MD 3,856 Wet scrubber 

Geneva Steel Provo, UT 816 Fabric filter 

Inland Steel East Chicago, IN 1,089 Fabric filter 

LTV Steel East Chicago, IN 1,270 Wet scrubber 

U.S. Steel Gary, IN 3,992 Fabric filter 

Weirton Steela Weirton, WV 1,179a NA 

Wheeling-Pittsburgh Steel East Steubenville, WV 519 Wet scrubber 

WCI Steel Warren, OH 477 Fabric filter 

TOTAL 17,597b 

a

Not in operation during 1998 (Calcagni et al., 1998).
 
 b

When the Ashland, KY, and Weirton, WV, facilities are excluded, total 1998 capacity was 15,600,000 metric tons.
 


NA = Not available 

Sources: Metal Producing (1991, 1996); Calcagni et al. (1998). 

chlorinated organics in the oil) at that time. Consequently, a low confidence rating is assigned to 

the emission factor and the emissions estimate. 

In 2000, a total of 10,600 million metric tons of sinter were consumed in blast furnaces 

(Fenton, 2001). This activity level has a high confidence rating because it is based on a 

comprehensive survey. Assuming the same proportions for facilities with WSs and facilities 

with FFs as in 1995 and 1987, then estimated TEQ emissions from WS-equipped facilities were 

3.9 g TEQDF-WHO98 (3.4 g I-TEQDF) and emissions from FF-equipped facilities were 23.7 g 

TEQDF-WHO98 (21.3 g I-TEQDF), for a total of 27.6 TEQDF-WHO98 (24.4 g I-TEQDF) for 2000. 

This emissions estimate is assigned a high confidence rating on the basis of the high ratings 

given to the activity level and emission factor for reference year 2000. 

7.3.2. Coke Production 

Coke is the principal fuel used in the manufacture of iron and steel. It is the solid 

carbonaceous material produced by the destructive distillation of coal in high-temperature ovens. 

7-32
 

No testing of CDD/CDF emissions from U.S. coke facilities has been reported. However, at a 

facility in the Netherlands, Bremmer et al. (1994) measured a CDD/CDF emission rate to air 

during the water quenching of hot coke of 0.23 ng I-TEQDF/kg of coal consumed. Bremmer et al. 

estimated minimal CDD/CDF air emissions (0.002 ng I-TEQDF/kg of coal) for flue gases 

generated during the charging and emptying of the coke ovens. 

The report commissioned by Environment Canada in 2000 to review steel production 

processes in Canadian plants (Charles E. Napier Company, Ltd., 2000) also provided information 

on emissions from coke ovens. Mean emission factors (ng I-TEQ/kg) of the four Canadian coke 

oven facilities (as indicated in Table 7-10) were 0.3 ng I-TEQ/kg coke produced. 

Although there are no testing data on which to base an estimate of CDD/CDF emissions 

in the United States, a preliminary estimate of potential TEQ annual emissions from U.S. coke 

plants can be made by combining the estimated consumption values of 33.5 million metric tons 

in 1987, 29.9 million metric tons in 1995, and 26.2 million metric tons in 2000 (EIA, 2002) with 

the emission factor reported by Bremmer et al. (1994) for a Dutch coke plant (0.23 ng ITEQDF/kg of coal consumed). These calculations yield annual emissions of 7.7, 6.9, and 6.03 g ITEQDF for 1987, 1995, and 2000, respectively. These estimates should be regarded as 

preliminary indications of possible emissions from this source category; further testing is needed 

to confirm the true magnitude of these emissions. 

7.4. SECONDARY FERROUS METAL SMELTING/REFINING 

Electric arc furnaces in Europe have been reported to be sources of CDD/CDF emissions; 

no testing has been reported at U.S. facilities. Electric arc furnaces are used to produce carbon 

and steel alloys, primarily from scrap material, using a batch process. The input material is 

typically 100% scrap. Scrap, alloying agents, and fluxing materials are loaded into the 

cylindrical, refractory-lined furnace, and then carbon electrodes are lowered into the mix. The 

current of the opposite-polarity electrodes generates heat through the scrap. Processing time of a 

batch ranges from about 1.5 to 5 hr to produce carbon steel and from 5 to 10 hr to produce alloy 

steel (U.S. EPA, 1995b). 

The melting of scrap ferrous material contaminated with metalworking fluids and plastics 

that contain chlorine provides the conditions conducive to formation of CDDs/CDFs. Tysklind 

et al. (1989) studied the formation and release of CDDs/CDFs at a pilot 10-ton electric furnace in 

Sweden. Scrap ferrous metal feedstocks containing varying amounts of chlorinated compounds 

(PVC plastics, cutting oils, or calcium chloride) were charged into the furnace under different 

operating conditions (continuous feed, batch feed into the open furnace, or batch feed through the 

furnace lid). During continuous charging operations, the highest emissions, 1.5 ng Nordic 

7-33
 

Table 7-10. CDD/CDF emission estimates for Canadian coke oven facilities, 

blast furnace facilities, and electric arc furnaces 

Company/facility Location 

Plant 

capacity 

(1,000 net 

tonnes/yr) 

Estimated 

production 

(1,000 net 

tonnes) 

Estimated 

CDD/CDF 

emissions 

(g I-TEQ) 

Estimated 

CDD/CDF 

emission 

factor 

(ng ITEQ/kg) 

Coke oven facilities 

Algoma Steel Inc. Sault Ste. Marie, Ontario 1,021 979 0.29 0.296 

Dofasco Inc. Hamilton, Ontario 1,656 1,588 0.48 0.302 

Stelco Inc., Lake Erie Steel Nanticoke, Ontario 563 540 0.16 0.296 

Stelco Inc., Hilton Works Hamilton, Ontario 1,035 993 0.30 0.302 

TOTAL 4,275 4,100 1.23 

Blast furnace facilities 

Algoma Steel Inc. Sault Ste. Marie, Ontario 2,270 2,177 <0.01 NA 

Dofasco Inc. Hamilton, Ontario 2,725 2,613 <0.01 NA 

Stelco Inc., Lake Erie Steel Nanticoke, Ontario 1,680 1,611 <0.01 NA 

Stelco Inc., Hilton Works Hamilton, Ontario 2,720 2,608 <0.01 NA 

TOTAL 9,395 9,009 <0.01 

Electric arc furnaces 

AltaSteel Ltd. Edmonton, Alberta 295 256 0.67 2.62 

Atlas Specialty Steels Welland, Ontario 218 189 0.49 2.59 

Atlas Stainless Steels Tracy, Quebec 118 103 0.27 2.62 

Co-Steel Lasco Whitby, Ontario 907 788 0.79 1.00 

Dofasco Inc. Hamilton, Ontario 1,225 1,065 0.50 0.469 

Gerdau MRM Steel Inc. Cambridge, Ontario 290 252 0.66 2.62 

Gerdan MRM Steel Inc. Selkirk, Manitoba 281 244 0.63 2.58 

IPSCO Inc. Regina, Saskatchewan 907 788 1.13 1.43 

Ispat Sidbec Inc. Contrecoeur, Quebec 1,633 1,419 3.69 2.60 

Ivanco Rolling Mills Inc. L’Original, Ontario 408 355 0.92 2.59 

Slater Steels, Hamilton 

Specialty Bar Div. 

Hamilton, Ontario 363 315 0.82 2.60 

Stelco-McMaster Ltèe Contrecoeur, Quebec 499 434 1.13 2.60 

Sydney Corp. Sydney, Nova Scotia 454 395 0.40 1.01 

TOTAL 7,598 6,603 12.10 

NA = Not available 

Source: Charles E. Napier Company, Ltd. (2000). 

7-34 
TEQ/dry Nm3

 (after an FF), were observed with a feedstock consisting of scrap metal with PVC 

plastics (1.3 g chlorine/kg feedstock). This emission rate equates to 7.7 ng Nordic TEQ/kg of 

feedstock. 

The highest emissions during batch charging also occurred when the scrap metal with 

PVC plastic was combusted (0.3 ng Nordic TEQ/dry Nm3 [1.7 ng Nordic TEQ/kg] feedstock). 

Much lower emissions (0.1 ng Nordic TEQ/dry Nm3 [0.6 ng Nordic TEQ/kg] feedstock) were 

observed when scrap metal with cutting oils that contained chlorinated additives (0.4 g 

chlorine/kg feedstock) was melted. Although these cutting oil-related emissions were not 

significantly different from the emissions observed from the melting of no-chlorine scrap metal, 

relatively high levels of CDDs/CDFs (110 ng Nordic TEQ/dry Nm3) were detected in flue gases 

prior to the FF. 

The congener profiles of raw flue gas samples (prior to the APCD) showed that CDFs 

rather than CDDs were predominant in all three feedstock types. The congener profile from the 

test burn with PVC-containing feedstock showed a higher chlorinated congener content than was 

observed with the other feedstocks. 

Eduljee and Dyke (1996) used a range of 0.7 to 10 ng I-TEQDF/kg of scrap feed to 

estimate national emissions for the United Kingdom. This range was assumed to be 

representative of no-chlorine and high-chlorine operations. However, the study authors provided 

little information on the supporting emission test studies (i.e., tested facility operational 

materials, feed rates, congener-specific emission rates). 

Umweltbundesamt (1996) reported stack testing results for a variety of electric arc 

furnaces in Germany. Sufficient data were provided in the report to enable calculation of TEQ 

emission factors for six of the tested facilities. Two facilities had emission factors exceeding 1 

ng I-TEQDF/kg of scrap processed, and two facilities had emission factors of less than 0.1 ng I

TEQDF/kg of scrap. The mean emission factor was 1.15 ng I-TEQDF/kg of scrap. The TEQ 

concentrations in the stack gases at these facilities (corrected to 7% oxygen) ranged from less 

than 0.1 to 1.3 ng I-TEQDF/m3. 

The report commissioned by Environment Canada in 2000 to review steel production 

processes at Canadian plants (Charles E. Napier Company, Ltd., 2000) included information on 

emissions from iron sintering and provided information on emissions from electric arc furnaces, 

which were estimated on the basis of plant capacity and estimated production. Mean emission 

factors (ng I-TEQ/kg) of the 13 Canadian electric arc furnace facilities (as indicated in Table 

7-10) were 2.1 ng I-TEQDF/kg steel produced. 

In March 2000, Environment Canada reported on source testing to determine CDD/CDF 

emissions from a facility in Ontario (Cianciarelli, 2000). Sampling was conducted on the 

exhaust stack of the electric arc furnace of Dofasco Inc., and both concentrations and emission 

­

7-35
 

rates were provided (see Table 7-11). Total CDD/CDF concentrations were reported to be 51.15 

pg TEQ/m3, and the total emission rate was reported to be 0.47 ng TEQ/kg steel produced. In 

August 2000, the Emissions Research and Measurement Division of Environment Canada 

conducted source testing to determine CDDs/CDFs from the electric arc furnace of another 

facility, Gerdau Courtice Steel Inc. (Cianciarelli, 2001). These results (presented in Table 7-11) 

are being used to support the Canadian dioxin/furan inventory for electric arc furnaces. The total 

CDD/CDF concentrations were reported to be 125.5 pg TEQ/m3

, and the total emission rate was 

reported to be 1.1 ng TEQ/kg steel produced. 

In 1987, electric arc furnaces accounted for 38.1% of U.S. steel production, or 30.8 

billion kg of raw steel produced (Peters, 1988). In 1995, electric arc furnaces accounted for 

40.4% of U.S. steel production, or 38.4 of the total 95.2 million metric tons of raw steel produced 

(Fenton, 1996). In 2000, electric arc furnaces accounted for 46.2% of U.S. steel production, or 

49 of the 106 million metric tons of raw steel produced (USGS, 2002). 

No testing of CDD/CDF emissions from U.S. electric arc furnaces on which to base an 

estimate of national emissions has been reported. A preliminary estimate of potential TEQ 

annual emissions from U.S. electric arc furnaces can be made by combining the production 

estimate of steel and an average emission factor of 1.21 ng I-TEQDF/kg steel derived from the 

data reported in Umweltbundesamt (1996) and the three Environment Canada reports (Charles E. 

Napier Company, Ltd., 2000; Cianciarelli, 2000, 2001). This calculation yields an annual 

emissions estimate of 37.3 g I-TEQDF in 1987, 46.5 g I-TEQDF in 1995, and 59.3 g I-TEQDF in 

2000. These estimates should be regarded as preliminary indications of possible emissions from 

this source category; further testing is needed to confirm the true magnitude of these emissions. 

7.5. FERROUS FOUNDRIES 

Ferrous foundries produce high-strength iron and steel castings used in industrial 

machinery, pipes, and heavy transportation equipment. Iron and steel castings are solid solutions 

of iron, carbon, and various alloying materials. Castings are produced by injecting or pouring 

molten metal into cavities of a mold made of sand, metal, or ceramic material. Metallic raw 

materials are pig iron, iron and steel scrap, foundry returns, and metal turnings (U.S. EPA, 

1995b, 1997a). 

The melting process takes place primarily in cupola (or blast) furnaces and to a lesser 

extent in electric arc furnaces. About 70% of all iron castings are produced using cupolas, 

although steel foundries rely almost exclusively on electric arc furnaces or induction furnaces for 

melting. The cupola is typically a vertical, cylindrical steel shell with either a refractory-lined or 

a water-cooled inner wall. Charges are loaded at the top of the unit; the iron is melted as it flows 

7-36
 

Table 7-11. CDD/CDF emission concentrations and rates for Canadian 

electric arc furnaces 

Congener 

Mean facility concentration 

(pg TEQ/m3

) 

Mean facility emission rate 

(ng TEQ/tonne steel) 

Run 1 Run 2 Run 3 Avg. Run 1 Run 2 Run 3 Avg. 

Dofasco Inc. 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0 

1.96 

0 

0 

0.25 

0.11 

0.04 

1.99 

4.09 

0 

0.33 

0 

0 

0.01 

0 

6.44 

0 

1.13 

0.63 

0.05 

0 

0.66 

4.16 

0 

0.49 

0.29 

0.05 

0.02 

0 

20.5 

0 

0 

2.7 

1.2 

0.4 

17.2 

35.3 

0 

2.8 

0 

0 

0 

0 

60.5 

0 

10.6 

5.9 

0.5 

0 

5.7 

38.8 

0 

4.5 

2.9 

0.6 

0.1 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

9.96 

0.48 

5.88 

1.12 

0.51 

0 

0 

0.11 

0 

0.01 

37.11 

1.23 

14.9 

2.23 

0.81 

0.52 

0 

0.05 

0 

0 

29.45 

1.60 

22.98 

3.86 

2.02 

1.43 

0 

0.16 

0 

0 

25.51 

1.1 

14.59 

2.4 

1.11 

0.65 

0 

0.11 

0 

0 

104.2 

5 

61.6 

11.7 

5.3 

0 

0 

1.1 

0 

0.1 

320.9 

10.6 

128.8 

19.3 

7 

4.5 

0 

0.4 

0 

0 

276.5 

15 

215.7 

36.3 

19 

13.4 

0 

1.5 

0 

0 

233.9 

10.2 

135.4 

22.4 

10.4 

6 

0 

1 

0 

0 

TOTAL 20.43 63.27 69.75 51.14 213.8 546.8 654.9 471.9 

Gerdau Courtice Steel Inc. 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

6.3 

8.3 

0.6 

1 

0.8 

0.2 

0 

2.7 

4.7 

0.4 

0.7 

0.5 

0.1 

0 

2.6 

3.1 

0.2 

0.4 

0.3 

0.1 

0 

3.9 

5.4 

0.4 

0.7 

0.5 

0.1 

0 

57 

75 

5 

9 

8 

1 

0 

21 

37 

3 

6 

4 

1 

0 

22 

27 

2 

3 

3 

1 

0 

33.3 

46.3 

3.3 

6 

5 

1 

0 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

65 

5.5 

95.5 

12.5 

6.9 

5.8 

0.5 

0.7 

0.1 

0 

29.4 

2.9 

46.2 

7.9 

4.6 

4.3 

0.5 

0.5 

0.1 

0 

18 

1.7 

26 

4.3 

2.4 

1.7 

0.2 

0.2 

0 

0 

37.5 

3.4 

55.9 

8.2 

4.6 

3.9 

0.4 

0.5 

0.1 

0 

588 

50 

864 

113 

62 

52 

5 

6 

1 

0 

232 

23 

364 

63 

36 

34 

4 

4 

1 

0 

154 

15 

222 

37 

20 

14 

2 

2 

0 

0 

324.7 

29.3 

483.3 

71 

39.3 

33.3 

3.7 

4 

0.7 

0 

TOTAL 209.7 105.5 61.2 125.5 1,896 833 524 1,084.2 

Source: Cianciarelli (2000). 

7-37
 

down the cupola and is removed at the bottom. Electric induction furnaces are batch-type 

furnaces in which the charge is melted by a fluctuating electromagnetic charge produced by 

electrical coils surrounding the unit (U.S. EPA, 1995b, 1997a). 

Iron and steel foundries, particularly those using electric arc furnaces, are highly 

dependent on iron and steel scrap. Of the estimated 72 million metric tons of iron and steel scrap 

consumed by the iron and steel industry in 1995, 25% (18 million metric tons) were used by 

ferrous foundries. The other 75% were used by primary ferrous metal smelters (principally those 

using electric arc furnaces) (USGS, 1997f). In 2000, 20% (12.4 million metric tons) were used 

by ferrous foundries; the remaining 80% were used by primary ferrous smelters (USGS, 2000). 

In 2000, there were approximately 1,100 ferrous foundries in the United States producing 

1.3 million metric tons of steel castings and 10 metric tons of iron castings. Thus, foundries face 

the same potential for CDD/CDF emissions as do electric arc furnaces because of their use of 

scrap that contains chlorinated solvents, plastics, and cutting oils (see Section 7.4) The potential 

for formation and release of CDDs/CDFs during the casting process is not known. 

In 1993, emissions testing was conducted at a U.S. ferrous foundry (CARB, 1993, as 

reported in U.S. EPA, 1997a). The tested facility consisted of a batch-operated, coke-fired 

cupola furnace charged with pig iron, scrap iron, scrap steel, coke, and limestone. Emission 

control devices operating during the testing were an oil-fired afterburner and an FF. The 

congener and congener group emission factors derived from the testing are presented in Table 

7-12. The calculated TEQ emission factor for this set of tests is 0.42 ng TEQDF-WHO98 (0.37 ng 

I-TEQDF/kg) of metal charged to the furnace. 

Umweltbundesamt (1996) reported stack testing results for a variety of ferrous foundries 

in Germany. Sufficient data were provided to enable calculation of TEQ emission factors for 

eight of the tested facilities. Three facilities had emission factors exceeding 1 ng I-TEQDF/kg of 

metal charge, and four facilities had emission factors less than 0.1 ng I-TEQDF/kg of metal 

charge. The emission factors span more than four orders of magnitude. The mean emission 

factor was 1.26 ng I-TEQDF/kg of metal feed. 

In 1997, EPA conducted testing for emissions of dioxins at two ferrous foundries (U.S. 

EPA, 1999c, f). One study was conducted on the cupola’s WS; the second study was performed 

on the cupola’s FF. A summary of the results is presented in Table 7-13. The emission factor 

developed from these tests is 2.05 ng I-TEQ/kg of metal processed. 

Because of the wide range of emissions for the tested German foundries reported in 

Umweltbundesamt (1996), the confidence in the degree to which the three tested U.S. facilities 

represent the mean emission factor for the approximately 1,100 U.S. foundries is considered very 

low. Therefore, the limited data available were judged inadequate for developing national 

emission estimates that could be included in the national inventory. However, a preliminary 

7-38
 

Table 7-12. CDD/CDF emission factors for a U.S. ferrous foundry 

Congener/congener group 

Mean facility emission factor 

(ng/kg scrap feed) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.033 

0.086 

NR 

0.051 

NR 

0.093 

NR 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.52 

0.305 

0.35 

0.19 

0.17 

NR 

0.101 

0.193 

NR 

0.059 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF (for reported congeners) 

Total TEQDF-WHO98 

0.263 

1.888 

0.372 

0.415 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

3.96 

1.76 

0.55 

0.19 

NR 

25.8 

850 

1.74 

0.24 

0.06 

Total CDD/CDF (not including OCDD) 884.3 

NR = Not reported 

Source: CARB (1993), as reported in U.S. EPA (1997a). 

estimate of potential TEQ annual emissions from U.S. ferrous foundries can be made by 

combining the mean emission factor (1.23 ng I-TEQDF/kg of metal feed) derived from the data 

reported in Umweltbundesamt (1996), CARB (1993), and U.S. EPA (1997a) with an estimated 

activity level for U.S. foundries. 

7-39
 
Table 7-13. Congener-specific profile for ferrous foundries 

Congener 

Mean emission factor (2 facilities) 

(ng I-TEQ/kg) 

Nondetect set to zero Nondetect set to 1⁄2 detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.11 

0.15 

0.012 

0.023 

0.028 

0.0033 

0.16 

0.11 

0.15 

0.012 

0.023 

0.028 

0.0033 

0.16 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.084 

1.08 

0.21 

0.1 

0.0079 

0.075 

0.0082 

0.0014 

0.00009 

0.00007 

0.084 

1.08 

0.21 

0.1 

0.0079 

0.075 

0.0082 

0.0014 

0.00009 

0.00007 

Total I-TEQ 2.05 2.05 

Sources: U.S. EPA (1999c, f). 

In 1987, U.S. shipments from ferrous foundries were 9.19 million metric tons, of which 

about 90% were iron castings and 10% were steel castings (Houck, 1991). In 1995, U.S. 

shipments from the approximately 1,000 U.S. ferrous foundries were 13.9 million metric tons, of 

which about 90% were iron castings and 10% were steel castings (Fenton, 1996). In 2000, U.S. 

shipments from the approximate 1,100 U.S. ferrous foundries were 11.3 million metric tons, of 

which about 89% were iron castings and 11% were steel castings (USGS, 2001). Using the mean 

emission factors and these activity levels yields annual emission estimates of 11.3 g I-TEQDF, 

17.1 g I-TEQDF, and 13.9 g I-TEQDF for 1987, 1995, and 2000, respectively. These estimates 

should be regarded as preliminary indications of possible emissions from this source category; 

further testing is needed to confirm the true magnitude of these emissions. 

7.6. SCRAP ELECTRIC WIRE RECOVERY 

The objective of wire recovery is to reclaim the metal (copper, lead, silver, and gold) in 

the electric wire by removing the insulating material. The recovery facility then sells the 

reclaimed metal to a secondary metal smelter. Wire insulation commonly consists of a variety of 

7-40
 

plastics, asphalt-impregnated fabrics, or burlap. Chlorinated organics are used to preserve the 

cable casing in below-ground cables. The combustion of chlorinated organic compounds in the 

cable insulation, 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). 

Although, in the past, scrap electric wire was commonly recovered using thermal 

processing to burn off the insulating material, current recovery operations typically no longer 

involve thermal treatment, according to industry and trade association representatives. Instead, 

scrap electric wire is mechanically chopped into fine particles. The insulating material is then 

removed by mixing, followed by settling of the heavier metal (telephone conversations between 

T. Leighton, Versar, Inc., and R. Garino, Institute of Scrap Recycling Industries, March 2, 1993, 

and T. Leighton and J. Sullivan, Triple F. Dynamics, March 8, 1993). 

EPA measured dioxin-like compounds emitted to the air from a scrap wire reclamation 

incinerator during its 1986 National Dioxin Study of combustion sources (U.S. EPA, 1987a). 

EPA determined that the tested facility was 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 operated in a batch mode, with the combustion cycles for each batch of scrap 

feed lasting between 1 and 3 hr. Natural gas was used to incinerate the material. Although most 

of the wire had a tar-based insulation, PVC-coated wire was also fed to the incinerator. 

Temperatures during combustion in the primary chamber furnace were about 570°C. The tested 

facility was equipped with a high-temperature, natural gas-fired afterburner (980 to 1090°C). 

Emission factors estimated for this facility are presented in Table 7-14. The estimated TEQ 

emission factor (based only on 2,3,7,8-TCDD, 2,3,7,8-TCDF, OCDD, and OCDF) is 15.8 ng 

TEQDF-WHO98 (16.9 ng I-TEQDF/kg) of scrap feed. Figure 7-5 presents a congener group profile 

based on these emission factors. 

Bremmer et al. (1994) reported emission factors for three facilities in the Netherlands that 

subsequently ceased operations. Emission rates at a facility burning underground cables and 

cables containing PVC ranged from 3.7 ng I-TEQDF/kg to 14 ng I-TEQDF/kg. The emission rates 

at a second facility ranged from 21 ng I-TEQDF/kg of scrap (when burning copper core coated 

with greasy paper) to 2,280 ng I-TEQDF/kg of scrap (when burning lead cable). The third facility, 

which burned motors, was reported to have an emission rate of 3,300 ng I-TEQDF/kg of scrap. 

On the basis of these measurements, Bremmer et al. used emission rates of 40 ng I-TEQDF/kg of 

scrap and 3,300 ng I-TEQDF/kg of scrap for estimating national emissions in the Netherlands for 

facilities burning wires and cables and those burning motors, respectively. 

Although limited emissions testing has been conducted at one U.S. facility, the activity 

level for this industry sector in reference years 1987, 1995, and 2000 is unknown; therefore, an 

7-41
 

Table 7-14. CDD/CDF emission factors for a scrap wire incinerator 

Congener/congener group Mean facility emission factora (ng/kg scrap feed) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.374 

NR 

NR 

NR 

NR 

NR 

1,000 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

2.67 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

807 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

NR 

NR 

16.9b 

15.8 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

1,000 

4.42 

13.7 

71.1 

347 

107 

97.4 

203 

623 

807 

Total CDD/CDF 3,274 a

No nondetect values were reported for 2,3,7,8-TCDD, 2,3,7,8-TCDF, or any congener group in the three test runs. b

Estimated on the basis of the measured data for 2,3,7,8-TCDD, 2,3,7,8-TCDF, OCDD, and OCDF and congener

 group emissions (i.e., for the penta-, hexa-, and hepta-CDDs and CDFs, it was assumed that the measured

 emission factor within a congener group was the sum of equal emission factors for all congeners in that group,

 including non-2,3,7,8-substituted congeners). 

NR = Not reported 

Source: U.S. EPA (1987a). 

7-42
 

Figure 7-5. Congener group profile for air emissions from a scrap wire 

incinerator. 

Source: U.S. EPA (1987a). 

estimate of national emissions cannot be made. It is uncertain how many facilities in the United 

States still combust scrap wire. Trade association and industry representatives state that U.S. 

scrap wire recovery facilities now burn only minimal quantities of scrap wire. However, an 

inventory of CDD/CDF sources in the San Francisco Bay area noted that two facilities thermally 

treated electric motors to recover electrical windings (BAAQMD, 1996). 

In addition to releases from regulated recovery facilities, CDD/CDF releases from smallscale burning of wire at unregulated facilities and open air sites have occurred; however, the 

current magnitude of these types of activity in the United States is not known. Harnly et al. 

(1995) analyzed soil/ash mixtures from three closed metal recovery facilities and from three 

closed sites using open burning for copper recovery near a California desert town. The geometric 

means of the total CDD/CDF concentrations at the facility sites and the open burning sites were 

86,000 and 48,500 ng/kg, respectively. The geometric mean TEQ concentrations were 2,900 and 

1,300 ng I-TEQDF/kg, respectively. A significantly higher geometric mean concentration (19,000 

ng I-TEQDF/kg) was found in fly ash located at two of the facility sites. The congener-specific 

7-43
 

and congener group results from this study are presented in Table 7-15. The results show that the 

four dominant congeners in the soil samples at both the facility and the open burning sites were 

OCDF, 1,2,3,4,6,7,8-HpCDF, 1,2,3,4,7,8-HxCDF, and 2,3,7,8-TCDF. A slightly different profile 

was observed in the fly ash samples, with 1,2,3,7,8-PeCDF and 1,2,3,4,7,8,9-HpCDF replacing 

OCDD and 2,3,7,8-TCDF as the dominant congeners. 

Table 7-15. CDD/CDF concentrations in fly ash and ash/soil at metal recovery 

sites 

Congener/congener 

group 

Metal recovery facilities Open burn sites 

Fly ash (2 sites) Ash/soil (3 sites) Ash/soil (3 sites) 

Geometric

 mean 

(:g/kg) 

Relative 

percent of 

total 

CDD/CDF 

Geometric

 mean 

(:g/kg) 

Relative 

percent of 

total 

CDD/CDF 

Geometric

 mean 

(:g/kg) 

Relative 

percent of 

total 

CDD/CDF 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

a 

400 

1,200 

2,300 

1,700 

12,000 

18,000 

0.1 

0.2 

0.5 

0.3 

2.4 

3.5

 a 

0.24 

0.25 

0.49 

1.3 

2.6 

7.2 

0.3 

0.3 

0.6 

1.5 

3.1 

8.5

 a 

0.24 

0.13 

0.33 

0.39 

1.2 

3.4 

0.5 

0.3 

0.7 

0.8 

2.5 

7 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

15,000 

35,000 

10,000 

46,000 

12,000 

5,000 

5,000 

71,000 

25,000 

100,000 

2.9 

6.9 

2 

9 

2.4 

1 

1 

13.9 

4.9 

19.6 

6.4 

2.9 

1.4 

5.9 

1.8 

0.92 

1.6 

12 

3 

14 

7.5 

3.4 

1.6 

6.9 

2.1 

1.1 

1.9 

14.1 

3.5 

16.5 

1.7 

0.58 

0.66 

2.7 

0.76 

0.66 

0.49 

4.3 

0.71 

6.6 

3.5 

1.2 

1.4 

5.6 

1.6 

1.4 

1 

8.9 

1.5 

13.6 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

a 

2,000 

4,000 

24,000 

18,000 

23,000 

110,000 

88,000 

110,000 

100,000 

a 

0.4 

0.8 

4.7 

3.5 

4.5 

21.6 

17.3 

21.6 

19.6 

a 

1.4 

2.7 

4.1 

7.2 

14 

12 

12 

17 

14 

a 

1.6 

3.2 

4.8 

8.5 

16.5 

14.1 

14.1 

20 

16.5 

a 

2.8 

0.98 

2 

3.4 

5.6 

7 

7.6 

7.4 

6.6 

a 

5.8 

2 

4.1 

7 

11.5 

14.4 

15.7 

15.3 

13.6 

Total I-TEQDF 

Total CDD/CDF 

16,968 

479,000 

3 

84.4 

1.3 

48.4 

a

Analytical method used had low sensitivity for TCDDs; results were not reported. 

Source: Harnly et al. (1995). 

7-44 
Van Wijnen et al. (1992) reported similar results for soil samples collected from 

unpermitted incineration sites of former scrap wire and cars in the Netherlands. Total CDD/CDF 

concentrations in the soil ranged from 60 to 98,000 ng/kg, with 9 of the 15 soil samples having 

levels above 1,000 ng/kg. Chen et al. (1986) reported finding high levels of CDDs/CDFs in 

residues from open air burning of wire in Taiwan, and Huang et al. (1992) reported elevated 

levels in soil near wire scrap recovery operations in Japan. Bremmer et al. (1994) estimated an 

emission rate to air of 500 ng I-TEQDF/kg of scrap for illegal, unregulated burning of cables in the 

Netherlands. 

7.7. DRUM AND BARREL RECLAMATION FURNACES 

Hutzinger and Fiedler (1991a) reported detecting CDDs/CDFs in stack gas emissions 

from drum and barrel reclamation facilities at levels ranging from 5 to 27 ng/m3

. EPA measured 

dioxin-like compounds in the stack gas emissions of a drum and barrel reclamation furnace as 

part of the National Dioxin Study (U.S. EPA, 1987a). 

Drum and barrel reclamation facilities operate a burning furnace to thermally clean used 

55-gallon steel drums of residues and coatings. The drums processed at these facilities come 

from a variety of sources in the petroleum and chemical industries. The thermally cleaned drums 

are then repaired, repainted, relined, and sold for reuse. The drum-burning process subjects the 

used drums to an elevated temperature in a tunnel furnace fired by auxiliary fuel for a sufficient 

time so that the paint, interior linings, and previous contents are burned or disintegrated. 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, residual 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 aircooled as they exit the furnace. Exhaust gases from the burning furnace are typically drawn 

through a breeching fan to a high-temperature afterburner. 

The afterburner at the facility tested by EPA operated at an average of 827°C during 

testing and achieved a 95% reduction in CDD/CDF emissions (U.S. EPA, 1987a). Emission 

factors estimated for this facility are presented in Table 7-16. On the basis of the measured 

congener and congener group emissions, the average TEQ emission factor is estimated to be 17.5 

ng TEQDF-WHO98/drum (16.5 ng I-TEQDF/drum). The congener group profile is presented in 

Figure 7-6. 

Approximately 2.8 to 6.4 million 55-gallon drums are reclaimed by incineration annually 

in the United States (telephone conversation between C. D. Ruiz, Versar, Inc., and P. Rankin, 

Association of Container Reconditioners, December 21, 1992). This estimate is based on the 

assumption that 23 to 26 incinerators are in operation; each incinerator, on average, handles 500 

7-45
 

Table 7-16. CDD/CDF emission factors for a drum and barrel reclamation 

facility 

Congener/congener group Mean facility emission factora (ng/drum) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

2.09 

NR 

NR 

NR 

NR 

NR 

37.5 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

36.5 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

NR 

22.4 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

NR 

NR 

16.5b 

17.5 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

50.29 

29.2 

32.2 

53.4 

37.5 

623 

253 

122 

82.2 

22.4 

Total CDD/CDF 1,305.2 a

No nondetect values were reported for 2,3,7,8-TCDD, 2,3,7,8-TCDF, or any congener group in the three test runs. b

Estimated on the basis of the measured data for 2,3,7,8-TCDD, 2,3,7,8-TCDF, OCDD, and OCDF and congener

 group emissions (i.e., for the penta-, hexa-, and hepta-CDDs and CDFs, it was assumed that the measured

 emission factor within a congener group was the sum of equal emission factors for all congeners in that group,

 including non-2,3,7,8-substituted congeners). 

NR = Not reported 

Source: U.S. EPA (1987a). 

7-46
 

Figure 7-6. Congener group profile for air emissions from a drum incinerator. 

Source: U.S. EPA (1987a). 

to 1,000 drums each day and, on average, each incinerator operates 5 days/wk, with 14 downtime 

days annually for maintenance activities. The weights of 55-gallon drums vary considerably; 

however, on average, a drum weighs 38 lb (or 17 kg); therefore, an estimated 48 to 109 million 

kg of drums are incinerated annually. For 1987 and 1995, EPA assumed that 4.6 million drums 

were burned each year (i.e., the midpoint of the range); applying the emission factors developed 

above, the estimated annual emission is 0.08 g TEQDF-WHO98 (0.08 g I-TEQDF). 

The Reusable Industrial Packaging Association estimated that approximately 35 million 

55-gallon barrels were reclaimed in 1997 (RIPA, 1997). Assuming the number of drums treated 

has remained constant through 2000, the estimate for 2000 would be 0.61 g TEQDF-WHO98 (0.58 

g I-TEQDF). 

A low confidence rating is assigned to the activity level estimates for all reference years 

because they are based on expert judgment rather than a published reference. A low confidence 

7-47
 

rating is also assigned to the emission factor because it was developed from stack tests conducted 

at just one U.S. drum and barrel furnace and thus may not represent average emissions from 

current operations in the United States. Based on these ratings, the emission estimates are 

assigned a low confidence rating. 

7.8. SOLID WASTE FROM PRIMARY/SECONDARY IRON/STEEL 

MILLS/FOUNDRIES 

Table 17 in Quab and Fermann (1997) contains summary data on the typical annual 

quantities and ranges of TEQ (Norwegian-TEQ [NTEQ] and I-TEQ) from various solid residuals 

from the metallurgical industries in Europe, but support information and specific congeners were 

not discussed. The summary data for annual TEQ generation are presented below (in grams) for 

informational purposes only and are not included in the inventory of dioxin releases presented in 

this report because they are disposed of in permanent landfills and are not considered an 

environmental release. 

• Grey iron foundries, FF dust and scrubber sludge: 0.817 NTEQ 

• Steel mill coke oven door leakage dust: 0.31 NTEQ 

• Steel mill coke oven door leakage dust: 0.04 I-TEQ 

• Pig iron tapping slag: 0.041 NTEQ 

• Basic oxygen furnace scrubber sludge: 1.53 NTEQ (range, 0.3–7.81) 

• Electric furnace FF dust: 3.1 I-TEQ (range of 0.4–2.4) 

• Electric furnace slag or FF dust: 19.2 NTEQ 

7-48
 

8. CHEMICAL MANUFACTURING AND PROCESSING SOURCES
 

8.1. BLEACHED CHEMICAL WOOD PULP AND PAPER MILLS
 

In March 1988, EPA and the U.S. pulp and paper industry jointly released the results of a 

screening study that provided the first comprehensive data on the formation and discharge of 

CDDs/CDFs from pulp and paper mills (U.S. EPA, 1988d). This early screening study of five 

bleached kraft mills (the Five Mill Study) confirmed that the pulp bleaching process was 

primarily responsible for the formation of CDDs/CDFs. The study results showed that 2,3,7,8­

TCDD was present in seven of nine bleached pulps, five of five wastewater treatment sludges, 

and three of five treated wastewater effluents. The study results also indicated that 2,3,7,8­

TCDD and 2,3,7,8-TCDF were the principal CDDs/CDFs formed. 

To provide EPA with more complete data on the release of these compounds, EPA and 

the pulp and paper industry jointly conducted a survey during 1988 of 104 pulp and paper mills 

in the United States to measure levels of 2,3,7,8-TCDD and 2,3,7,8-TCDF in effluent, sludge, 

and pulp. The survey 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 mills where chemically produced wood pulps were bleached with chlorine 

or chlorine derivatives. The final study report (U.S. EPA, 1990e) was released in July 1990. 

An initial phase of the 104 Mill Study involved the analysis of bleached pulp (10 

samples), wastewater sludge (9 samples), and wastewater effluent (9 samples) from eight kraft 

mills and one sulfite mill for all 2,3,7,8-substituted CDDs/CDFs. These analyses were conducted 

to test the conclusion drawn in the Five Mill Study that 2,3,7,8-TCDD and 2,3,7,8-TCDF were 

the principal CDDs/CDFs found in pulp, wastewater sludge, and wastewater effluent on a TEQ 

basis. Although at the time of the study there were no reference analytical methods for many of 

the 2,3,7,8-substituted CDDs/CDFs, the data obtained were considered valid by EPA for the 

purposes intended because of the identification and quantification criteria used, duplicate sample 

results, and limited matrix spike experiments. Table 8-1 presents a summary of the results 

obtained in terms of the median concentrations and the range of concentrations observed for each 

matrix (pulp, sludge, and effluent). Figures 8-1, 8-2, and 8-3 present congener profiles for each 

matrix (normalized to total CDD/CDF) using the median reported concentrations. 

After examination of the raw, mill-specific data, EPA concluded that the congener 

profiles were fairly consistent across matrices within mills and that 2,3,7,8-TCDD and 2,3,7,8­

TCDF accounted for the majority of TEQ in the samples (U.S. EPA, 1990e). Using the median 

concentrations and treating nondetect values as either zero or one-half the detection limit (DL), 

8-1
 

Table 8-1. CDD/CDF concentrations in pulp and paper mill bleached pulp, wastewater sludge, and 

effluent (circa 1988) 

8-2
 


Congener/congener 

group 

Bleached pulp Wastewater sludge Wast

Median 

(ng/kg) 

Range 

(ng/kg) 

No. of 

detects 

(10 samples) 

Median 

(ng/kg) 

Range 

(ng/kg) 

No. of 

detects 

(9 samples) 

Median 

(pg/L) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

6.4 

ND(0.3) 

ND(0.4) 

ND(0.5) 

ND(0.5) 

3.3 

46 

0.4 to 124 

ND(0.1) to 1.4 

ND(0.2) to 0.4 

ND(0.2) to 1.6 

ND(0.2) to 0.5 

2.3 to 8.4 

28 to 81 

10 

2 

1 

2 

1 

10 10 

63 

ND(2.5) 

ND(3.1) 

ND(3.2) 

ND(3.9) 

37 

698 

ND(6.3) to 180 

ND(1.4) to 28 

ND(1.5) to 40 

ND(1.7) to 95 

ND(1.7) to 80 

18 to 490 

263 to 1,780 

8 

1 

1 

1 

1 

9 9 

42 

ND(9.6) 

ND(12) 

ND(12) 

ND(12) 

170 

3,000 

N

ND(

ND(

ND(

ND(

1,

2,3,7,8-TCDF 1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

18 

ND(0.7) ND(0.2) ND(0.3) 

ND(0.3) ND(0.3) ND(0.3) ND(0.6) ND(0.6) 

2.2 

1.4 to 716 

ND(0.1) to 3.9 ND(0.1) to 4.7 

ND(0.2) to ND(0.6) 

ND(0.1) to ND(0.4) ND(0.1) to ND(0.4) ND(0.2) to ND(0.4) 

ND(0.1) to 0.8 

ND(0.1) to ND(2.1) 

ND(2.8) to 4.3 

10 

4 3 0 

0 0 0 3 0 8 

233 

6.2 4.7 

ND(2.5) 

ND(1.4) ND(1.7) ND(1.7) 

6.6 

ND(1.6) 

22 

13 to 1150 

ND(1.2) to 22 ND(0.9) to 38 ND(0.9) to 31 

ND(0.9) to 33 

ND(0.9) to ND(4) 

ND(0.9) to 34 ND(3.6) to 70 ND(1.2) to 10 

ND(54) to 168 

9 6 6 

2 1 0 1 7 1 8 

120 

ND(7.2) ND(6.3) 

ND(8.4) ND(7.1) ND(6.2) ND(8.2) 

ND(23) ND(22) 

190 

NN

ND(ND(ND(ND(

N

ND(

ND

Total 2,3,7,8-CDDa,b 55.7 798 3,212 

Total 2,3,7,8-CDFa,b 20.2 272.5 310 Total I-TEQDF 8.28 90.12 58.89 

 (nondetect = 0)b 91.72 66.57 Total I-TEQDF 8.56 

 (nondetect = 1⁄2 DL)b 

Total TEQDF-WHO98 8.24 89.47 56.02 

 (nondetect = 0)b 

Total TEQDF-WHO98 8.59 91.7 66.09 

 (nondetect = 1⁄2 DL)b 

Total CDD/CDFb 75.9 1,070.5 3,522 

aCalculated assuming nondetect values were zero.bSum of median values. 


 

DL = Detection limit 


 


 

ND = Not detected (value in parenthesis is the detection limit)
 Source: U.S. EPA (1990e).
 
2,3,7,8-TCDD

1,2,3,7,8-PeCDD

1,2,3,4,7,8-HxCDD1,2,3,6,7,8-HxCDD

1,2,3,7,8,9-HxCDD

1,2,3,4,6,7,8-HpCDD

OCDD

2,3,7,8-TCDF

1,2,3,7,8-PeCDF

2,3,4,7,8-PeCDF

1,2,3,4,7,8-HxCDF1,2,3,6,7,8-HxCDF1,2,3,7,8,9-HxCDF

2,3,4,6,7,8-HxCDF

1,2,3,4,6,7,8-HpCDF1,2,3,4,7,8,9-HpCDF

OCDF

2,3,7,8-TCDD

1,2,3,7,8-PeCDD

1,2,3,4,7,8-HxCDD1,2,3,6,7,8-HxCDD

1,2,3,7,8,9-HxCDD

1,2,3,4,6,7,8-HpCDD

OCDD

2,3,7,8-TCDF

1,2,3,7,8-PeCDF2,3,4,7,8-PeCDF

1,2,3,4,7,8-HxCDF

1,2,3,6,7,8-HxCDF1,2,3,7,8,9-HxCDF

2,3,4,6,7,8-HxCDF

1,2,3,4,6,7,8-HpCDF1,2,3,4,7,8,9-HpCDF

OCDF

0.00%

10.00%

20.00%

30.00%

40.00%

50.00%

60.00%

70.00%

Percent of Total CDDs and CDFs 

Congener

Figure 8-1. 104 Mill Study full congener analysis results for pulp (nondetects 

equal to zero). 

Source: Median concentrations from U.S. EPA (1990e). 

0%

10%

20%

30%

40%

50%

60%

70%

Percent of Total CDDs and CDFs 

Congener

Figure 8-2. 104 Mill Study full congener analysis results for sludge. 

Source: Median concentrations from U.S. EPA (1990e). 

8-3 
2,3,7,8-TCDD

1,2,3,7,8-PeCDD

1,2,3,4,7,8-HxCDD

1,2,3,6,7,8-HxCDD

1,2,3,7,8,9-HxCDD

1,2,3,4,6,7,8-HpCDD

OCDD

2,3,7,8-TCDF

1,2,3,7,8-PeCDF2,3,4,7,8-PeCDF

1,2,3,4,7,8-HxCDF1,2,3,6,7,8-HxCDF

1,2,3,7,8,9-HxCDF2,3,4,6,7,8-HxCDF

1,2,3,4,6,7,8-HpCDF1,2,3,4,7,8,9-HpCDF

OCDF

0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

Congener

Figure 8-3. 104 Mill Study full congener analysis results for effluent. 

Source: Median concentrations from U.S. EPA (1990e). 

EPA concluded that 2,3,7,8-TCDF accounted for 77.8 to 99.5% of the total TEQDF-WHO98 (77.3 

to 99% of the total I-TEQDF) in pulp, 70.4 to 96.5% of the TEQDF-WHO98 (69.9 to 95.8% of the 

I-TEQDF) in sludge, and 75.0 to 96.4% of the TEQDF-WHO98 (71.3 to 91.7% of the I-TEQDF) in 

effluent. 

NCASI reported on a similar full-congener analysis study for samples collected from 

eight mills during the mid-1990s (letter dated February 10, 1997, from W. Gillespie, National 

Council of the Paper Industry for Air and Stream Improvement, Inc., to G. Schweer, Versar, 

Inc.). The results of these analyses are presented in Table 8-2. The frequency of detection of 

2,3,7,8-TCDD and 2,3,7,8-TCDF was significantly lower than in the 1988 study; therefore, 

deriving meaningful summary statistics concerning the relative importance of 2,3,7,8-TCDD and 

2,3,7,8-TCDF to the total TEQ is difficult. With all nondetect values assumed to be zero, 

2,3,7,8-TCDD and 2,3,7,8-TCDF accounted for 97% of the total effluent TEQDF-WHO98 (91% of 

the I-TEQDF), 53% of the total sludge TEQDF-WHO98 (46% of the I-TEQDF), and 87% of the total 

pulp TEQDF-WHO98 (87% of the I-TEQDF). Because of the high frequency of nondetects when all 

nondetect values are one-half the DL, 2,3,7,8-TCDD and 2,3,7,8-TCDF accounted for only 13% 

of the total effluent I-TEQDF, 13% of the total sludge I-TEQDF, and 28% of the total pulp I

TEQDF. 

­

8-4
 

Table 8-2. CDD/CDF concentrations in pulp and paper mill bleached pulp, wastewater sludge, and 

effluent (mid-1990s) 

8-5
 


Congener/congener 

group 

Bleached pulp Wastewater sludge W

Mean 

nondetect 

= 0 

(ng/kg) 

Median 

(ng/kg) 

Range 

(ng/kg) 

No. of 

detects/ 

 samples 

Mean 

nondetect 

= 0 

(ng/kg) 

Median 

(ng/kg) 

Range 

(ng/kg) 

No. of 

detects/ 

 samples 

Mean 

nondetect 

= 0 

(pg/L) 

Med

(ng/k

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 1,2,3,4,7,8-HxCDD 1,2,3,6,7,8-HxCDD 1,2,3,7,8,9-HxCDD 1,2,3,4,6,7,8-HpCDD 

OCDD 

0.3 

0 0 0 0 0 

2.4 

ND(1) 

ND(5) ND(5) ND(5) ND(5) ND(5) 

ND(10) 

ND(1) to 5 

ND(3) to ND(7) ND(3) to ND(7) ND(3) to ND(7) ND(3) to ND(7) ND(3) to ND(7) 

ND(10) to 15 

1/18 

0/18 0/18 0/18 0/18 0/18 

3/16 

0.8 

0 0.5 2.3 1.6 

41.4 

445 

ND(1) 

ND(5) ND(5) ND(5) ND(5) 

7 

150 

ND(1) to 4 

ND(4) to ND(52) 

ND(4) to 7 

ND(4) to 18 ND(4) to 14 

ND(4) to 330 

21 to 2,900 

4/12 

0/12 1/13 2/13 2/13 9/13 

10/10 

1.2 

0 0 0 0 3.2 

99 

ND

NDNDNDNDND

ND(

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 1,2,3,4,7,8-HxCDF 1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF OCDF 

10.3 

0 0.4 0 0 0 0 0 0 0 

ND(1) 

ND(5) ND(5) ND(5) ND(5) ND(5) ND(5) ND(5) ND(5) 

ND(10) 

ND(1) to 170 

ND(3) to ND(7) 

ND(3) to 7 

ND(3) to ND(7) ND(3) to ND(7) ND(3) to ND(7) ND(3) to ND(7) ND(3) to ND(7) ND(3) to ND(7) 

ND(6) to ND(14) 

7/18 

0/18 1/18 0/18 0/18 0/18 0/18 0/18 0/18 0/18 

6.2 

0 0.5 0 0 0 0.5 1.2 0 0 

3 

ND(5) ND(5) ND(5) ND(5) ND(5) ND(5) ND(5) ND(5) 

ND(10) 

ND(1) to 31 

ND(4) to ND(52) 

ND(4) to 7 

ND(4) to ND(52) ND(4) to ND(52) ND(4) to ND(52) 

ND(4) to 6 

ND(4) to 10 

ND(4) to ND(52) 

ND(9) to ND(100) 

9/12 

0/13 1/13 0/13 0/13 0/13 1/13 2/13 0/13 0/13 

2.3 ND

NDNDNDNDNDNDNDND

ND(

Total 2,3,7,8-CDDa 2.7 491.6 103 

Total 2,3,7,8-CDFa 10.7 8.4 2.3 

Total I-TEQDF 1.53 3 1.5 

 (nondetect = 0)a 53.6 

Total I-TEQDF 6.4 12.9 1.4 

 (nondetect = 1⁄2 DL)a 66.5 

Total TEQDF-WHO98 1.5 2.6 

 (nondetect = 0)a 

Total TEQDF-WHO98 7.6 15.2 

 (nondetect = 1⁄2 DL)a 

a Sum of mean values. 

DL = Detection limit 
 

ND = Not detected (value in parenthesis is the detection limit)
 Source: Letter dated February 10, 1997, from W. Gillespie, National Council of the Paper Industry for Air and Stream Improvement, Inc., to G. Schw
In 1992, the pulp and paper industry conducted its own NCASI-coordinated survey of 

2,3,7,8-TCDD and 2,3,7,8-TCDF emissions (NCASI, 1993). Ninety-four mills participated in 

the study, and NCASI assumed that the remaining 10 (of 104) operated at the same levels as 

measured in the 1988 104 Mill Study. All nondetect values were counted as one-half the DL. If 

a DL was not reported, it was assumed to be 10 pg/L for effluent and 1 ng/kg for sludge or 

bleached pulp. The data used in the report were provided by individual pulp and paper 

companies that had been requested by NCASI to generate the data using the same protocols used 

in the 104 Mill Study. 

In 1993, as part of its efforts to develop revised effluent guidelines and standards for the 

pulp, paper, and paperboard industry, EPA published the development document for the 

guidelines and standards being proposed for this industry (U.S. EPA, 1993b). The development 

document presented 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. To estimate these discharges, 

EPA used the most recent information about each mill from four databases (104 Mill Study, EPA 

short-term monitoring studies at 13 mills, EPA long-term monitoring studies at eight mills, and 

industry self-monitoring data submitted to EPA). The 104 Mill Study data were used for only 

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. 

Gillespie (1994) and Gillespie (1995) reported the results of 1993 and 1994 updates, 

respectively, to the 1992 NCASI survey. As in the 1992 survey, companies were requested to 

follow the same protocols for generating data that were used in the 104 Mill Study. Gillespie 

(1994, 1995) reported that fewer than 10% of mills had 2,3,7,8-TCDD and 2,3,7,8-TCDF 

concentrations in effluent above the nominal DLs of 10 pg/L and 100 pg/L, respectively. EPA 

obtained similar results in its short- and long-term sampling for 18 mills; 2,3,7,8-TCDD was 

detected at four mills, and 2,3,7,8-TCDF was detected at nine mills (U.S. EPA, 1993b). 

Gillespie (1994) reported that wastewater sludges at most mills (90%) contained less than 

31 ng/kg of 2,3,7,8-TCDD and less than 100 ng/kg of 2,3,7,8-TCDF. Gillespie (1995) reported 

that 90% of the mills had 2,3,7,8-TCDD and 2,3,7,8-TCDF concentrations in sludge of less than 

17 ng/kg and 76 ng/kg, respectively, in 1994. U.S. EPA (1993b) reported similar results but 

found detectable levels of 2,3,7,8-TCDD and 2,3,7,8-TCDF in sludges from 64% and 85%, 

respectively, of the facilities sampled. 

Gillespie (1994) reported that nearly 90% of the bleached pulps contained less than 2 

ng/kg of 2,3,7,8-TCDD and less than 160 ng/kg of 2,3,7,8-TCDF. Gillespie (1995) reported that 

90% of the bleached pulps contained 1.5 ng/ng or less of 2,3,7,8-TCDD and 5.9 ng/kg or less 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 ng/kg of 2,3,7,8­

TCDD. 

8-6
 
On April 15, 1998, EPA promulgated effluent limitation guidelines and standards for 

certain segments of the pulp, paper, and paperboard industry (Federal Register, 1998c). The 

industry segments covered by this rulemaking (i.e., the bleached paper-grade kraft and soda 

subcategory and the paper-grade sulfite subcategory) are those segments responsible for more 

than 90% of the bleached chemical pulp production in the United States. For this rule, EPA 

updated the estimates of baseline loadings made in 1993 for the proposed rule by using more 

recent data collected by EPA, NCASI (including the 1994 NCASI survey), and individual 

facilities (U.S. EPA, 1997d). These revised estimates are presented in the last column in Table 

8-3. EPA projects that, after full compliance with these rules, annual TEQ discharges will be 

reduced to 5 g in effluent and 7 g in sludge. 

Table 8-3. Summary of bleached chemical pulp and paper mill discharges 

(g/yr) of 2,3,7,8-TCDD and 2,3,7,8-TCDF 

Matrix Congener 

U.S. EPA 

1988a 

NCASI 

1992b 

U.S. EPA 

1993c 

NCASI 

1993b 

NCASI 

1994b 

U.S. EPA 

1995d 

Effluent 2,3,7,8-TCDD 

2,3,7,8-TCDF 

TEQ 

201 

1,550 

356 

22 

99 

32 

71 

341 

105 

19 

76 

27 

14.6 

49.0 

19.5 

16 

120 

28 

Sludgee 2,3,7,8-TCDD 

2,3,7,8-TCDF 

TEQ 

210 

1,320 

343 

33 

118 

45 

-­

-­

177 

24 

114 

35 

18.9 

95.2 

28.4 

-­

-­

50 

Pulp 2,3,7,8-TCDD 

2,3,7,8-TCDF 

TEQ 

262 

2,430 

505 

24 

124 

36 

-­

-­

149 

22 

106 

33 

16.2 

78.8 

24.1 

-­

-­

40 a

Data from the 104 Mill Study (U.S. EPA, 1990e). The total discharge rate of congener or TEQ (based only on 

2,3,7,8-TCDD and 2,3,7,8-TCDF concentrations) was summed across all 104 mills. b

The total discharge rate of congener or TEQ (based only on 2,3,7,8-TCDD and 2,3,7,8-TCDF concentrations) was 

summed across all 104 mills. The daily discharge rates reported in NCASI (1993) and Gillespie (1994, 1995) were 

multiplied by a factor of 350 days/yr to obtain estimates of annual discharge rates. NCASI 1992 survey (NCASI, 

1993), 1993 update (Gillespie, 1994), and 1994 update (Gillespie, 1995). c

The discharges in effluent and sludge were estimated in U.S. EPA (1993b, 1997d) for January 1, 1993. The TEQ 

discharge in pulp was estimated by multiplying the 1988 discharge estimate by the ratio of the 1988 and1993 

effluent discharge estimates (i.e., the estimate of the reduction in 1988 discharges achieved by pollution prevention 

measures taken by the industry between 1988 and 1993). d

The discharges in effluent and sludge were estimated in U.S. EPA (1997d) for mid-1995. The TEQ discharge in 

pulp was estimated by multiplying the 1988 discharge estimate by the ratio of the 1988 and1995 effluent discharge 

estimates (i.e., the estimate of the reduction in 1988 discharges achieved by pollution prevention measures taken by 

industry between 1988 and 1995). e

Approximately 20.5% of the sludge generated in 1990 was incinerated. The remaining 79.5% was predominantly 

landfilled (56.5%) or placed in surface impoundments (18.1%); 4.1% was land-applied directly or as compost, and 

0.3% was distributed or marketed (U.S. EPA, 1993e). 

NCASI = National Council of the Paper Industry for Air and Stream Improvement, Inc. 

-- = No information given 

8-7
 
8.1.1. Estimates of National Emissions in 1987 and 1995 

The U.S. annual discharges of 2,3,7,8-TCDD and 2,3,7,8-TCDF are summarized in Table 

8-3 for each of the six surveys discussed above. EPA release estimates for 1988 (U.S. EPA, 

1990e) and 1995 (U.S. EPA, 1997d) are believed to best represent emissions in reference years 

1987 and 1995, respectively. During the period between EPA’s 104 Mill Study and issuance of 

the development document (U.S. EPA, 1993b), the U.S. pulp and paper industry reduced releases 

of CDDs/CDFs, primarily by instituting numerous process changes to reduce the formation of 

CDDs/CDFs during the production of chemically bleached wood pulp. Details on the process 

changes implemented are provided in U.S. EPA (1993b) and Gillespie (1995). Much of the 

reduction between 1988 and 1995 can be attributed to process changes for pollution prevention. 

The confidence ratings for these release estimates are judged to be high because direct 

measurements were made at virtually all facilities, providing a high level of confidence in both 

the production and the emission factor estimates. The best estimates of annual emissions in 1987 

(i.e., the 1988 estimates presented in Table 8-3) are 356 g TEQ/yr for effluent and 343 g TEQ/yr 

for sludge. The best estimates of annual emissions in 1995 (i.e., the 1995 estimates presented in 

Table 8-3) are 28 g TEQ/yr for effluent and 50 g TEQ/yr for sludge. The CDD/CDF content in 

bleached chemical wood pulp as a product is estimated to be approximately 505 g TEQ and 40 g 

TEQ in 1987 and 1995, respectively. Although EPA provided an estimate of contaminant levels 

of CDDs/CDFs in wood pulp, it is currently not known whether the dioxin contamination in the 

product actually resulted in a release to the open and circulating environment. 

In 1990, the majority (75.5%) of the wastewater sludge generated by these facilities was 

placed in landfills or in surface impoundments, with the remainder incinerated (20.5%), applied 

to land directly or as compost (4.1%), or distributed as a commercial product (less than 1%) 

(U.S. EPA, 1993e). Data on the disposition of wastewater sludges are available only for years 

1988 through 1995. On the basis of these data, the best estimate of TEQ applied to land (i.e., not 

incinerated or landfilled) is 14.1 g TEQ (4.1% of 343 g) for 1987 and 2 g (4.1% of 50 g) for 

1995. These emission estimates are assigned a high level of confidence on the basis of the high 

confidence ratings given to both the activity level and the emission factor estimates. 

8.1.2. Estimates of National Emissions in 2000 

In 2000, NCASI provided estimates of congener-specific CDD/CDF releases from the 

pulp and paper industry in effluent, wastewater residuals, and pulp (Gillespie, 2002). Emission 

factors were taken from “NCASI Handbook of Chemical Specific Information for SARA 

(Superfund Amendments and Reauthorization Act) Section 313 Form R Reporting.” Emission 

factors were compiled from valid test data supplied to NCASI by a variety of sources, including 

member companies that had performed the tests in response to a regulatory program. The mass 

8-8
 

throughput parameter of total pulp production (31.9 million metric tons/yr) was provided by the 

American Forest and Paper Association and included data from 12 elemental chlorine-free mills. 

The effluent flow from chemical pulp mills with aerated stabilization basins (1,509 million 

gal/day) and with activated sludge treatment (660 million gal/day) was taken from the NCASI 

database and included data from five aerated stabilization basin mills and three activated sludge 

treatment mills. The primary waste treatment residuals from pulp mills (0.974 million dry metric 

tons/yr) and the combined, secondary, and dredged waste treatment residuals from pulp mills 

(1.37 million dry metric tons/yr) were also taken from the NCASI database and included data 

from five mills for the primary residuals and data from three mills for the secondary residuals 

(Gillespie, 2002). 

Table 8-4 provides a breakdown of TEQDF-WHO98 concentrations and emissions by 

congener. Total TEQDFWHO98 concentrations were reported to be 0.49 pg/L, 1.76 ng/kg, and 

0.02 pg/g for effluent, sludge, and pulp, respectively. CDD/CDF emission estimates were 

reported as 1.02 g TEQDFWHO98/yr, 1.93 g TEQDFWHO98/yr, and 0.582 g TEQDFWHO98/yr for 

effluent, sludge, and pulp, respectively. 

Fifty-one percent of the sludge generated in 2000 was sent to landfills or lagoons. It is 

uncertain how much of the remaining 49% of the sludge was applied to land. However, a 

conservative estimate can be developed by applying the 4.1% used to develop the 1987 and 1995 

estimates. In this case, 0.08 g TEQDFWHO98/yr of sludge is estimated to have been applied to 

land in 2000. These estimates are assigned a high confidence rating because they are based on 

recent industry survey data; however, EPA is working with NCASI to develop a QA/QC protocol 

to monitor the data being collected. 

8.2. MANUFACTURE OF CHLORINE, CHLORINE DERIVATIVES, AND METAL 

CHLORIDES 

Testing of CDD/CDF emissions to air, land, or water from U.S. manufacturers of 

chlorine, chlorine derivatives, and metal chlorides on which to base estimates of national 

emissions has not been reported. Sampling of graphite electrode sludges from European chlorine 

manufacturers indicates high levels of CDFs. Limited sampling of chlorine derivatives and metal 

chlorides in Europe indicates low-level contamination in some products. 

8.2.1. Manufacture of Chlorine 

Chlorine gas is produced by electrolysis of brine electrolytic cells. Until the late 1970s, 

the primary type of electrolytic process used in the chlor-alkali industry to produce chlorine 

consisted of the use of mercury cells containing graphite electrodes. As shown in Table 8-5, high 

levels of CDFs have been found in several samples of graphite electrode sludge from facilities in 

8-9
 

Table 8-4. CDD/CDF TEQ concentrations and emissions for the paper and 

pulp industry by source 

Congener 

Effluent 

Waste treatment 

residuals not lagooned 

or landfilled (sludge) 

(49% not landfilled) Pulp 

TEQ 

conc.a 

(pg/L) 

TEQ 

emissions 

(ng/yr) 

TEQ 

conc.a 

(ng/kg) 

TEQ 

emissions 

(ng/yr) 

TEQ 

conc.a 

(pg/g pulp) 

TEQ 

emissions 

(ng/yr) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

1,2,3,4,6,7,8,9-OCDD 

0 

0.00e+00 

0.00e+00 

1.30e!01 

9.00e!02 

7.00e!02 

7.37e!02 

0.00e+00 

0.00e+00 

0.00e+00 

2.71e+08 

1.88e+08 

1.46e+08 

1.54e+08 

4.00e!01 

5.00e!02 

1.00e!02 

8.00e!02 

9.00e!02 

1.82e!01 

2.80e!01 

4.63e+08 

6.24e+07 

1.25e+07 

8.53e+07 

9.05e+07 

1.97e+08 

2.81e+08 

1.00e!02 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

3.00e!03 

3.04e!03 

2.90e+08 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

8.69e+07 

8.80e+07 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

1,2,3,4,6,7,8,9-OCDF 

1.00e!01 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

1.90e!02 

5.00e!03 

2.00e!03 

2.08e+08 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

3.96e+07 

1.04e+07 

4.17e+06 

4.00e!01 

1.00e!02 

1.00e!01 

4.00e!02 

1.00e!02 

5.00e!02 

0.00e+00 

1.70e!02 

0.00e+00 

3.70e!03 

4.66e+08 

1.25e+07 

1.25e+08 

4.63e+07 

1.25e+07 

5.15e+07 

0.00e+00 

1.83e+07 

0.00e+00 

3.93e+06 

1.00e!03 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

2.00e!03 

0.00e+00 

1.00e!03 

0.00e+00 

6.00e!05 

2.90e+07 

0.00e+00 

0.00e+00 

0.00e+00 

0.00e+00 

5.79e+07 

0.00e+00 

2.90e+07 

0.00e+00 

1.74e+06 

TEQDF-WHO98 0.49 1.02e+09 1.76 0.02 5.83e+08 

Residuals total 

Residuals not landfilled 

1.93e+09 

9.44e+08 a

TEQ concentrations are in TEQDF-WHO98. 

Source: Gillespie (2002). 

Europe. The CDFs predominate in these sludges, and the 2,3,7,8-substituted congeners account 

for a large fraction of the respective congener totals (Rappe et al., 1990a, 1991; Rappe, 1993; 

Strandell et al., 1994). During the 1980s, titanium metal anodes were developed to replace 

graphite electrodes (U.S. EPA, 1982b; Curlin and Bommaraju, 1991). Currently, no U.S. facility 

is 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). 

8-10
 

Table 8-5. CDD/CDF concentrations (μg/kg) in graphite electrode sludge 

from chlorine production 

Congener/congener group Sludge 1 Sludge 2 Sludge 3 Sludge 4 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

ND (0.006) 

ND (0.007) 

ND (0.018) 

ND (0.012) 

ND (0.016) 

0.095 

0.92 

26 

25 

12 

32 

7 

1.3 

0.87 

9.1 

8.1 

31 

ND (0.009) 

ND (0.009) 

ND (0.026) 

ND (0.016) 

ND (0.022) 

0.21 

2 

56 

55 

25 

71 

16 

2.8 

1.9 

19 

19 

76 

ND (0.009) 

ND (0.009) 

ND (0.029) 

ND (0.019) 

ND (0.025) 

0.25 

2.2 

57 

56 

24 

73 

15 

2.6 

2 

19 

20 

71 

ND 

ND (0.033) 

ND (0.49) 

ND (0.053) 

ND (1.2) 

0.055 

0.65 

52 

55 

27 

44 

12 

1.7 

1.3 

15 

14 

81 

Total 2,3,7,8-CDDa 

Total 2,3,7,8-CDFa 

Total I-TEQDF 

a 

Total TEQDF-WHO98 

a 

1.02 

152.37 

14.2 

14.1 

2.21 

341.7 

30.5 

30.4 

2.45 

339.6 

30.2 

30.2 

0.7 

303 

27.7 

27.6 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

ND (0.006) 

ND (0.070) 

ND (0.046) 

0.22 

0.92 

64 

75 

68 

24 

31 

ND (0.009) 

ND (0.009) 

ND (0.064) 

0.48

2 

150 

240 

140 

53 

76 

ND (0.009) 

ND (0.009) 

ND (0.074) 

0.56 

2.2 

140 

240 

140 

54 

71 

-­

-­

-­

-­

0.65 

-­

-­

-­

-­

81 

Total CDD/CDFa 263.14 661.48 647.76 -­ a

Calculated assuming nondetect values were zero. 

ND = Not detected (value in parenthesis is the reported detection limit) 

-- = No information given 

Sources: Rappe et al. (1991); Rappe (1993). 

Although the origin of the CDFs in graphite electrode sludge is uncertain, chlorination of 

the cyclic aromatic hydrocarbons (such as dibenzofuran) present in the coal tar used as a binding 

agent in the graphite electrodes has been proposed as the primary source (Strandell et al., 1994). 

8-11
 

For this reason, sludges produced using metal electrodes were not expected to contain CDFs. 

However, results of an analysis of metal electrode sludge from a facility in Sweden, analyzed as 

part of the Swedish Dioxin Survey, showed that the sludge contained high levels of CDFs 

(similar to those of the graphite sludge) and primarily nondetectable levels of CDDs (Strandell et 

al., 1994). The sludge showed the same type of CDF congener pattern reported by Rappe et al. 

(1991) and Rappe (1993). Strandell et al. suggested that chlorination of polyaromatic 

hydrocarbons present in the rubber linings of the electrolytic cell may have produced the CDFs 

found in the one sample analyzed. 

Although EPA does not regulate CDDs/CDFs specifically, it issued restrictions under 

RCRA on the land disposal of wastewater and sludges generated by chlorine manufacturers that 

use the mercury cell process and the diaphragm process (with graphite electrodes) (waste codes 

K071, K073, and K106) (40 CFR 268). 

The Chlorine Chemistry Council (CCC), a trade association representing manufacturers 

that produce and/or use chlorine, sampled the treated wastewater discharges to surface waters at 

seven chlor-alkali facilities in the U.S. (CCC, 2004). The results of this study are presented in 

Table 8-6. The measurements are intended to represent wastewater discharges in both 2000 and 

1995. As indicated in Table 8-6, 1.79 g TEQDF-WHO98 (1.81 g I-TEQ) were released to water 

from chlor-alkali facilities in 2000 and 1995. The emission estimates for 2000 and 1995 are 

assigned a high confidence rating, because they were derived from the testing of chlor-alkali 

facility treated wastewater. 

The CCC reported on air emissions from two chlor-alkali production facilities in 2000 

(CCC, 2004). Table 8-7 summarizes the CDD/CDF congener-specific and TEQ annual releases 

to air from the Dow Chemical chlor-alkali facility in Midland, MI, and the PPG Industries facility 

in Natrium, WV. Total releases to air in 2000 were 0.08 g TEQDF-WHO98 (0.08 g I-TEQ). There 

are no data showing air releases from chlor-alkali facilities in 1995 and 1987. A high 

confidence rating is to assigned the estimate for 2000 because it is based on actual measurements 

of chlor-alkali facilities. 

8.2.2. Manufacture of Chlorine Derivatives and Metal Chlorides 

The limited sampling of chlorine-derivative products indicates that they contain very low, 

if any, concentrations of CDDs/CDFs. Rappe et al. (1990b) analyzed a sample of chlorine bleach 

consisting of 4.4% sodium hypochlorite. Most of the 2,3,7,8-substituted CDD/CDF congeners 

were below the limits of detection (0.3 to 7 pg/L for all congeners except OCDD and OCDF, 

which were 12 and 20 pg/L, respectively). No 2,3,7,8-substituted CDDs were detected. Tetra-, 

penta-, and hexa-CDFs were detected at levels of 13 pg/L or lower. The TEQ content of the 

sample was 4.9 pg I-TEQDF/L. Hutzinger and Fiedler (1991b) reported finding no CDDs/CDFs at 

8-12
 

Table 8-6. Releases of dioxin-like compounds in wastewater discharges from 

chlor-alkali and mixed chemical manufacturing facilities to surface water in 

reference years 2000 and 1995 

PPG 

Congener 

Occidental Chemical Corporation Industries 

Total 

Battleground, 

TX 

Deer 

Park, TX 

Delaware 

City, DE 

Hahnville, 

LA 

Mobile, 

AL 

Muscle 

Shoals, 

AL 

Natrium, 

WV 

2,3,7,8 TCDD 0.00 0.02 0.00 0.00 0.00 0.00 0.00 0.02 

1,2,3,7,8 PeCDD 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 

1,2,3,4,7,8 HxCDD 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 

1,2,3,6,7,8 HxCDD 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 

1,2,3,7,8,9 HxCDD 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 

1,2,3,4,6,7,8 HpCDD 0.00 0.08 0.00 0.01 0.00 0.00 0.22 0.31 

OCDD 0.48 21.50 4.09e!03 0.10 1.15e!03 1.13e!09 3.13 25.22 

2,3,7,8 TCDF 0.00 0.63 1.02e!03 0.19 2.88e!04 3.94e!08 0.06 0.89 

1,2,3,7,8 PeCDF 0.00 1.20 0.00 0.90 0.00 1.33e!07 0.06 2.16 

2,3,4,7,8 PeCDF 0.00 0.21 0.00 0.85 0.00 7.99e!08 0.33 1.39 

1,2,3,4,7,8 HxCDF 0.00 2.11 0.00 2.96 0.00 1.85e!07 0.11 5.18 

1,2,3,6,7,8 HxCDF 0.00 0.38 0.00 1.18 0.00 9.76e!08 0.00 1.56 

1,2,3,7,8,9 HxCDF 0.00 2.03e!03 0.00 0.63 0.00 2.29e!08 0.00 0.63 

2,3,4,6,7,8 HxCDF 0.00 0.00 0.00 0.60 0.00 3.28e!08 0.00 0.60 

1,2,3,4,6,7,8 HpCDF 0.00 0.59 2.31e!03 4.47 6.49e!04 1.32e!07 0.15 5.22 

1,2,3,4,7,8,9 HpCDF 0.00 5.66e!03 0.00 0.69 0.00 6.30e!08 0.00 0.69 

OCDF 0.00 4.88 0.00 1.75 0.00 1.34e!07 0.66 7.29 

Total I-TEQ 4.83e!04 0.53 1.29e!04 1.08 3.64e!05 8.65e!08 0.19 1.80 

Total TEQDF-WHO98 4.83e!05 0.51 1.26e!04 1.08 3.54e!05 8.63e!08 0.19 1.59 

Source: CCC (2004). 

a DL of 4 μg/kg in chlorine gas or in samples of 10% sodium hypochlorite, 13% sodium 

hypochlorite, and 31 to 33% hydrochloric acid at a DL of 1 μg/kg. 

Hutzinger and Fiedler (1991b) reported the results of analyses of samples of ferric 

trichloride (FeCl3), aluminum trichloride (AlCl3), copper chloride (CuCl2), cuprous chloride 

(CuCl), silicon tetrachloride (SiCl4), and titanium tetrachloride (TiCl4) for their HpCDF, OCDF, 

HpCDD, and OCDD content. The sample of FeCl3 contained HpCDF and OCDF in the low 

micrograms per kilogram range, but no HpCDD or OCDD was detected at a DL of 0.02 μg/kg. 

One of the two samples of AlCl3 analyzed also contained a low (μg/kg) concentration of OCDF. 

The samples of CuCl2 and CuCl contained concentrations of HpCDF, OCDF, and OCDD of less 

than 1 μg/kg. The results are presented in Table 8-8. 

8-13
 
Table 8-7. Congener-specific and TEQ annual releases to air (g/yr) from 

chlor-alkali and mixed chemical production facilities in 2000 

Congener 

Dow Chemical Co. 

Midland, MI 

PPG Industries 

Natrium, WV 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.027 

0.004 

0.008 

0.000 

0.003 

0.020 

0.086 

0.003 

0.000 

0.000 

0.002 

0.002 

0.087 

0.208 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDD 

1,2,3,4,7,8,9-HpCDD 

OCDF 

0.023 

0.009 

0.008 

0.066 

0.000 

0.002 

0.009 

0.148 

0.028 

0.225 

0.044 

0.003 

0.030 

0.044 

0.006 

0.006 

0.022 

0.142 

0.039 

0.064 

Total I-TEQDF 0.047 0.034 

Total TEQDF-WHO98 0.049 0.033 

Source: Chlorine Chemistry Council (2004). 

8.3. MANUFACTURE OF HALOGENATED ORGANIC CHEMICALS 

Several chemical production processes generate CDDs/CDFs (Versar, Inc., 1985; 

Hutzinger and Fiedler, 1991b). CDDs/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/CDFs to the environment. Also, the products 

themselves may contain these compounds, and their use or consumption may result in additional 

releases to the environment. 

CDD/CDF congener distribution patterns indicative of noncombustion sources have been 

observed in sediments in southwest Germany and the Netherlands. According to Ree et al. 

(1988), the congener patterns found suggest that wastes from the production of chlorinated 

organic compounds may be important historical sources of CDD/CDF contamination in these 

8-14
 

Table 8-8. CDD/CDF concentrations (μg/kg) in metal chlorides 

Congener group FeCl3 AlCl3 

a AlCl3 a CuCl2 CuCl TiCl4 SiCl4 

Total TCDD -­ -­ -­ -­ -­ -­ -­

Total PeCDD -­ -­ -­ -­ -­ -­ -­

Total HxCDD -­ -­ -­ -­ -­ -­ -­

Total HpCDD ND ND ND 0.03 ND ND ND 

Total OCDD ND ND 0.1 0.60 0.03 ND ND 

Total TCDF -­ -­ -­ -­ -­ -­ -­

Total PeCDF -­ -­ -­ -­ -­ -­ -­

Total HxCDF -­ -­ -­ -­ -­ -­ -­

Total HpCDF 12 ND ND 0.10 0.08 ND ND 

Total OCDF 42 ND 34.0 0.50 0.20 ND ND a

AlCl3 was tested twice. 

ND = Not detected; detection limit of 0.02 μg/kg 

-- = No information given 

Source: Hutzinger and Fiedler (1991b). 

regions. The production and use of many of the chlorophenols, chlorophenoxy herbicides, and 

PCB products are now 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 under certain limited use and disposal 

conditions (Rappe, 1992b). 

8.3.1. Chlorophenols 

Chlorophenols have been widely used for a variety of pesticidal applications. The higherchlorinated phenols (tetrachlorophenol and pentachlorophenol [PCP]) and their sodium salts 

have been used primarily for wood preservation. The lower-chlorinated phenols have been used 

primarily as chemical intermediates in the manufacture of other pesticides. For example, 2,4­

dichlorophenol is used to produce the herbicides 2,4-dichlorophenoxyacetic acid (2,4-D), 4-(2,4­

dichlorophenoxy)butanoic acid (2,4-DB), 2-(2,4-dichlorophenoxy)-propanoic acid (2,4-DP), 

Nitrophen, Genite, and Zytron and 2,4,5-trichlorophenol was used to produce hexachlorophene, 

2,4,5-T, Silvex, Erbon, Ronnel, and Gardona (Gilman et al., 1988; Hutzinger and Fiedler, 

1991b). (Sections 8.3.7 and 8.3.8 contain information on EPA actions to control CDD/CDF 

contamination of pesticides, including PCP and its salts, and to obtain additional data on 

CDD/CDF contamination of pesticides.) 

8-15
 
The two major commercial methods used to produce chlorophenols are (1) electrophilic 

chlorination of molten phenol by chlorine gas in the presence of catalytic amounts of a metal 

chloride and organic chlorination promoters and stabilizers, and (2) alkaline hydrolysis of 

chlorobenzenes under heat and pressure using aqueous methanolic sodium hydroxide. Other 

manufacturing methods include conversion of diazonium salts of various chlorinated anilines and 

chlorination of phenolsulfonic acids and benzenesulfonic acids, followed by the removal of the 

sulfonic acid group (Gilman et al., 1988; Hutzinger and Fiedler, 1991b). 

Because of the manufacturing processes employed, commercial chlorophenol products 

can contain appreciable amounts of impurities (Gilman et al., 1988). During the direct 

chlorination of phenol, CDDs/CDFs can form either by the condensation of tri-, tetra-, and 

pentachlorophenols or by the condensation of chlorophenols with hexachlorocyclohexadienone 

(which forms from excessive chlorination of phenol). During alkaline hydrolysis of 

chlorobenzenes, CDDs/CDFs can form through chlorophenate condensation (Ree et al., 1988; 

Gilman et al., 1988; Hutzinger and Fiedler, 1991b). 

The limited information on CDD/CDF concentrations in chlorophenols published in the 

1970s and early 1980s was compiled by Versar, Inc. (1985) and Hutzinger and Fiedler (1991b). 

The results of several major studies cited by these reviewers (Firestone et al., 1972; Rappe et al., 

1978a, b) are presented in Table 8-9. Typically, CDDs/CDFs were not detected in mono- and 

dichlorophenols but were reported in tri- and tetrachlorophenols. More recent results of testing 

of 2,4-dichlorophenol, performed in response to the Toxic Substances Control Act (TSCA) 

dioxin/furan test rule, showed no detectable concentrations of 2,3,7,8-substituted tetra- through 

hepta-CDD/CDFs. 

Other than a study by Hagenmaier (1986), which reported finding 2,3,7,8-TCDD at a 

concentration of 0.3 μg/kg in a sample of 2,3,4,5-tetrachlorophenol, no more recent data on 

concentrations of CDDs and CDFs could be found in the literature for the mono- through 

tetrachlorophenols. Tables 8-10 and 8-11 present summaries of several studies that reported 

CDD/CDF concentrations in PCP and in PCP-Na products, respectively. Many of these studies 

do not report congener-specific concentrations, and many are based on products obtained from 

non-U.S. sources. 

8.3.1.1. Regulatory Actions for Chlorophenols 

Section 8.3.8 of this report describes regulatory actions taken by EPA to control the 

manufacture and use of chlorophenol-based pesticides. In the mid-1980s, EPA’s Office of Solid 

Waste (OSW) promulgated, under RCRA, land disposal restrictions on wastes (wastewaters and 

nonwastewaters) resulting from the manufacture of chlorophenols (40 CFR 268). Table 8-12 

lists all wastes in which CDDs/CDFs are specifically regulated by EPA as hazardous 

8-16
 
Table 8-9. CDD/CDF concentrations (mg/kg) in mono- through tetrachlorophenols 

8-17
 


Congener/ 

congener 

group 2-CPa 2,4-DCPa 2,6-DCPa 

2,4,5-TrCP 

(Na salt)a 2,4,5-TrCPa 2,4,6-TrCPa 

2,4,6-TrCP 

(Na salt)b 2,3

Total TCDD ND (0.02) ND (0.02) ND (0.02) ND (0.02) to 14 ND (0.02) to 6.5 ND (0.02) to 49 <0.02 NTotal PeCDD ND (0.02) ND (0.02) ND (0.02) ND (0.02) ND (0.02) to 1.5 ND (0.02) <0.03 N

Total HxCDD ND (0.02) ND (0.02) ND (0.02) ND (0.02) ND (0.02) ND (0.02) <0.03 ND 

Total HpCDD ND (0.02) ND (0.02) ND (0.02) ND (0.02) ND (0.02) ND (0.02) <0.1 ND 

Total OCDD ND (0.02) ND (0.02) ND (0.02) ND (0.02) ND (0.02) ND (0.02) <0.1 ND (

Total TCDF + ND ND ND ND + 1.5 

Total PeCDF ND ND ND ND ND + 17.5 Total HxCDF ND ND ND ND ND + 36 Total HpCDF ND ND ND ND ND ND 4.8 

Total OCDF ND ND ND ND ND ND -­ 

TOTAL -­ -­ -­ -­ -­ -­ -­ 

aSource: Firestone et al. (1972); because of poor recoveries, the authors stated that actual CDD/CDF levels may have been considerably 

reported. 

bSources: Rappe et al. (1978a, b); common Scandinavian commercial chlorophenols. 

ND = Not detected (value in parenthesis is the detection limit, if reported) + = Detected but not quantified -- = No information given 
Table 8-10. CDD/CDF concentrations (historical and current) (μg/kg) in technical-grade pentachlo

(PCP) products 

8-18
 


Congener/ 

congener group 1973a 1978b 1979c 1984d 1985e 1986e 1987f 1987g 1985–88h 19

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

ND (10) 

ND (10) 

-­

2,200 

100 

100,000 

610,000 

ND (0.05) 

ND (1) 

6 

2,565 

44 

210,000 

1,475,000 

ND (0.05) 

ND (1) 

8 

1,532 

28 

106,000 930,000 

ND (0.03) 

1 

ND (1) 

831 

28 

78,000 

733,000 

ND (0.05) 

2 

ND (1) 

1,480 

53 

99,900 

790,000 

ND (0.05) 

ND (1) 

8 

600 

13 

89,000 

2,723,000 

1,100

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

130,000 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

ND (10) 

-­

-­

-­

-­

-­

-­

-­

-­

130,000 

ND (0.5) 

ND (1) 

ND (1) 

49 

5 

5 

ND (1) 

34,000 

4,100 

222,000 

ND (0.5) 

ND (1) 

ND (1) 

34 

4 

ND (1) 

ND (1) 

29,000 

6,200 

233,000 

ND (0.1) 

0.5 

1.5 

125 

ND (1) 

32 

ND (1) 

11,280 

637 

118,000 

ND (0.1) 

0.2 

0.9 

163 

ND (1) 

146 

ND (1) 

19,940 

980 

137,000 

ND (0.5) 

ND (1) 

ND (1) 

67 

2 

ND (1) 

ND (1) 

22,000 

3,400 

237,000 

170

Total 2,3,7,8-CDD1 

Total 2,3,7,8-CDF1 

Total I-TEQDF 1 

Total TEQDF-WHO98 1 

-­-­-­-­

-­-­-­-­

-­-­-­-­

712,300 130,0001,970

1,687,615 260,159 4,445

 1,037,568 

268,238 

2,735 

1,689 

811,860 130,076 

1,853 1,088 

891,435 158,230 

2,321 1,488 

2,812,621 

262,469 

4,173 

1,509 

1,10

170

$

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD Total OCDD Total TCDF 

Total PeCDF 

Total HxCDF Total HpCDF 

Total OCDF 

ND (20) 

ND (30) 

5,500 

98,000 

220,000 

40 

250 

22,000 

150,000 

160,000 

-­

-­

-­-­

-­

900 

4,000 

32,000 

120,000 

130,000 

-­

-­

10,100 

296,000 

1,386,000 

-­

1,400 

9,900 

88,000 

43,000 

1,304

ND (10) 

ND (10) 

4,500 

135,000 

610,000 

ND (10) 

-­

-­

62,000 

130,000 

2,918

ND 

ND 

4,694 

283,000 

1,475,000 

6 

10 

1,982 

125,000 

222,000 

ND 

ND 

2,925 

134,000 930,000 

ND 

3 

1,407 

146,000 

233,000 

1.9 

6.5 

1,700 

154,000 733,000 

0.8 

141 

4,300 

74,000 

118,000 

0.4 

15.2 

3,300 

198,000 790,000 

0.4 

343 

13,900 

127,000 137,000 

ND 

ND 

912 

117,000 

2,723,000 

ND 

200 

1,486 

99,000 

237,000 

ND 

ND 

130

1,100

ND 

ND 

1436

170

Total CDD/CDFl 655,790 286,900 1,834,400 941,500 2,111,692 1,447,335 1,085,150 1,269,559 3,178,598 1,45

aSource: Buser and Bosshardt (1976); mean of 10 samples of “high” CDD/CDF-content PCP received from Swiss commercial sources bSource: Rappe et al. (1978b); sample of U.S. origin, “presumably prepared by alkaline hydrolysis of hexachlorobenzene.” cSource: U.S. Department of Health and Human Services (1989); composite of technical-grade materials produced in 1979 by Monsant

(St. Louis, MO), Reichhold Chemicals, Inc. (White Plains, NY), and Vulcan Materials Co. (Birmingham, AL). 
Table 8-10. CDD/CDF concentrations (historical and current) in technical-grade pentachloropheno

products (μg/kg) (continued) 

d Source: Cull et al. (1984); mean of four “recent” production batches from each of two manufacturers of technical PCP using three diff

ANOVA showed no statistically significant difference in CDD/CDF concentrations between the eight samples (samples obtained in the 

e Source: Letter dated February 7, 1997, from John Wilkinson, Pentachlorophenol Task Force, to Matthew Lorber, U.S. EPA; average o

for the period January 1987 to August 1996. 

f Source: Hagenmaier and Brunner (1987); sample of Witophen P (Dynamit Nobel - Lot no. 7777) (obtained in Germany). g Source: Hagenmaier and Brunner (1987); sample of PCP produced by Rhone Poulenc (obtained in Germany). h Source: Letter dated February 7, 1997, from John Wilkinson, Pentachlorophenol Task Force, to Matthew Lorber, U.S. EPA; samples 

1985, 1986, and 1988. 

i Source: Harrad et al. (1991); PCP-based herbicide formulation from the New York State Department of Environmental Conservation. j Source: Letter dated March 5, 1997, from Thomas Mitchell, KMG-Bernuth, to Matthew Lorber, U.S. EPA; average of monthly batch 

February 1987 to December 1996 (excluding the following months, for which data were not available: February 1993, January 1992, DSeptember 1991, December 1988, and September 1988). 

k Source: Schecter et al. (1997); sample found stored in a barn in Vermont. l Calculated assuming nondetects were zero. 

ND = Not detected (value in parenthesis is the detection limit) -- = No information given 

8-19
 

Table 8-11. Historical CDD/CDF concentrations (μg/kg) in pentachlorophenol-Na 

(PCP-Na) 

Congener/congener 

group 1969a 1973b 1973c 1987d 1987e 1992f 1980sg 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

-­

-­

-­

-­

-­

-­

3,600 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

0.23 

18.2 

28.3 

2,034 

282 

9,100 

41,600 

0.51 

3.2 

13.3 

53 

19 

3,800 

32,400 

0.076 

18.7 

96 

4,410 

328 

175,400 

879,000 

ND (1.4) 

28.3 

ND (6.1) 

4,050 

ND (1.4) 

33,800 

81,000 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

1.8 

8.2 

6.6 

48 

69 

ND (1) 

87 

699 

675 

37,200 

0.79 

1.9 

1.1 

4.6 

1.3 

1.3 

4.6 

197 

36 

4,250 

ND (1) 

ND (4) 

ND (4) 

27.6 

21.9 

9.8 

103 

9,650 

2,080 

114,600 

149 

319 

324 

ND (2.8) 

225 

480 

ND (385) 

6,190 

154 

36,000 

Total 2,3,7,8-CDDh -­ -­ -­ 53,062.7 36,289 1,059,252.8 118,878.3 

Total 2,3,7,8-CDFh -­ -­ -­ 38,794.6 4,498.6 126,492.3 43,841 

Total I-TEQDF h -­ -­ -­ 452 89.5 3,374 1,201 

Total TEQDF-WHO98 h -­ -­ -­ 390 58.1 2,489 1,110 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

-­

-­

17,000 

9,600 

3,600 

-­

-­

-­

-­

-­

140 

40 

140 

1,600 

4,000 

ND (20) 

60 

1,400 

4,300 

4,300 

50 

ND (30) 

3,400 

38,000 

110,000 

ND (20) 

40 

11,000 

47,000 

26,500 

27 

213 

3,900 

18,500 

41,600 

82 

137 

3,000 

13,200 

37,200 

52 

31 

230 

5,800 

32,400 

12 

27 

90 

860 

4,250 

3.6 

142.7 

9,694 

260,200 

879,000 

10.1 

88.4 

9,082.3 

75,930 

114,600 

1.9 

140 

14,000 

100,000 

81,000 

1,200 

6,400 

49,000 

91,000 

36,000 

Total CDD/CDFh 30,200 15,980 235,990 117,859 43,752 1,348,751 378,742 

a

Source: Firestone et al. (1972); mean of two samples of PCP-Na obtained in the United States between 1967 and 

1969. 

b

Source: Buser and Bosshardt (1976); mean of five samples of “low” CDD/CDF-content PCP-Na received from 

Swiss commercial sources. c

Source: Buser and Bosshardt (1976); sample of “high” CDD/CDF-content PCP-Na received from a Swiss 

commercial source. d

Source: Hagenmaier and Brunner (1987); sample of Dowicide-G purchased from Fluka; sample obtained in 

Germany. e

Source: Hagenmaier and Brunner (1987); sample of Preventol PN (Bayer AG); sample obtained in Germany. f

Source: Santl et al. (1994); 1992 sample of PCP-Na from Prolabo, France. g

Source: Palmer et al. (1988); sample of a PCP-Na formulation collected from a closed sawmill in California in the 

late 1980s. h

Calculated assuming nondetect values were zero. 

ND = Not detected (value in parenthesis is the detection limit) 

-- = No information given 

8-20
 

Table 8-12. Summary of specific dioxin-containing wastes that must comply 

with land disposal restrictionsa 

EPA 

hazardous 

waste 

number Waste description 

Land disposal 

restriction 

effective date 

Regulated 

waste 

constituent 

F020 Wastes (except wastewater and spent carbon 

from HCl purification) from the production or 

manufacturing use (as a reactant, chemical 

intermediate, or component in a formulating 

process) of tri- or tetrachlorophenol or of 

intermediates used to produce their pesticide 

derivatives. (This listing does not include 

wastes from the production of hexachlorophene 

from highly purified 2,4,5-trichlorophenol.) 

November 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

F021 Wastes (except wastewater and spent carbon 

from HCl purification) from the production or 

manufacturing use (as a reactant, chemical 

intermediate, or component in a formulating 

process) of pentachlorophenol or of 

intermediates used to produce its derivatives. 

November 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

F022 Wastes (except wastewater and spent carbon 

from HCl purification) from the manufacturing 

use (as a reactant, chemical intermediate, or 

component in a formulating process) of tetra-, 

penta-, or hexachlorobenzenes under alkaline 

conditions. 

November 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

F023 Wastes (except wastewater and spent carbon 

from HCl purification) from the production of 

materials on equipment previously used for the 

production or manufacturing use (as a reactant, 

chemical intermediate, or component in a 

formulating process) of tri- and 

tetrachlorophenols. (This listing does not 

include wastes from equipment used only for the 

production or use of hexachlorophene from 

highly purified 2,4,5-trichlorophenol.) 

November 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

F026 Wastes (except wastewater and spent carbon 

from HCl purification) from the production of 

materials on equipment previously used for the 

manufacturing use (as a reactant, chemical 

intermediate, or component in a formulating 

process) of tetra-, penta-, or hexachlorobenzene 

under alkaline conditions. 

November 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

8-21
 

Table 8-12. Summary of specific dioxin-containing wastes that must comply 

with land disposal restrictionsa (continued) 

EPA 

hazardous 

waste 

number Waste description 

Land disposal 

restriction 

effective date 

Regulated 

waste 

constituent 

F027 Discarded unused formulations containing tri-, 

tetra-, or pentachlorophenol or discarded unused 

formulations containing compounds derived 

from these chlorophenols. (This listing does not 

include formulations containing 

hexachlorophene synthesized from prepurified 

2,4,5-trichlorophenol as the sole component.) 

November 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

F028 Residues resulting from the incineration or 

thermal treatment of soil contaminated with 

EPA Hazardous Wastes No. F020–F023, F026, 

and F027 

November 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

F039 Leachate (liquids that have percolated through 

land-disposed wastes) resulting from the 

disposal of more than one restricted waste 

classified as hazardous under Subpart D of 40 

CFR 268. (Leachate resulting from the disposal 

of one or more of the following EPA hazardous 

wastes and no other hazardous wastes retains its 

EPA hazardous waste number(s): F020, F021, 

F022, F026, F027, and/or F028.) 

August 8, 1990 

(wastewater) 

May 8, 1992 

(nonwastewater) 

TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

K043 2,6-Dichlorophenol waste from the production 

of 2,4-D 

June 8, 1989 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs 

K099 Untreated wastewater from the production of 

2,4-D 

August 8, 1988 TCDDs 

PeCDDs 

HxCDDs 

TCDFs 

PeCDFs 

HxCDFs a

For wastewater, the treatment standard for all regulated waste constituents except PeCDFs is 0.063 μg/L; the 

 standard for PeCDFs is 0.035 μg/L. For nonwastewater, the treatment standard for all regulated waste

 constituents is 1 μg/kg. Treatment standards are based on incineration to 99.9999% destruction and removal

 efficiency. 

Source: 40 CFR 268. 

8-22 
constituents, including chlorophenol wastes (waste codes F020 and F021). The regulations 

prohibit the land disposal of these wastes until they are treated to a level below the routinely 

achievable DLs for the EPA hazardous waste numbers listed in Table 8-12 for each of the 

following congener groups: TCDDs, PeCDDs, HxCDDs, TCDFs, PeCDFs, and HxCDFs. 

Wastes from PCP-based wood-preserving operations (waste codes K001 and F032) are also 

regulated as hazardous wastes under RCRA (40 CFR 261). 

EPA’s Office of Water promulgated effluent limitations for facilities that manufacture 

chlorinated phenols and discharge treated wastewater (40 CFR 414.70). These effluent 

limitations do not specifically regulate CDDs or CDFs. The effluent limitations for the 

individually regulated chlorinated phenols are less than or equal to 39 μg/L for facilities that use 

biological end-of-pipe treatment. 

Di- and trichlorophenols are subject to reporting under the dioxin/furan test rule, which is 

discussed in Section 8.3.7 of this report. Since the effective date of that rule (June 5, 1987), only 

the 2,4-dichlorophenol isomer has been commercially produced in (or imported to) the United 

States, and as noted in Table 8-9, no CDDs/CDFs were detected in the product. Testing is 

required for the other di- and trichlorophenols if manufacture or importation resumes. Similarly, 

tetrachlorophenols were subject to reporting under the Dioxin/Furan Pesticide Data Call-In (DCI) 

(discussed in Section 8.3.8 of this report). Since issuance of the DCI, the registrants of 

tetrachlorophenol-containing pesticide products have elected to no longer support the registration 

of their products in the United States. 

In January 1987, EPA entered into a settlement agreement with PCP manufacturers that 

set limits, effective in February 1989, on the allowed uses of PCP and its salts and the maximum 

allowable concentrations of 2,3,7,8-TCDD and HxCDDs. Section 8.3.8 discusses the 1987 PCP 

settlement agreement and includes estimates of current releases of CDDs/CDFs associated with 

the use of PCP in the United States. Section 11.3.1 (Chapter 11) provides an estimate of the 

amount of CDDs/CDFs that may have entered the environment or that are contained in treated 

wood products as a result of prior use of PCP and PCP-Na. 

Since the late 1980s, U.S. commercial production of chlorophenols has been limited to 

2,4-dichlorophenol and PCP. As noted above, disposal of wastes generated during the 

manufacture of chlorophenols is strictly regulated, and thus releases to the environment are 

expected to be negligible. With regard to releases associated with the use of 2,4-dichlorophenol, 

no CDDs/CDFs have been detected in 2,4-dichlorophenol. Releases associated with the use of 

PCP are presented in Sections 8.3.8 and 11.3.1. 

8-23
 

8.3.2. Chlorobenzenes 

Chlorobenzenes have been produced in the United States since 1909. U.S. production 

operations were developed primarily to provide chemical raw materials for the production of 

phenol, aniline, and various pesticides based on the higher-chlorinated benzenes. Because of 

(incremental) changes in the processes used to manufacture phenol and aniline and the phaseout 

of highly chlorinated pesticides such as DDT and hexachlorobenzene, U.S. production of 

chlorobenzenes in 1988 had decreased to 50% of the peak production level, in 1969. 

Chlorobenzenes can be produced via three methods: (1) electrophilic substitution of 

benzene (in liquid or vapor phase) with chlorine gas in the presence of a metal salt catalyst, (2) 

oxidative chlorination of benzene with HCl at 150 to 300°C in the presence of a metal salt 

catalyst, and (3) dehydrohalogenation of hexachlorocyclohexane wastes at 200 to 240°C with a 

carbon catalyst to produce trichlorobenzene, which can be further chlorinated to produce higherchlorinated benzenes (Ree et al., 1988; Hutzinger and Fiedler, 1991b; Bryant, 1993). 

All chlorobenzenes currently manufactured in the United States are produced by the 

electrophilic substitution process using liquid-phase benzene (i.e., temperature is at or below 

80°C). FeCl3 is the most common catalyst employed. Although this method can be used to 

produce mono- through hexachlorobenzene, the extent of chlorination is controlled to yield 

primarily monochlorobenzene and dichlorobenzene. The finished product is a mixture of 

chlorobenzenes, and refined products must be obtained by distillation and crystallization (Bryant, 

1993). 

CDDs/CDFs can be produced inadvertently during the manufacture of chlorobenzenes by 

nucleophilic substitution and pyrolysis mechanisms (Ree et al., 1988). The criteria required for 

production of CDDs/CDFs via nucleophilic substitution are oxygen as a nuclear substituent (i.e., 

presence of chlorophenols) and production or purification of the substance under alkaline 

conditions. Formation via pyrolysis requires reaction temperatures above 150°C (Ree et al., 

1988; Hutzinger and Fiedler, 1991b). The liquid-phase electrophilic substitution process 

currently used in the United States does not meet either of these criteria. Although Ree et al. and 

Hutzinger and Fiedler state that the criteria for formation of CDDs/CDFs via nucleophilic 

substitution may be present in the catalyst neutralization and purification/distillation steps of the 

manufacturing process, Opatick (1995) states that the chlorobenzene reaction product in U.S. 

processes remains mildly acidic throughout these steps. 

Table 8-13 summarizes the very limited published information on CDD/CDF 

contamination of chlorobenzene products. The presence of CDDs/CDFs has been reported in 

tri-, penta-, and hexachlorobenzene. No CDDs/CDFs have been reported in mono- or 

dichlorobenzene. Conflicting data exist concerning the presence of CDDs/CDFs in 

8-24
 
Table 8-13. CDD/CDF concentrations (μg/kg) in chlorobenzenes 

8-25
 


Congener/ congener group MCBza 

1,2-DCBz 

(for 

synthesis)a 

1,2,4-TrCBz 

(“pure”)b 

Mixed 

TrCBz 

(47%)a 

1,2,4,5-TCBz 

(99%)a 

PeCBz 

(98%)a 

HCB

(97%

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF Total OCDF 

ND (0.02) 

ND (0.02) 

ND (0.02) 

ND (0.02) ND (0.02) 

ND (0.02) 

ND (0.02) 

ND (0.02) 

ND (0.02) ND (0.02) 

0.3 

ND (0.02) 

ND (0.02) 

ND (0.02) ND (0.02) 

ND (0.02) 

0.5 

ND (0.02) 

ND (0.02) ND (0.02) 

ND (0.1) 

ND (0.1) 

ND (0.1) 

ND (0.1) ND (0.1) 

ND (0.1) 

ND (0.1) 

ND (0.1) 

ND (0.1) ND (0.1) 

0.027 

0.14 

0.259 

0.253 0.081 

0.736 

0.272 

0.091 

0.03 

0.016 

ND (0.02) 0.2 

0.5 

0.8 0.4 

0.03 

0.2 

0.8 

1.5 2.1 

ND (0.02) 

ND (0.02) 0.02 

0.02 0.05 

0.02 

ND (0.02) 

ND (0.02) 0.1 

0.1 

ND 

ND 

ND 

6

ND 

ND 

ND 

2

Total CDD/CDF ND 0.8

 ND 

1.9 

6.5 10

aSource: Hutzinger and Fiedler (1991b); unpublished results of tests performed at the University of Bayreuth, Germ0.3any, and by Dr. H. HbSource: Villanueva et al. (1974); range of three samples of commercially available HCBz. 

ND = Not detected (value in parenthesis is the detection limit, if reported) -- = No information given 
trichlorobenzene. One study (Villanueva et al., 1974) detected no CDDs/CDFs in one sample of 

1,2,4-TCBz at a DL of 0.1 μg/kg. Hutzinger and Fiedler (1991b) reported unpublished results of 

a study by Dr. Hans Hagenmaier showing CDD/CDF congener group concentrations ranging 

from 0.02 to 0.074 μg/kg in a sample of mixed trichlorobenzene. Because the TCBz examined 

by Hagenmaier contained about 2% hexachlorocyclohexane, it is reasonable to assume that it was 

produced by dehydrohalogenation of hexachlorocyclohexane (a manufacturing process not 

currently used in the United States). 

8.3.2.1. Regulatory Actions for Chlorobenzenes 

EPA determined, as part of the Federal Insecticide, Fungicide, and Rodenticide Act 

(FIFRA) DCI (discussed in Section 8.3.8), that the 1,4-dichlorobenzene manufacturing processes 

used in the United States are not likely to form CDDs/CDFs. Mono-, di-, and trichlorobenzene 

are listed as potential precursor chemicals under the TSCA dioxin/furan test rule and are subject 

to reporting (see Section 8.3.7). In addition, EPA issued a Significant New Use Rule (SNUR) 

under Section 5(a)(2) of TSCA on December 1, 1993 (effective January 14, 1994) for 

pentachlorobenzene and 1,2,4,5-tetrachlorobenzene (Federal Register, 1993c). This rule requires 

that EPA be notified at least 90 days before the manufacture, import, or processing of either of 

these compounds in amounts of 10,000 pounds or greater per year per facility for any use. All 

registrations of pesticide products containing hexachlorobenzene were cancelled in the mid1980s (Carpenter et al., 1986). 

OSW promulgated land disposal restrictions on wastes (i.e., wastewaters and 

nonwastewaters) resulting from the manufacture of chlorobenzenes (40 CFR 268). Table 8-12 

lists all solid wastes for which EPA specifically regulates CDDs and CDFs, including 

chlorobenzene wastes, as hazardous constituents. The regulations prohibit the land disposal of 

these wastes until they are treated to a level below the routinely achievable DLs in the waste 

extract listed in Table 8-12 for each of the following congener groups: TCDDs, PeCDDs, 

HxCDDs, TCDFs, PeCDFs, and HxCDFs. 

EPA’s Office of Water promulgated effluent limitations for facilities that manufacture 

chlorinated benzenes and discharge treated wastewater (40 CFR 414.70). These effluent 

limitations do not specifically address CDDs and CDFs. The following chlorinated benzenes are 

regulated: chlorobenzene; 1,2-dichlorobenzene; 1,3-dichlorobenzene; 1,4-dichlorobenzene; 

1,2,4-trichlorobenzene; and hexachlorobenzene. The effluent limitations for the individual 

regulated chlorinated benzenes are less than or equal to 77 μg/L for facilities that use biological 

end-of-pipe treatment and less than or equal to 196 μg/L for facilities that do not use biological 

end-of-pipe treatment. 

8-26
 
Since at least 1993, U.S. commercial production of chlorobenzenes has been limited to 

monochlorobenzene, 1,2-dichlorobenzene, 1,4-dichlorobenzene, and, to a much lesser extent, 

1,2,4-trichlorobenzene. As noted above, CDD/CDF formation is not expected under the normal 

operating conditions of the processes currently used in the United States to produce these four 

chemicals. No tetra-, penta-, or hexachlorinated benzenes are now intentionally produced or used 

in the United States (Bryant, 1993). Thus, releases of CDDs/CDFs from the manufacture of 

chlorobenzenes in 1995 were estimated to be negligible. Because the information available on 

CDD/CDF content of mono- through pentachlorobenzene is very limited and is based primarily 

on unpublished European data, and because information on the chlorobenzene manufacturing 

processes in place during 1987 is not readily available, no emission estimates can be made for 

1987. 

8.3.3. Chlorobiphenyls 

PCBs are manufactured by the direct batch chlorination of molten biphenyl in the 

presence of a catalyst, followed by separation and purification of the desired chlorinated biphenyl 

fractions. During the manufacture of PCBs, the inadvertent production of CDFs also occurs. 

This section focuses on levels of CDD/CDF contamination that may have been present in PCB 

products. 

CDFs have been shown to form when PCB-containing transformers and capacitors 

undergo malfunctions or are subjected to fires that result in accidental combustion of the 

dielectric fluid. The direct releases of dioxin-like PCBs from sources is reviewed in Chapter 10. 

During the commercial production of PCBs, thermal oxidative cyclization under alkaline 

conditions resulted in the inadvertent production of CDFs in most of the commercial PCB 

mixtures (Brown et al., 1988; ATSDR, 1993). Bowes et al. (1975a) first reported detection of 

CDFs in Aroclor products; samples of unused Aroclors manufactured in 1969 and 1970 were 

found to have CDF (TCDF through HxCDF) concentrations ranging from 0.8 to 2 mg/kg. Bowes 

et al. used congener-specific analytical methodology and detected 2,3,7,8-TCDF and 2,3,4,7,8­

PeCDF at concentrations ranging from 0.11 to 0.33 mg/kg and 0.12 to 0.83 mg/kg, respectively, 

in unused samples of Aroclor 1254 and Aroclor 1260. The presence of CDDs in commercial 

PCB mixtures, although at much lower concentrations than those of the CDFs, was reported by 

Hagenmaier (1987) and Malisch (1994). Table 8-14 presents the CDF and CDD congener group 

concentrations reported by Bowes et al. (1975a) and those reported in subsequent years for 

unused PCBs by Erickson (1986), ATSDR (1993), Hagenmaier (1987), and Malisch (1994). 

Several researchers have reported concentrations of specific CDD/CDF congeners in 

commercial PCB mixtures (Bowes et al., 1975b; Brown et al., 1988; Hagenmaier, 1987; Malisch, 

1994). Table 8-15 presents the results of these four studies. Only the Hagenmaier and Malisch 

8-27
 

Table 8-14. Concentrations of CDD/CDF congener groups in unused commercial polychlorinated b

mixtures (mg/kg) 

8-28
 


Year of 

CDD congener group concentrations CDF congener group concentrat

Total 

PCB mixture manufacture TCDD PeCDD HxCDD HpCDD OCDD CDD TCDF PeCDF HxCDF HpCDF O

Aroclor 1016 1972 -­ -­ -­ -­ -­ -­ ND ND ND -­ 

Aroclor 1242 -­ -­ -­ -­ -­ -­ -­ 0.07 0.03 0.003 -­ Aroclor 1242 -­ -­ -­ -­ -­ -­ -­ 2.3 2.2 ND -­ Aroclor 1242 -­ -­ -­ -­ -­ -­ -­ 0.25 0.7 0.81 -­ Clophen A-30 -­ 0.0007 ND 0.001 0.006 0.031 0.039 6.377 2.402 0.805 0.108 Clophen A-30 -­ ND ND ND 0.005 0.025 0.030 0.713 0.137 0.005 0.001 

Aroclor 1248 1969 -­ -­ -­ -­ -­ -­ 0.5 1.2 0.3 -­ Clophen A-40 -­ ND ND ND 0.012 0.030 0.042 1.289 0.771 0.144 0.02 

Kanechlor 400 -­ -­ -­ -­ -­ -­ -­ -­ -­ -­ -­ 

Aroclor 1254 1969 -­ -­ -­ -­ -­ -­ 0.1 0.2 1.4 -­ Aroclor 1254 1970 -­ -­ -­ -­ -­ -­ 0.2 0.4 0.9 -­ 

Aroclor 1254 -­ -­ -­ -­ -­ -­ -­ 0.02 0.2 0.6 -­ Aroclor 1254 -­ -­ -­ -­ -­ -­ -­ 0.05 0.1 0.02 -­ 

Clophen A-50 -­ ND ND ND 0.011 0.027 0.038 5.402 2.154 2.214 0.479 Aroclor 1260 

Aroclor 1260 Aroclor 1260 

Aroclor 1260 Clophen A-60 

Clophen A-60 

Clophen A-60 Phenoclor DP-6 

-­

1969 

-­

-­-­

-­

-­-­

-­

-­-­

-­

0.0004 

ND 

-­-­

-­

-­-­

-­

0.002 

ND 

-­-­

-­

-­-­

-­

0.002 

ND 

-­-­

-­

-­-­

-­

0.003 

0.014 

-­-­

-­

-­-­

-­

0.015 

0.032 

-­-­

-­

-­-­

-­

0.022 

0.046 

-­-­

0.3 

0.1 0.8 

0.2 

15.786 

16.34 

1.4 0.7 

1 

0.4 0.9 

0.3 

11.655 

21.164 

5 

10 

1.1 

0.5 0.5 

0.3 

4.456 

7.63 

2.2 2.9 

1.35 

-­

-­

-­

1.517 

2.522 

-­-­

Clophen T-64 -­ -­ -­ ­ ­ -­ -­ -­ 0.3 1.73 2.45 0.82 

Prodelec 3010 -­ -­ -­ ­ ­ -­ -­ -­ 1.08 0.35 0.07 -­ 

aSource: Bowes et al. (1975a). bSource: Erickson (1986). cSource: ATSDR (1993). dSource: Malisch (1994). eSource: Hagenmaier (1987). 

ND = Not detected -- = No information given 
Table 8-15. 2,3,7,8-Substituted congener concentrations in unused polychlorinated biphenyl (PCB) 

8-29
 


Congener 

Congener concentrations in Clophens Congener concentrations in Aroclo

A-30a A-30b A-40b A-50b A-60a A-60b 1016c 1242c 1248d 1254c 1254c 1254c 125

2,3,7,8-TCDD ND ND ND ND ND ND -­ -­ -­ -­ -­ -­ -­ 1,2,3,7,8-PeCDD ND ND ND ND 0.1 ND -­ -­ -­ -­ -­ -­ -­ 1,2,3,4,7,8-HxCDD ND ND ND ND 0.2 ND -­ -­ -­ -­ -­ -­ -­ 1,2,3,6,7,8-HxCDD 0.8 ND ND ND ND ND -­ -­ -­ -­ -­ -­ -­ 

1,2,3,7,8,9-HxCDD ND ND ND ND ND ND -­ -­ -­ -­ -­ -­ -­ 1,2,3,4,6,7,8-HpCDD 5.6 2.4 4.4 5.3 2.5 6.8 -­ -­ -­ -­ -­ -­ -­ OCDD 31.1 24.7 30.3 26.9 14.9 32.3 -­ -­ -­ -­ -­ -­ -­ 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

1,032.6 

135.8 509.2 

301.4 

65.3 

ND 

50.6 43.7 

22.5 

15.7 

36.9 14.9 13.1 

1.9 0.8 

ND 

0.1 0.6 

ND 

ND 

250.2 

52.7 

171.3 

48.4 19.6 

0.7 6.8 7 

2.8 

11.4 

1,005.7 

155.2 407.5 

647.5 227.5 

8.3 

62.5 

205.5 

72.2 69.2 

2,287.7 

465.2 

1,921.9 

1,604.2 

157.6 

42.8 

369.5 

480.6 

321.7 639.2 

3,077.2 1,750.8 2,917.0 

2,324.1 

351.3 

19 

4,08.3 

1,126.1 

304 

1,024.3 

0.1 

-­

1.75 

-­-­

0.08 

-­

-­

-­

-­

40.1 

-­

40.8 

-­-­

0.26 

-­

-­

-­

-­

330 

-­

830 

-­-­

-­-­

-­

-­

-­

28 

-­

110 

-­-­

28.8 

-­

-­

-­

-­

20.9 

-­

179 

-­-­

28.7 

-­

-­

-­

-­

55.8 

-­

105 

-­-­

19.4 

-­

-­

-­

-­

11

-­

12

-­-­

-­

-­

-­

-­

-­

Total TCDD 0.7 ND ND ND 0.4 ND -­ -­ -­ -­ -­ -­ -­ 

Total PeCDD ND ND ND ND 2 ND -­ -­ -­ -­ -­ -­ -­ Total HxCDD 1.2 ND ND ND 1.8 ND -­ -­ -­ -­ -­ -­ -­ 

Total HpCDD 5.6 5.4 11.6 11 3 13.5 -­ -­ -­ -­ -­ -­ -­ Total OCDD 31.1 24.7 30.3 26.9 14.9 32.3 -­ -­ -­ -­ -­ -­ -­ Total TCDF 6,376.6 713 1,289.4 5,402.3 15,785.7 16,340 -­ -­ -­ -­ -­ -­ -­ 

Total PeCDF 2,402.4 136.5 770.8 2,153.7 11,654.6 21,164 -­ -­ -­ -­ -­ -­ -­ 

Total HxCDF 804.8 5.1 143.6 2,213.8 4,455.8 7,630.2 -­ -­ -­ -­ -­ -­ -­ Total HpCDF 108.3 0.8 19.5 478.8 1,517 2,522.3 -­ -­ -­ -­ -­ -­ -­ 

Total OCDF 15.7 ND 11.4 69.2 639.2 1,024.3 -­ -­ -­ -­ -­ -­ -­ Total CDD/CDFe 9,746.4 885.5 2,276.6 10,355.7 34,074.4 48,726.6 -­ -­ -­ -­ -­ -­ -­ 

Total I-TEQDF e 407.2 11.3 409.6 1,439.2 2,179 -­ -­ -­ -­ -­ -­ -­ 

Total TEQDF-WHO98 e 407.2 11.3 409.5 1,439 2,178 -­ -­ -­ -­ -­ -­ -­ 

aSource: Malisch (1994).
 bSource: Hagenmaier (1987).
 cSource: Brown et al. (1988).
 dSource: Bowes (1975b).
 eCalculated assuming nondetect values were zero.

ND = Not detected 


 


 

-- = No information given
 
studies, however, reported the concentrations of all 2,3,7,8-substituted CDDs and CDFs. It is 

evident from the table that major variations are found in the levels of 2,3,7,8-TCDF and 

2,3,4,7,8-PeCDF in the Clophen mixtures reported by Hagenmaier and Malisch and the 

corresponding levels in the Aroclor mixtures reported by Bowes et al. and Brown et al. 

Brown et al. (1988) compared the levels of 2,3,7,8-TCDF; 2,3,4,7,8-PeCDF; and 

1,2,3,7,8,9-HxCDF in used samples (from previously used capacitors and transformers) and 

unused samples of Aroclor 1016, 1242, 1254, and 1260. The concentration ranges reported for 

the used and unused Aroclors were similar, leading Brown et al. (1988) to conclude that CDFs 

are not formed during the normal use of PCBs in electrical equipment. 

8.3.4. Ethylene Dichloride/Vinyl Chloride Monomer/Polyvinyl Chloride Manufacturing 

In the United States the manufacture of PVC is an integrated manufacturing process. 

This means that most manufacturing facilities produce all the precursors and chemical 

intermediates necessary to manufacture PVC as well as PVC resins and products. For example, 

ethylene dichloride (EDC) is used to produce vinyl chloride monomer (VCM). VCM is then 

converted to PVC resins. The resins are used to manufacture various PVC products. 

 PVC resins are produced from the polymerization of VCM. VCM is typically produced 

by the thermal dehydrochlorination (commonly known as cracking) of EDC. The cracking of 

EDC requires elevated pressure (20 to 30 atm) and temperature (450 to 650°C) and yields VCM 

and HCl at about a 1:1 molar ratio. EDC is produced by two different methods: (1) direct 

chlorination of ethylene with chlorine in the presence of a catalyst at a temperature of 50 to 60°C 

and pressure of 4 to 5 atm, and (2) oxychlorination, which involves reaction of ethylene with HCl 

and oxygen in the presence of a catalyst at temperatures generally less than 325°C. The primary 

source of HCl for the oxychlorination process is the HCl produced from the cracking of EDC to 

form VCM. Most VCM manufacturing facilities are integrated with EDC production facilities 

(The Vinyl Institute, 1998). 

Although it has been generally recognized that CDDs/CDFs can be formed during the 

manufacture of EDC, VCM, and PVC, manufacturers and environmental public interest groups 

have disagreed as to the quantity of CDDs/CDFs that are formed and released to the environment 

in wastes and possibly in PVC products. Although EPA regulates emissions from EDC/VCM 

production facilities under the Clean Water Act (40 CFR 61), the Clean Air Act (40 CFR 414), 

and RCRA (40 CFR 268, waste codes F024, K019, and K020), CDDs/CDFs are not specifically 

regulated pollutants; as a consequence, monitoring data for CDDs/CDFs in emissions were 

generally lacking until the early 1990s. 

Greenpeace International initially determined that CDDs and CDFs can be formed during 

the manufacture of PVC. In 1993, it issued a report on CDD/CDF emissions associated with the 

8-30
 

production of EDC/VCM (Greenpeace, 1993). Greenpeace estimated that 5 to 10 g I-TEQDF 

were released to the environment (air, water, and ground combined) annually for every 100,000 

metric tons of VCM produced. This emission factor was based on data gathered by Greenpeace 

on four European plants. The Vinyl Institute responded with a critique of the Greenpeace report 

(ChemRisk, 1993). Miller (1993) summarized the differing views of the two parties. According 

to Miller, European PVC manufacturers claimed the emission factor was 0.01 to 0.5 g ITEQDF/100,000 metric tons of VCM, but although Greenpeace and ChemRisk used basically the 

same monitoring information to develop their emission factors, Greenpeace adjusted the 

emission factor to account for unquantified fugitive emissions and waste products that contain 

unspecified amounts of CDDs/CDFs. 

In 1995, Greenpeace issued a second report (Stringer et al., 1995) reiterating the 

organization’s concern that the generation and emission of CDDs/CDFs may be significant and 

urging that further work be initiated to quantify and prevent emissions. Stringer et al. presented 

the results of analyses of three samples of chlorinated wastes obtained from U.S. EDC/VCM 

manufacturing facilities. The three wastes were characterized according to EPA hazardous waste 

classification numbers as an F024 waste (waste from the production of short-chain aliphatics by 

free radical-catalyzed processes), a K019 waste (heavy ends from the distillation of ethylene from 

EDC production), and a probable K020 waste (heavy ends from distillation of vinyl chloride in 

VCM manufacture). Table 8-16 presents the analytical results reported by Stringer et al. This 

study acknowledged that because EDC/VCM production technologies and waste treatment and 

disposal practices are very site-specific, the limited information available on CDD/CDF 

generation and emissions made it difficult to quantify amounts of CDDs/CDFs generated and 

emitted. 

In response to the lack of definitive studies, and at the recommendation of EPA, U.S. 

PVC manufacturers began an extensive monitoring program, the Dioxin Characterization 

Program (DCP). The objective of the DCP was to evaluate the extent and magnitude of 

potential CDD/CDF releases to air, water, and land, as well as the potential for PVC product 

contamination. Manufacturers performed emissions and product testing at several facilities that 

were representative of various manufacturing and process control technologies. In 1998, The 

Vinyl Institute completed studies of CDD/CDF releases in wastewater, wastewater treatment 

plant solids, and stack gases, as well as studies of the CDD/CDF content of products (PVC resins 

and EDC sold as products) (The Vinyl Institute, 1998). 

In September 2002, the CCC met to review dioxin release estimates for 2000 submitted 

by various EDC/VCM manufacturing facilities. Several companies provided stack gas emissions 

and wastewater release data as well as a discussion of how they generated the release and transfer 

estimates reported in the Toxics Release Inventory for 2000. In March 2004, the CCC met again 

8-31
 

Table 8-16. Reported CDD/CDF concentrations (μg/kg) in wastes from 

polyvinyl chloride (PVC) manufacture 

Congener/congener group F024 waste K019 waste K020 waste 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.37 

0.14 

0.3 

0.14 

0.11 

4.2 

15 

260 

890 

260 

330 

620 

920 

1,060 

0.06 

0.05 

0.08 

0.06 

0.07 

0.89 

3 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.91 

9.5 

1.6 

110 

24 

9.5 

3.1 

250 

51 

390 

680 

975 

1,050 

10,100 

9,760 

21,800 

930 

13,400 

1,340 

43,500 

0.44 

1.8 

0.58 

11 

2.4 

1.3 

0.89 

38 

6 

650 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

20.3 

849.6 

20 

19.7 

4,340 

103,535 

5,928 

6,333 

4.21 

712.4 

3.2 

2.6 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

3.1 

3.6 

1.3 

5 

15 

15 

65 

300 

450 

390 

1,230 

3,540 

3,950 

1,270 

1,060 

20,600 

45,300 

63,700 

16,600 

43,500 

1.9 

1.7 

a 

1.7 

3 

6 

11 

27 

58 

650 

Total CDD/CDF 1,248 200,750 760.3 a

Congener group concentration reported in source is not consistent with reported congener concentrations. 

Source: Stringer et al. (1995). 

8-32 
to discuss the results, to date, of its CDD/CDF data validation study for PVC/EDC/VCM and 

chlor-alkali facilities. The study’s goal was to provide facility-specific water, air, and land 

release estimates for 2000 and 2002. As of the date of this report, data validation studies were 

provided for 17 of 20 facilities in the CCC that were considered chlor-alkali production facilities 

and PVC/EDC/VCM manufacturing plants. 

8.3.4.1. Water Releases 

This section presents estimates of releases of dioxin-like compounds in wastewater 

discharges to surface waters for 2000 and 1995 from the integrated EDC, VCM, and PVC 

manufacturing facilities in the United States. Site-specific testing of the wastewater from 

facilities operated by Dow Chemical Co., Occidental Chemical Corp., Georgia Gulf Corp., PPG 

Industries, and DuPont provide the basis for estimating annual releases of dioxin-like compounds 

in 2000 (CCC, 2005, 2003a, b, c, d). In total, these tests represent site-specific analysis of 17 

manufacturing facilities throughout the United States. The site-specific releases of dioxin-like 

compounds for these facilities are shown in Table 8-17. In 2000, approximately 23.08 g TEQDFWHO98 (23.94 g I-TEQ) were released from 17 EDC/VCM/PVC manufacturing facilities in the 

United States. These estimates are assigned a high confidence rating because they were derived 

from the testing of individual manufacturing facilities and the activity level is known with a high 

degree of certainty. 

Estimating TEQ wastewater discharges for 1995 from EDC/VCM/PVC integrated 

chemical manufacturing facilities is problematic. A report by The Vinyl Institute (1998) did not 

represent a comprehensive testing of the wastewaters from all existing EDC/VCM/PVC 

manufacturing facilities in 1995. The report presented results for treated wastewater samples 

collected during April and May of 1995 at only 10 manufacturing sites (6 that manufactured only 

PVC, 3 that manufactured EDC and VCM, and 1 that manufactured EDC, VCM, and PVC). In 

terms of production, the 10 sites represented only about 27% of the total estimated 1995 

EDC/VCM/PVC production. 

The representativeness of these sites to total EDC/VCM/PVC manufacturing is 

questionable because the testing program did not include some of the higher-emitting facilities 

discovered in 2000 (CCC, 2003a). For example, the CCC reports that provided the basis for the 

2000 release estimates (CCC, 2005, 2003a, b, c, d) indicated that the Dow facility in Freemont, 

TX, and the Occidental facility in Ingleside, TX, released approximately 3 g and 1.6 g TEQDFWHO98, respectively, as wastewater discharges into surface waters. These two facilities are 

about an order of magnitude higher in wastewater releases than are the 10 facilities tested in 1995 

(The Vinyl Institute, 1998). In an attempt to overcome this deficiency, it is assumed that the 

wastewater releases from EDC/VCM/PVC integrated chemical production facilities in 1995 were 

8-33
 

Table 8-17. Releases of dioxin-like compounds (g/yr) in wastewater discharges from EDC/VCM/PV

integrated chlorine chemical manufacturing facilities to surface water in reference years 2000 and 

8-34
 


Company Facility 

2,3,7,8 TCDD

1,2,3,7,8 PeCDD

1,2,3,4,7,8 HxCDD

1,2,3,6,7,8 HxCDD

1,2,3,7,8,9 HxCDD

1,2,3,4,6,7,8 HpCDD

OCDD

2,3,7,8 TCDF

1,2,3,7,8 PeCDF

2,3,4,7,8 PeCDF

1,2,3,4,7,8 HxCDF

1,2,3,6,7,8 HxCDF

1,2,3,7,8,9 HxCDF

2,3,4,6,7,8 HxCDF

1234678 HpCDF

 Dowa Freeport, TX 0.00 0.00 0.57 0.00 0.35 15.61 74.61 2.63 3.57 1.68 26.28 0.00 0.00 5.00 125

 Dow Midland, MI 0.00 0.00 0.02 0.00 0.00 0.54 3.63 0.02 0.00 0.01 0.06 0.00 0.00 0.04 0

 Dow Plaquemine, LA 0.06 0.12 0.87 0.00 0.33 18.72 285.79 4.78 3.77 1.97 21.42 0.00 0.98 3.14 161

 DuPontb DeLisle, MS 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0

 DuPont Edge Moor, DE 0.00 0.00 0.00 0.00 0.00 0.00 0.10 0.00 0.00 0.00 0.08 0.00 0.00 0.00 0

 DuPont Johnsonville, TN 0.00 0.00 0.00 0.00 0.00 0.00 1.41 0.00 0.00 0.00 0.44 0.00 0.00 0.00 1

 Georgia Gulf c Plaquemine, LA 0.00 0.00 0.00 0.00 0.00 0.03 0.12 0.01 0.01 0.02 0.02 0.01 0.01 0.01 0

 Occidentald Convent, LA 0.00 0.00 0.00 0.00 0.00 0.00 0.15 0.00 0.00 0.00 0.02 0.00 0.00 0.00 0

 Occidental Deer Park, TX 0.00 0.00 0.00 0.00 0.00 0.08 1.19 0.00 0.00 0.00 0.04 0.03 0.03 0.01 0

 Occidental Ingleside, TX 0.02 0.09 0.08 0.09 0.08 0.09 0.18 0.02 0.09 0.09 0.09 0.10 0.10 0.10 0

 Occidental LaPorte, TX 0.00 0.00 0.00 0.00 0.00 0.00 0.05 0.00 0.00 0.00 0.04 0.00 0.00 0.00 0 Occidental Mobile, AL 0.00 0.00 0.00 0.00 0.00 0.00 0.00 3e!04 0.00 0.00 0.00 0.00 0.00 0.00 6e!

 Occidental Battleground, TX 0.00 0.00 0.00 0.00 0.00 0.00 0.48 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0 Occidental Delaware City, DE 0.00 0.00 0.00 0.00 0.00 0.00 0.00 1e!03 0.00 0.00 0.00 0.00 0.00 0.00 2e! Occidental Hahnville, LA 0.00 0.00 0.00 0.00 0.00 0.01 0.10 0.19 0.90 0.85 2.96 1.18 0.63 0.60 4

 Occidental Muscle Shoals, AL 0.00 0.00 0.00 0.00 0.00 0.00 1e!09 4e!08 1e!07 8e!08 2e!07 1e!07 2e!08 3e!08 1e!

 PPG Industries Lake Charles, LA 0.00 0.00 0.01 0.01 0.00 0.40 2.97 6.47 14.30 10.83 12.29 3.96 3.23 1.57 4

TOTAL 0.08 0.22 1.55 0.10 0.77 35.49 370.78 14.13 22.64 15.46 63.73 5.29 4.98 10.47 299
Table 8-17. Releases of dioxin-like compounds (g/yr) in wastewater discharges from EDC/VCM/PV

integrated chlorine chemical manufacturing facilities to surface water in reference years 2000 and (continued) 

aDow Chemical Company.bDuPont. cGeorgia Gulf Corporation.dOccidental Chemical Corporation.Source: Chlorine Chemistry Council (2004).EDC = Ethylene dichloridePVC = Polyvinyl chlorideVCM = Vinyl chloride monomer


 


 


 


 
 


 


 


 

8-35
 

equal to wastewater discharges in 2000. Based on this assumption, approximately 23.08 g 

TEQDF-WHO98 (23.94 g I-TEQ) were released from 17 EDC/VCM/PVC manufacturing facilities 

in the United States in reference year 1995. These estimates are assigned a low confidence rating 

because they may not be representative of wastewater discharges from facilities operating in 

1995, although the activity level is known with a high degree of certainty and is assigned a high 

confidence rating. The site-specific releases of dioxin-like compounds for facilities operating in 

1995 are shown in Table 8-17. 

A number of EDC/VCM/PVC manufacturing facilities have reported significant 

reductions in CDD/CDF wastewater discharges to surface water in 2002 from 2000 levels. For 

example, the Dow facility in Freemont, TX, achieved a 47% reduction (from 6.4 g TEQDFWHO98 in 2000 to 3.43 g TEQDF-WHO98 in 2002) (CCC, 2003a), the Dow facility in 

Plaquemine, LA, achieved a 69% reduction (from 6.86 g TEQDF-WHO98 in 2000 to 2.16 g 

TEQDF-WHO98 in 2002) (CCC, 2003a), and the PPG Industries facility in Lake Charles, LA, 

achieved a 28% reduction (from 8.97 g TEQDF-WHO98 in 2000 to 6.47 g TEQDF-WHO98 in 2002) 

(CCC, 2003b). 

8.3.4.2. Land Releases 

Only one EDC/VCM/PVC manufacturing facility (the Georgia Gulf facility in 

Plaquemine, LA) reported CDD/CDF releases to land from the land application of wastewater 

sludge in 2000 (CCC, 2003c). The congener-specific and TEQ releases are presented in Table 

8-18. Releases to land from EDC/VCM/PVC manufacturing in 2000 and 1995 were 1.36 g 

TEQDF-WHO98 (1.45 g I-TEQDF), assuming that the land application rates were the same in both 

years. 

These emission estimates for 2000 are assigned a high confidence rating because the 

releases to land were determined from the actual measurements taken from the single facility that 

applies the wastewater sludge to land. The emission estimates for 1995 are assigned a medium 

confidence rating because the estimates were based on assuming that the CDD/CDF levels of 

contamination in the wastewater solids and the land application rates were the same in both 

reference years. 

8.3.4.3. Air Releases 

The Vinyl Institute conducted a study of releases of EDC/VCM/PVC to air (The Vinyl 

Institute, 1998). Based on similarities in design and service, thermal destruction units at 

EDC/VCM and/or PVC manufacturing units were subcategorized into three types: type A, vent 

gas incinerators at PVC-only resin plants; type B, vent gas thermal oxidizers at EDC/VCM 

plants; and type C, liquid-only and liquid/vent gas thermal oxidizers at EDC/VCM plants. Using 

8-36
 

Table 8-18. Congener-specific releases to land from an EDC/VCM/PVC 

integrated chemical manufacturing facilitya in reference years 2000 and 1995 

Congener Annual release to land (g/yr) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.004 

0.039 

0.123 

0.122 

0.078 

1.710 

8.640 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDD 

1,2,3,4,7,8,9-HpCDD 

OCDF 

0.077 

0.354 

0.370 

2.690 

2.110 

1.540 

0.595 

28.100 

6.540 

118.000 

Total I-TEQDF 1.450 

Total TEQDF-WHO98 1.360 a

Georgia Gulf, Plaquemine, LA. 

Source: CCC (2004). 

EDC = Ethylene dichloride 

PVC = Polyvinyl chloride 

VCM = Vinyl chloride monomer 

an industry-wide survey, The Vinyl Institute identified 22 type A units at 11 facilities, 23 type B 

units at 10 facilities, and 17 type C units at 10 facilities. Test data were gathered from 5 of the 22 

type A units (3 facilities representing 7% of total U.S. and Canadian EDC/VCM/PVC production 

in 1995), 14 of the 23 type B units (8 facilities), and 13 of the 17 type C units (7 facilities). The 

sampled type B and C units represented 70% of total U.S. and Canadian EDC/VCM/PVC 

production in 1995. 

Annual I-TEQDF emission estimates were generated by combining estimated emissions 

from tested units (based on measured stack gas results and plant-specific activity data) with an 

estimate of emissions from untested units. The emissions from the untested units were estimated 

by multiplying the average emission factor for the tested units in the category (the most likely 

8-37
 

estimate) or the average emission factor of the tested units with the highest emissions in each 

class (the upper-bound estimate) by the activity level for the untested units. It is not possible to 

calculate emission factors for TEQDF-WHO98 using the data presented in The Vinyl Institute 

report. 

The Vinyl Institute estimates of most likely and upper-bound emissions during 1995 for 

these three categories are as follows: 

Most likely emissions Upper-bound emissions 

Category estimate (g I-TEQDF/yr) estimate (g I-TEQDF/yr) 

PVC-only incinerators 0.0014 0.0019 

EDC/VCM liquid and liquid/vents 3.7 7.2 

EDC/VCM vents for VCM only 6.9 21.6 

The study also estimated emissions that may have resulted from incineration of 

EDC/VCM/PVC wastes processed by off-site, third-party processing. Using the emission factors 

for liquid and liquid/vents developed in its study, it was estimated that potential emissions to air 

from this source category would be 0.65 g I-TEQDF/yr (most-likely estimate) and 2.3 g ITEQDF/yr (upper-bound estimate). Combining these third-party release estimates with those 

developed above yields a 1995 estimate of 11.2 g I-TEQDF/yr. 

Data validation studies by the CCC indicate that eight EDC/VCM/PVC manufacturing 

facilities released 5.51 g TEQDF-WHO98 (5.56 g I-TEQDF) to air (CCC, 2004); more than 85% of 

the releases occurred at two facilities. Congener-specific and TEQ release estimates to air from 

EDC/VCM/PVC manufacturing facilities are shown in Table 8-19. The emission estimates for 

1995 and 2000 are assigned a high confidence rating because they were based on emissions 

testing of on-site incinerator and vent releases. 

8.3.4.4. Transfers to Secure Landfills 

The CCC reported on the amount of CDDs/CDFs contained in wastewater treatment plant 

sludges used to secure landfills in 2000 (CCC, 2004). These data were determined on the basis 

of sampling wastewater sludges at 16 integrated EDC/VCM/PVC manufacturing facilities. Table 

8-20 summarizes the estimated CDD/CDF congener-specific and TEQ amounts (based on actual 

test data) transferred from specific facilities in 2000. It should be noted that, because the wastes 

were transferred to secure landfills, this is not considered to have been an environmental release; 

therefore, these transfers were not incorporated into the inventory. 

8-38
 

Table 8-19. Congener-specific and TEQ releases to air (g/yr) from EDC/VCM/PVC integrated chem

manufacturing facilities in reference years 2000 and 1995 

Company Facility 

2,3,7,8 TCDD

1,2,3,7,8 PeCDD

1,2,3,4,7,8 HxCDD

1,2,3,6,7,8 HxCDD

1,2,3,7,8,9 HxCDD

1,2,3,4,6,7,8 HpCDD

OCDD

2,3,7,8 TCDF

1,2,3,7,8 PeCDF

2,3,4,7,8 PeCDF

1,2,3,4,7,8 HxCDF

1,2,3,6,7,8 HxCDF

1,2,3,7,8,9 HxCDF

2,3,4,6,7,8 HxCDF

1234678 HpCDF

Dowa Freeport, TX 0.01 0.20 1.01 0.00 0.39 3.73 9.46 1.26 1.80 1.28 11.98 0.00 0.34 2.50 32

Dow Midland, MI 0.026 0.004 0.008 0.000 0.003 0.02 0.086 0.023 0.009 0.008 0.066 0.000 0.002 0.009 0.1Dow Plaquemine, LA 0.00 0.00 0.03 0.00 0.01 0.10 0.27 0.03 0.04 0.03 0.40 0.00 0.00 0.05 1

DuPontb DeLisle, MS 0.000 0.001 0.000 0.000 0.000 0.004 0.011 0.004 0.004 0.006 0.007 0.005 0.002 0.013 0.0

DuPont Edge Moor, DE 0.000 0.001 0.000 0.000 0.000 0.003 0.016 0.001 0.002 0.002 0.003 0.002 0.005 0.003 0.0

DuPont Johnsonville, TN 0.000 0.001 0.000 0.000 0.000 0.003 0.007 0.005 0.006 0.008 0.010 0.008 0.004 0.023 0.0

Georgia Gulf c Plaquemine, LA 0.00 0.00 0.00 0.00 0.00 0.02 0.12 0.09 0.07 0.03 0.17 0.07 0.04 0.01 0

Occidentald Convent, LA 0.00 0.01 0.00 0.01 0.00 0.01 0.00 0.02 0.02 0.01 0.01 0.01 0.00 0.00 0

Occidental Deer Park, TX 0.00 0.01 0.05 0.03 0.06 1.37 13.54 0.06 0.12 0.14 1.38 0.58 0.55 0.06 11

Occidental Ingleside, TX 0.01 0.03 0.11 0.12 0.09 1.58 9.51 0.04 0.35 0.37 3.58 3.38 0.77 1.94 25

Occidental LaPorte, TX 0.00 0.01 0.01 0.02 0.01 0.15 0.48 0.01 0.02 0.03 0.04 0.04 0.01 0.04 0

TOTAL 0.05 0.27 1.21 0.18 0.56 6.99 33.49 1.54 2.45 1.92 17.65 4.09 1.74 4.64 72

8-39
 



 


 


 


 


 
 

aDow Chemical Company.bDuPont. cGeorgia Gulf Corporation.dOccidental Chemical Corporation.Source: Chlorine Chemistry Council (2004).EDC = Ethylene dichloridePVC = Polyvinyl chloride


 


 

VCM = Vinyl chloride monomer
Table 8-20. Congener-specific and TEQ transfers to secure landfills (g/yr)a from EDC/VCM/PVC i

chemical production facilities in 2000 

8-40
 


Company Facility 

2,3,7,8 TCDD

1,2,3,7,8 PeCDD

1,2,3,4,7,8 HxCDD

1,2,3,6,7,8 HxCDD

1,2,3,7,8,9 HxCDD

1,2,3,4,6,7,8 HpCDD

OCDD

2,3,7,8 TCDF

1,2,3,7,8 PeCDF

2,3,4,7,8 PeCDF

1,2,3,4,7,8 HxCDF

1,2,3,6,7,8 HxCDF

1,2,3,7,8,9 HxCDF

2,3,4,6,7,8 HxCDF

Dowb Freeport, TX 0.00 0.00 0.57 0.00 0.35 15.61 74.61 2.63 3.57 1.68 26.28 0.00 0.00 5.00 

Dow Midland, MI 0.00 0.00 0.02 0.00 0.00 0.54 3.63 0.02 0.00 0.01 0.06 0.00 0.00 0.04 

Dow Plaquemine, LA 0.06 0.12 0.87 0.00 0.33 18.72 285.79 4.78 3.77 1.97 21.42 0.00 0.98 3.14 DuPontc DeLisle, MS 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 DuPont Edge Moor, DE 0.00 0.00 0.00 0.00 0.00 0.00 0.10 0.00 0.00 0.00 0.08 0.00 0.00 0.00 

DuPont Johnsonville, TN 0.00 0.00 0.00 0.00 0.00 0.00 1.41 0.00 0.00 0.00 0.44 0.00 0.00 0.00 

Georgia Gulf c Plaquemine, LA 0.00 0.00 0.00 0.00 0.00 0.03 0.12 0.01 0.01 0.02 0.02 0.01 0.01 0.01 

Occidentale Convent, LA 0.00 0.00 0.00 0.00 0.00 0.00 0.15 0.00 0.00 0.00 0.02 0.00 0.00 0.00 

Occidental Deer Park, TX 0.00 0.00 0.00 0.00 0.00 0.08 1.19 0.00 0.00 0.00 0.04 0.03 0.03 0.01 Occidental Ingleside, TX 0.02 0.09 0.08 0.09 0.08 0.09 0.18 0.02 0.09 0.09 0.09 0.10 0.10 0.10 Occidental LaPorte, TX 0.00 0.00 0.00 0.00 0.00 0.00 0.05 0.00 0.00 0.00 0.04 0.00 0.00 0.00 

Occidental Mobile, AL 0.00 0.00 0.00 0.00 0.00 0.00 0.00 3e!04 0.00 0.00 0.00 0.00 0.00 0.00 

Occidental Battleground, TX 0.00 0.00 0.00 0.00 0.00 0.00 0.48 0.00 0.00 0.00 0.00 0.00 0.00 0.00 Occidental Delaware City, DE 0.00 0.00 0.00 0.00 0.00 0.00 0.00 1e!03 0.00 0.00 0.00 0.00 0.00 0.00 Occidental Hahnville, LA 0.00 0.00 0.00 0.00 0.00 0.01 0.10 0.19 0.90 0.85 2.96 1.18 0.63 0.60 

Occidental Muscle Shoals, AL 0.00 0.00 0.00 0.00 0.00 0.00 1e!09 4e!08 1e!07 8e!08 2e!07 1e!07 2e!08 3e!08 

TOTAL 0.08 0.22 1.54 0.09 0.76 35.09 367.81 7.65 8.33 4.62 51.44 1.32 1.75 8.90 

aNot considered to be environmental releases. EDC = Ethylene dichloride bDow Chemical Company. PVC = Polyvinyl chloride cDuPont. VCM = Vinyl chloride monomer dGeorgia Gulf Corporation. eOccidental Chemical Corporation. 

Source: CCC (2004). 
8.3.4.5 Products 

The Vinyl Institute (1998) presented results for 22 samples from 14 of the 24 U.S. and 

Canadian facilities manufacturing suspension and mass PVC resins (13 pipe resins, 3 bottle 

resins, and 6 packaging resins). The results for U.S. manufacturers are summarized in Table 

8-21. The 14 sampled sites represented approximately 74% of estimated 1995 U.S. and 

Canadian suspension and mass PVC resin production. CDDs/CDFs were detected in only one 

sample (0.043 ng I-TEQDF/kg, assuming nondetects equal to zero). The overall mean TEQ 

concentrations were 0.002 ng I-TEQDF/kg (assuming nondetects equal to zero) and 0.7 ng 

I-TEQDF/kg (assuming nondetects equal to one-half the DL). The DLs were 2 ng/kg or less for 

all congeners in all samples except for OCDD and OCDF, which had DLs of 6 ng/kg or less. 

The same study also presented results for six samples from four of the seven U.S. 

facilities manufacturing dispersion PVC resins. CDDs/CDFs were detected in five of the 

samples. The results are summarized in Table 8-21. In terms of production, the four sampled 

sites represent approximately 61% of estimated 1995 U.S. dispersion PVC resin production. The 

results ranged from not detected to 0.008 ng I-TEQDF/kg (overall mean = 0.001 ng I-TEQDF/kg, 

assuming nondetects equal to zero, and 0.4 ng I-TEQDF/kg, assuming nondetects equal to onehalf the DL). The DLs were 2 ng/kg or less for all congeners in all samples except for OCDD 

and OCDF, which had DLs of 4 ng/kg or less. 

Results were also presented for five samples from 5 of the 15 U.S. facilities 

manufacturing EDC. The results are summarized in Table 8-21. In terms of production, the five 

sampled sites represented approximately 71% of the estimated EDC produced in the United 

States in 1995. CDDs/CDFs were detected in only one sample (0.03 ng I-TEQDF/kg). The 

overall mean TEQ concentrations were 0.006 ng I-TEQDF/kg (nondetects equal to zero) and 0.21 

ng I-TEQDF/kg (nondetects equal to one-half the DL). The DLs for all congeners were 1 ng/kg or 

less. 

Using 1995 U.S. production data, 4.846 million metric tons of suspension and mass PVC, 

0.367 million metric tons of dispersion PVC resins, and 1.362 million metric tons of EDC were 

produced. Based on the average TEQ concentration observed, The Vinyl Institute estimated that 

the total I-TEQDF contents of suspension/mass PVC resins, dispersion PVC resins, and EDC were 

0.01 g, 0.004 g, and 0.008 g, respectively (nondetects equal to zero), and 3.39 g, 0.15 g, and 0.29 

g, respectively (nondetects equal to one-half the DL). Therefore, total I-TEQDF present in PVC in 

1995 was estimated to be between 0.02 g (nondetects equal to zero) and 3.83 g (nondetects equal 

to one-half the DL). It is not possible using the data presented in The Vinyl Institute report to 

calculate emission factors for TEQDF-WHO98. However, because neither 1,2,3,7,8-PeCDD nor 

OCDD was detected in any sample, the TEQDF-WHO98 emission factors would be very similar to 

the I-TEQDF emission factors. 

8-41
 
Table 8-21. CDD/CDF concentrations in products from U.S. EDC/VCM/PVC manufacturers 

8-42
 

Congener/congener group 

Suspension and mass PVC resins Dispersion PVC resins EDC 

No. detects/ 

samplesa 

Rangeb (ng/kg) No. of 

detects/ 

samples 

Rangec (ng/kg) 

No. detects/ 

samples 

Min. Max. Min. Max. 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0/22 

0/22 

0/22 

0/22 

0/22 

1/22 

0/22 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

0.64 

ND 

0/6 

0/6 

0/6 

0/6 

0/6 

1/6 

0/6 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

0.8 

ND 

0/5 

0/5 

0/5 

0/5 

0/5 

0/5 

0/5 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0/22 

0/22 

0/22 

0/22 

0/22 

0/22 

1/22 

0/22 

0/22 

0/22 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

0.37 

ND 

ND 

ND 

0/6 

0/6 

0/6 

0/6 

0/6 

0/6 

0/6 

0/6 

0/6 

2/6 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

0.38 

0/5 

0/5 

0/5 

0/5 

0/5 

0/5 

0/5 

1/5 

1/5 

1/5 

Mean I-TEQDF (nondetect = 0) 

Mean I-TEQDF (nondetect = 1⁄2 DL) 

0.002 

0.7 

0.001 

0.4 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

0/22 

0/22 

0/22 

1/22 

0/22 

0/22 

0/22 

1/22 

0/22 

0/22 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

0.64 

ND 

ND 

ND 

0.37 

ND 

ND 

1/6 

1/6 

5/6 

1/6 

0/6 

0/6 

1/6 

0/6 

0/6 

2/6 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

0.24 

0.32 

0.97 

1.3 

ND 

ND 

0.3 

ND 

ND 

0.38 

0/5 

0/5 

0/5 

0/5 

0/5 

0/5 

0/5 

0/5 

1/5 

1/5 
Table 8-21. CDD/CDF concentrations in products from U.S. EDC/VCM/PVC manufacturers (continued) 

aTwo of these 22 samples were duplicate samples from two sites. The results were averaged and treated as one sample for each site. bDetection limits (DLs) for individual samples were less than 2 ng/kg for all congeners and congener groups except OCDD and OCDF, 

 DLs less than 6 ng/kg. 

cDLs for individual samples were less than 2 ng/kg for all congeners and congener groups except OCDD and OCDF, which had DLs lesd“Sales” EDC is defined as EDC sold commercially for non-VCM uses or exported from the United States. eDLs were less than 1 ng/kg for all congeners in all samples. 

DL = Detection limit EDC = Ethylene dichloride ND = Not detected PVC = Polyvinyl chloride VCM = Vinyl chloride monomer Source: The Vinyl Institute (1998). 

8-43
 

Approximately 6.55 million metric tons of PVC and 9.91 million metric tons of EDC 

were produced in North America in 2000 and approximately 5.58 million metric tons of PVC and 

7.83 million metric tons of EDC were produced in 1995 (C&EN, 2002). Approximately 94% of 

PVC production and approximately 17% of EDC production in 1995 occurred in the United 

States. Of the PVC produced, 87% was for suspension and mass PVC products and 7% was for 

dispersion PVC resins. Assuming these product percentages remained the same for 2000, it is 

estimated that approximately 5.69 million metric tons of suspension and mass PVC and 0.46 

million metric tons of dispersion PVC resins were produced and 1.69 million metric tons of EDC 

product were produced. Applying the same average TEQ observed in The Vinyl Institute 

samples from 1998, EPA estimated the total I-TEQDF contents of suspension/mass PVC resins, 

dispersion PVC resins, and EDC produced in 2000 to be 0.01 g, 0.0004 g, and 0.01 g, 

respectively (nondetects equal to zero) and 3.99 g, 0.17 g, and 0.36 g, respectively (nondetects 

equal to one-half the DL). Therefore, total I-TEQDF present in PVC in 2000 was estimated to be 

between 0.02 g (nondetects equal to zero) and 4.52 g (nondetects equal to one-half the DL). 

8.3.5. Other Aliphatic Chlorine Compounds 

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, 1991b). These compounds are produced in large quantities. In 1992, 14.6 million metric 

tons of halogenated hydrocarbons were produced in the United States (U.S. ITC, 1946–1994), 

with 1,2-dichloroethane and vinyl chloride accounting for 82% of total production. Highly 

chlorinated CDDs/CDFs (hexa- to octa-chlorinated congeners) have been found in nanogradequality samples of 1,2-dichloroethane (55 ng/kg of OCDF in one of five samples), 

tetrachloroethene (47 ng/kg of OCDD in one of four samples), epichlorohydrin (88 ng/kg of 

CDDs and 33 ng/kg of CDFs in one of three samples), and hexachlorobutadiene (360 to 425 

ng/kg of OCDF in two samples) obtained in Germany from Promochem (Hutzinger and Fiedler, 

1991b; Heindl and Hutzinger, 1987). No CDDs/CDFs were detected in two samples of allyl 

chloride, three samples of 1,1,1-trichloroethane, and four samples of trichloroethylene (DL 

ranged from 5 to 20 ng/kg) (Heindl and Hutzinger, 1987). Because no more recent or additional 

data could be found in the literature to confirm these values for products manufactured or used in 

the United States, no national estimates of CDD/CDF emissions were made for the inventory. 

EPA’s Office of Water promulgated effluent limitations for facilities that manufacture 

chlorinated aliphatic chlorine compounds and discharge treated wastewater (40 CFR 414.70). 

These effluent limitations do not specifically address CDDs/CDFs. Regulated limits for 

chlorinated aliphatic compounds are 68 μg/L for 1,2-dichloroethane and 22 μg/L for 

tetrachloroethylene. Similarly, OSW promulgated restrictions on land disposal of wastes 

8-44
 

generated during the manufacture of many chlorinated aliphatics (40 CFR 268); however, these 

restrictions do not specifically regulate CDDs/CDFs. 

8.3.6. Dyes, Pigments, and Printing Inks 

Several researchers have analyzed various dyes, pigments, and printing inks obtained in 

Canada and Germany for the presence of CDDs/CDFs (Williams et al., 1992; Hutzinger and 

Fiedler, 1991b; Santl et al., 1994). The following subsections discuss the findings of those 

studies. 

8.3.6.1. Dioxazine Dyes and Pigments 

Williams et al. (1992) analyzed the CDD/CDF content in dioxazine dyes and pigments 

available in Canada. As shown in Table 8-22, OCDD and OCDF concentrations in the 

nanogram-per-kilogram range and HpCDD, HxCDD, and PeCDD concentrations in the 

microgram-per-kilogram range were found in Direct Blue 106 dye (three samples), Direct Blue 

108 dye (one sample), and Violet 23 pigments (six samples). These dioxazine pigments are 

derived from chloranil, which has been found to contain high levels of CDDs/CDFs and has been 

suggested as the source of contamination among these dyes (Christmann et al., 1989b; Williams 

et al., 1992; U.S. EPA, 1992d). In May 1990, EPA received test results showing that chloranil 

was heavily contaminated with dioxins; levels as high as 2,903 μg TEQDF-WHO98/kg (3,065 μg ITEQDF/kg) were measured in samples from four importers (mean value of 1,388 μg TEQDFWHO98/kg [1,754 μg I-TEQDF/kg]) (U.S. EPA, 1992d; Remmers et al., 1992). (See Section 8.3.6 

for analytical results.) 

In the early 1990s, EPA learned that I-TEQDF levels in chloranil could be reduced by 

more than two orders of magnitude (to less than 20 μg/kg) through manufacturing feedstock and 

process changes. EPA’s Office of Pollution Prevention and Toxics subsequently began efforts to 

complete an industry-wide switch from the use of contaminated chloranil to low-dioxin chloranil. 

Although chloranil is not manufactured in the United States, significant quantities are imported. 

As of May 1992, EPA had negotiated agreements with all chloranil importers and domestic 

dye/pigment manufacturers known to EPA that used chloranil in their products to switch to lowdioxin chloranil. In May 1993, when U.S. stocks of chloranil with high levels of CDDs/CDFs 

had been depleted, EPA proposed a SNUR under Section 5 of TSCA that would require industry 

to notify EPA at least 90 days prior to the manufacture, import, or processing, for any use, of 

chloranil containing CDDs/CDFs at a concentration greater than 20 μg I-TEQDF/kg (Federal 

Register, 1993a; U.S. EPA, 1993d). 

8-45
 
Table 8-22. CDD/CDF concentrations (μg/kg) in samples of dioxazine dyes and pigments (Canada) 

8-46
 


Congener/congener 

group Blue 106 Blue 108 Violet 23 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

ND (0.3) 

ND (0.3) 

-­

ND (0.3) 

ND (0.3)

-­

ND (0.3) 

ND (0.3) 

-­

ND (0.3) 

ND (0.3) 

-­

ND (0.3) 

ND (0.3) 

-­

ND (0.3) 

ND (0.3) 

-­

ND (0.3) 

ND (0.3) 

-­

ND (0.3

ND (0.3

-­

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

ND (0.3) 

ND (0.3) 

31 

ND (0.3) 

ND (0.3) 6 

ND (0.3) 

ND (0.3) 

9 

ND (0.3) 

ND (0.3) ND (0.3) 

ND (0.3) 

ND (0.3) 

9 

ND (0.3) 

ND (0.3) 

1 

ND (0.3) 

ND (0.3) 

16 

ND (0.3

ND (0.3

10 

OCDD 41,953 28,523 18,066 23 7,180 806 11,022 7,929 

2,3,7,8-TCDF ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3

1,2,3,7,8-PeCDF ND (0.3) ND (0.3) ND (0.3) ND (0.3) 0.5 ND (0.3) ND (0.3) ND (0.3

2,3,4,7,8-PeCDF ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3

1,2,3,4,7,8-HxCDF 12 2 2 ND (0.3) 76 4 39 31 

1,2,3,6,7,8-HxCDFa 

1,2,3,7,8,9-HxCDF -­ -­ -­ -­ -­ -­ -­ -­ 

2,3,4,6,7,8-HxCDF ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3

1,2,3,4,6,7,8-HpCDF 50 10 14 9 13 10 11 4 

1,2,3,4,7,8,9-HpCDF -­ -­ -­ -­ -­ -­ -­ -­ 

OCDF 12,463 1,447 1,006 11 941 125 3,749 1,556 Total 2,3,7,8-CDD 41,984 28,529 18,075 23 7,189 807 11,038 7,939 

Total 2,3,7,8-CDF 12,525 1,459 1,022 20 1,030.5 139 3,799 1,591 

Total I-TEQDF b 56.4 30.3 19.5 0.1 16 1.4 18.9 12.7 

Total TEQDF-WHO98 b 7.45 3.4 2.3 0.1 8.7 0.6 5.6 4.2 

Total TCDD ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3

Total PeCDD ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3

Total HxCDD ND (0.3) ND (0.3) ND (0.3) 1 21 2 7 ND (0.3

Total HpCDD 34 8 12 ND (0.3) 30 5 36 11 

Total OCDD 41,953 28,523 18,066 23 7,180 806 11,022 7,929 

Total TCDF ND (0.3) 0.3 ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3) ND (0.3

Total PeCDF ND (0.3) ND (0.3) ND (0.3) ND (0.3) 0.5 ND (0.3) ND (0.3) ND (0.3

Total HxCDF 12 2 2 ND (0.3) 76 5 39 31 

Total HpCDF 71 32 26 12 26 14 29 13 

Total OCDF 12,463 1,447 1,006 11 941 125 3,749 1,556 

Total CDD/CDF b 54,533 30,012.3 19,112 47 8,274.5 957 14,882 9,540 
Table 8-22. CDD/CDF concentrations in samples of dioxazine dyes and pigments (μg/kg) (Canada) 

aResults listed for 1,2,3,4,7,8-HxCDF include concentrations for 1,2,3,6,7,8-HxCDF. bCalculations assume nondetected values were equal to zero. 

ND = Not detected (value in parenthesis is the detection limit) -- = Not reported Source: Williams et al. (1992). 

8-47
 

In 1983, approximately 36,500 kg of chloranil were imported (U.S. ITC, 1984). The U.S. 

International Trade Commission has not published quantitative import data for chloranil since 

1984. If it is assumed that this import volume reflects actual usage of chloranil in the United 

States during 1987 and that the CDD/CDF contamination level was 1,388 μg TEQDF-WHO98/kg 

(1,754 μg I-TEQDF/kg), then the maximum release to the environment via processing wastes and 

finished products was 50.6 g TEQDF-WHO98 (64 g I-TEQDF). If it is assumed that the import 

volume in 1995 was also 36,500 kg but that the imported chloranil contained 10 μg I-TEQDF/kg 

on average, then the total potential annual TEQ release associated with chloranil in 1995 was 

50.6 g TEQDF-WHO98 (64 g I-TEQDF). 

In 1986, EPA promulgated the Inventory Update Rule (IUR) that requires the partial 

updating of the TSCA Chemical Inventory database. Every four years, chemical manufacturers 

and importers of chemicals listed in the TSCA inventory that produce at one plant site or import 

at production volume levels of 10,000 lb or more must report the range of chemical production or 

import. According to information entered in the TSCA database, 10,000 to 500,000 lb (4,540 to 

227,000 kg) of chloranil were imported in 1994 and 2000 (http://www.epa.gov/opptintr/iur 

/iur02/search03.htm). Assuming the imported chloranil contained the same concentration of 

dioxin as the 1995 estimate (10 μg I-TEQDF/kg), the total potential annual TEQ release associated 

with chloranil in 2000 was 0.05 to 2.27 g I-TEQDF-WHO98 (mean of 1.16 g I-TEQDF). 

8.3.6.2. Phthalocyanine Dyes and Printing Inks 

Hutzinger and Fiedler (1991b) found CDDs/CDFs (tetra-, penta-, and hexa-chlorinated 

congeners) in the microgram-per-kilogram range in a sample of a Ni-phthalocyanine dye. No 

CDDs/CDFs were detected (DL of 0.1 to 0.5 μg/kg) in two samples of Cu-phthalocyanine dyes 

and in one Co-phthalocyanine dye (Hutzinger and Fiedler, 1991b). 

Santl et al. (1994) reported the results of analyses of four printing inks obtained from a 

supplier in Germany. Two of the inks are used for rotogravure printing and two are used for 

offset printing. The results of the analyses are presented in Table 8-23. The TEQDF-WHO98 

content of the inks ranged from 17.7 to 87.2 ng/kg (15 to 88.6 ng/kg on an I-TEQDF basis). 

Primarily non-2,3,7,8-substituted congeners were found. The identities of the dyes and pigments 

in these inks were not reported. 

Although EPA provided an estimate of potential environmental releases based on limited 

information of contaminant levels of CDDs/CDFs in the product, the estimate is still too 

uncertain to include in the quantitative inventory of sources. It is currently not known whether 

the dioxin contamination in the product actually results in a release to the open and circulating 

environment. 

8-48
 
Table 8-23. CDD/CDF concentrations (ng/kg) in printing inks (Germany) 

Congener/congener group 

Rotogravure 

(2-color) 

Rotogravure 

(4-color) 

Offset 

(4-color) 

Offset 

(4-color) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND (1) 

8 

19 

325 

155 

2,770 

5,810 

ND (1.5) 

ND (4) 

ND (5) 

310 

105 

1,630 

2,350 

ND (2) 

15 

16 

82 

42 

540 

890 

ND (2) 

6 

11 

21 

14 

240 

230 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

2.5 

ND (2) 

ND (2) 

4 

ND (3) 

ND (3) 

ND (3) 

40 

ND (4) 

129 

14 

ND (4) 

ND (4) 

7 

ND (5) 

ND (5) 

ND (5) 

14 

ND (7) 

ND (10) 

7 

ND (4) 

ND (4) 

27 

ND (5) 

ND (5) 

ND (5) 

315 

11 

960 

7 

ND (3) 

ND (3) 

35 

ND (5) 

ND (5) 

ND (5) 

42 

ND (6) 

165 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF a 

Total TEQDF-WHO98 

9,087 

175.5 

88.6 

87.2 

4,395 

35 

62.4 

60.3 

1,585 

1,320 

35.4 

41.2 

522 

249 

15 

18 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

4 

58 

2,679 

5,630 

5,810 

5.5 

13 

29 

64 

129 

ND (2) 

145 

2,485 

3,460 

2,350 

28 

ND (4) 

45 

14 

ND (10) 

77 

35 

660 

1,100 

890 

90 

340 

95 

566 

960 

38 

25 

246 

445 

230 

35 

110 

94 

63 

165 

Total CDD/CDF 14,421.5 8,527 4,813 1,451 a

Calculations assume nondetect values were zero.
 


ND = Not detected (value in parenthesis is the detection limit)
 


Source: Santl et al. (1994).
 


8-49
 

8.3.7. TSCA Dioxin/Furan Test Rule 

Citing evidence that halogenated dioxins and furans may be formed as by-products during 

chemical manufacturing processes (Versar, Inc., 1985), EPA issued a rule under Section 4 of 

TSCA that requires chemical manufacturers and importers to test for the presence of 

CDDs/CDFs and brominated dibenzo-p-dioxins (BDDs) and brominated dibenzofurans (BDFs) 

in certain commercial organic chemicals (Federal Register, 1987a). The rule listed 12 

manufactured or imported chemicals that required testing and 20 chemicals not currently 

manufactured or imported that would require testing if manufacture or importation resumed. 

These chemicals are listed in Table 8-24. The specific dioxin and furan congeners that require 

quantitation and the target limits of quantitation (LOQs) that are specified in the rule are listed in 

Table 8-25. 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 28 precursor chemicals listed in Table 8-26. 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. 

Twenty-three sampling and analytical protocols and test data for 10 of the 12 chemicals 

that required testing were submitted to EPA (U.S. EPA, 2003a, c). Manufacture or import of two 

substances (tetrabromobisphenol-A-bis-2,3-dibromopropylether and tetrabromobisphenol-Adiacrylate) have stopped since the test rule was promulgated. (All data and reports in the EPA 

TSCA docket are available for public review and inspection at EPA Headquarters in Washington, 

DC.) 

Table 8-27 presents the results of analytical testing for CDDs/CDFs for the chemicals that 

have data available in the TSCA docket. Five of these 10 chemicals contained CDDs/CDFs. 

Positive results were obtained for 2,3,5,6-tetrachloro-2,5-cyclohexadiene-1,4-dione (chloranil), 

pentabromodiphenyloxide, octabromodiphenyloxide, decabromodiphenyloxide, and 1,2­

bis(tribromophenoxy)-ethane. Table 8-28 presents the quantitative analytical results for four 

submitted chloranil samples, as well as the results of an EPA analysis of a sample of carbazole 

violet, which is manufactured from chloranil. 

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 μg/kg, 12 μg/kg, and 1 μg/kg, respectively. No hexa-, hepta-, or octa-BDFs 

were detected. The investigators also analyzed analytical-grade samples of two other brominated 

flame retardants, pentabromophenol and tetrabromophthalic anhydride; no BDDs or BDFs were 

detected (DLs not reported). 

8-50
 
Table 8-24. Chemicals requiring Toxic Substances Control Act Section 4 

testing under the dioxin/furan rule 

CAS No. Chemical name 

Currently manufactured or imported as of June 5, 1987 

79-94-7 

118-75-2 

118-79-6 

120-83-2 

1163-19-5 

4162-45-2 

21850-44-2 

25327-89-3 

32534-81-9 

32536-52-0 

37853-59-1 

55205-38-4 

Tetrabromobisphenol-A 

2,3,5,6-Tetrachloro-2,5-cyclohexadiene-1,4-dione 

2,4,6-Tribromophenol 

2,4-Dichlorophenol 

Decabromodiphenyloxide 

Tetrabromobisphenol-A-bisethoxylate 

Tetrabromobisphenol-A-bis-2,3-dibromopropylethera 

Allyl ether of tetrabromobisphenol-A 

Pentabromodiphenyloxide 

Octabromodiphenyloxide 

1,2-Bis(tribromophenoxy)-ethane 

Tetrabromobisphenol-A-diacrylatea 

Not manufactured or imported as of June 5, 1987b 

79-95-8 

87-10-5 

87-65-0 

95-77-2 

95-95-4 

99-28-5 

120-36-5 

320-72-9 

488-47-1 

576-24-9 

583-78-8 

608-71-9 

615-58-7 

933-75-5 

1940-42-7 

2577-72-2 

3772-94-9 

37853-61-5 

-

-

Tetrachlorobisphenol-A 

3,4',5-Tribromosalicylanide 

2,6-Dichlorophenol 

3,4-Dichlorophenol 

2,4,5-Trichlorophenol 

2,6-Dibromo-4-nitrophenol 

2[2,4-(Dichlorophenoxy)]-propanoic acid 

3,5-Dichlorosalicyclic acid 

Tetrabromocatechol 

2,3-Dichlorophenol 

2,5-Dichlorophenol 

Pentabromophenol 

2,4-Dibromophenol 

2,3,6-Trichlorophenol 

4-Bromo-2,5-dichlorophenol 

3,5-Dibromosalicylanide 

Pentachlorophenyl laurate 

Bismethylether of tetrabromobisphenol-A 

Alkylamine tetrachlorophenate 

Tetrabromobisphenol-B a

No longer manufactured in or imported into the United States (memorandum dated May 4, 1993, from Gordon 

Cash, U.S. EPA/OPPTS, to John Schaum, U.S. EPA/ORD). b

As of August 5, 1995, neither manufacture nor importation of any of these chemicals had resumed in the United 

 States (memorandum dated August 2, 1995, from T.S. Holderman, U.S. EPA, to A. Adenuga, Versar, Inc.). 

8-51 
Table 8-25. Congeners and limits of quantitation (LOQs) for which 

quantitation is required under the dioxin/furan test rule and pesticide Data 

Call-In 

Chlorinated dioxins 

and furans 

Brominated dioxins 

and furans 

LOQ 

(μg/kg) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

2,3,7,8-TBDD 

1,2,3,7,8-PeBDD 

1,2,3,4,7,8-HxBDD 

1,2,3,6,7,8-HxBDD 

1,2,3,7,8,9-HxBDD 

1,2,3,4,6,7,8-HpBDD 

0.1 

0.5 

2.5 

2.5 

2.5 

100 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

2,3,7,8-TBDF 

1,2,3,7,8-PeBDF 

2,3,4,7,8-PeBDF 

1,2,3,4,7,8-HxBDF 

1,2,3,6,7,8-HxBDF 

1,2,3,7,8,9-HxBDF 

2,3,4,6,7,8-HxBDF 

1,2,3,4,6,7,8-HpBDF 

1,2,3,4,7,8,9-HpBDF 

1 

5 

5 

25 

25 

25 

25 

1,000 

1,000 

8.3.8. Halogenated Pesticides and FIFRA Pesticides Data Call-In 

In the late 1970s and early 1980s, attention began to focus on pesticides as potential 

sources of CDDs/CDFs in the environment. Up to that time, CDD/CDF levels were not 

regulated in end-use pesticide products. However, some of the active ingredients in pesticides, 

particularly chlorinated phenols and their derivatives, were known or suspected to be 

contaminated with CDDs/CDFs. During the 1980s and 1990s, EPA took several actions to 

investigate and control CDD/CDF contamination of pesticides. 

In 1983, EPA cancelled the sale of Silvex and 2,4,5-T for all uses (Federal Register, 

1987a). Earlier, in 1979, EPA had ordered emergency suspension of the forestry, rights-of-way, 

and pasture uses of 2,4,5-T. Emergency suspensions of the forestry, rights-of-way, pasture, home 

and garden, commercial/ornamental turf, and aquatic weed control/ditch bank uses of Silvex 

were also ordered (Federal Register, 1979; Plimmer, 1980). The home and garden, 

commercial/ornamental turf, and aquatic weed control/ditch bank uses of 2,4,5-T had been 

suspended in 1970. 

8-52
 

Table 8-26. Precursor chemicals subject to reporting requirements under 

Toxic Substances Control Act Section 8(a)a 

CAS No. Chemical name 

85-22-3 

87-61-6 

87-84-3 

89-61-2 

89-64-5 

89-69-0 

92-04-6 

97-74-6 

94-81-5 

95-50-1 

95-56-7 

95-57-8 

95-88-5 

95-94-3 

95-50-7 

99-30-9 

99-54-7 

106-46-7 

108-70-3 

108-86-1 

108-90-7 

117-18-0 

120-82-1 

348-51-6 

350-30-1 

615-67-8 

626-39-1 

827-94-1 

Pentabromoethylbenzene 

1,2,3-Trichlorobenzene 

1,2,3,4,5-Pentabromo-6-chlorocyclohexane 

1,4-Dichloro-2-nitrobenzene 

4-Chloro-2-nitrophenol 

2,4,5-Trichloronitrobenzene 

2-Chloro-4-phenylphenol 

4-Chloro-o-toloxy acetic acid 

4-(2-Methyl-4-chlorophenoxy) butyric acid 

o-Dichlorobenzene 

o-Bromophenol 

o-Chlorophenol 

4-Chlororesorcinol 

1,2,4,5-Tetrachlorobenzene 

5-Chloro-2,4-dimethoxyaniline 

2,6-Dichloro-4-nitroaniline 

1,2-Dichloro-4-nitrobenzene 

p-Dichlorobenzene 

1,3,5-Trichlorobenzene 

Bromobenzene 

Chlorobenzene 

1,2,4,5-Tetrachloro-3-nitrobenzene 

1,2,4-Trichlorobenzene 

o-Chlorofluorobenzene 

3-Chloro-4-fluoronitrobenzene 

Chlorohydroquinone 

1,3,5-Tribromobenzene 

2,6-Dibromo-4-nitroaniline a

Dibromobenzene (CAS No. 106-37-6) was identified in the preamble to 52 FR 21412 as one of 29 precursor 

 chemicals; however, it was inadvertently omitted from the regulatory text. Because the regulatory text identified

 only 28 chemicals, 28 chemicals appear in 40 CFR 766.38 and in this table. 

In 1984, EPA issued a notice of intent to cancel registrations of pesticide products 

containing PCP (including its salts) for all wood preservative uses (Federal Register, 1984). This 

notice specified modifications to the terms and conditions of product registrations that were 

required in order to avoid cancellation of the products. In response to this notice, several trade 

associations and registrants requested administrative hearings to challenge EPA’s determinations. 

After carefully considering the comments and alternatives suggested during the prehearing stage 

of the administrative proceedings, EPA concluded that certain changes to the 1984 notice were 

appropriate. These changes, finalized in 1986 (Federal Register, 1986), included the following: 

8-53
 

Table 8-27. Results of analytical testing for dioxins and furans in the 

chemicals tested to date under Section 4 of the dioxin/furan test rule 

CAS No. Chemical name 

No. of 

chemical 

companies 

that 

submitted 

data 

No. of 

positive 

studies 

Congeners detected 

(detection range in μg/kg) 

79-94-7 Tetrabromobisphenol-A 3 0 a 

118-75-2 2,3,5,6-Tetrachloro-2,5­

cyclohexadiene1,4-dione (chloranil) 

6 5 See Table 8-28 

118-79-6 2,4,6-Tribromophenol 1 0 a 

120-83-2 2,4-Dichlorophenol 1 0 a 

1163-19-5 Decabromodiphenyl 

oxide 

3 3 2,3,7,8-PeBDD (ND–0.1)

 1,2,3,4,7,8/1,2,3,6,7,8-HxBDD (ND–0.5)

 1,2,3,7,8,9-HxBDD (ND–0.76)

 1,2,3,7,8-PeBDF (ND–0.7)

 1,2,3,4,7,8/1,2,3,6,7,8-HxBDF (ND–0.8)

 1,2,3,4,6,7,8-HpBDF (17–186) 

25327-89-3 Allyl ether of 

tetrabromobisphenol-A 

1 0 a 

32536-52-0 Octabromodiphenyl 

oxide 

3 3 2,3,7,8-TBDD (ND–0.71)

 1,2,3,7,8-PeBDD (ND–0.1)

 2,3,7,8-TBDF (ND–12.6)

 1,2,3,7,8-PeBDF (ND–6.3)

 2,3,4,7,8-PeBDF (ND–83.1)

 1,2,3,4,7,8/1,2,3,6,7,8-HxBDF (ND–67.8)

 1,2,3,7,8,9-HxBDF (ND–56.0)

 1,2,3,4,6,7,8-HpBDF (ND–330) 

378-53-59-1 1,2-Bis(tribromophenoxy)-ethane 1 1 2,3,7,8-TBDF (ND–0.04)

 1,2,3,4,7,8/1,2,3,6,7,8-HxBDF (ND–0.03)

 1,2,3,4,6,7,8-HpBDF (ND–0.33) 

32534-81-9 Pentabromodiphenyl 

oxide 

3 3 1,2,3,7,8-PeBDD (ND–5.9)

 1,2,3,4,7,8/1,2,3,6,7,8-HxBDD (ND–6.8)

 1,2,3,4,7,8/1,2,3,6,7,8-HxBDD (ND–6.8)

 1,2,3,7,8,9-HxBDD (ND–0.02)

 2,3,7,8-TBDF(ND–3.1)

 1,2,3,7,8-PeBDF (0.7–10.2)

 2,3,4,7,8-PeBDF (0.1–2.9)

 1,2,3,4,7,8/1,2,3,6,7,8-HxBDF (15.6–61.2)

 1,2,3,4,6,7,8-HpBDF (0.7–3.0) 

4162-45-2 Tetrabromobisphenol-Abisethoxylate 1 0 a 

a

No 2,3,7,8-substituted dioxins and furans detected above the test rule target limits of quantitation (see 

 Table 8-20). 

ND = Not detected 

Source: Holderman and Cramer (1995). 

8-54 
Table 8-28. CDD/CDF concentrations (μg/kg) in chloranil and carbazole 

violet samples analyzed pursuant to the EPA dioxin/furan test rule 

Congener 

Chloranil 

Carbazole 

violet 

Importer 

1 

Importer 

2 

Importer 

3 

Importer 

4 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND (1) 

ND (2) 

ND (3) 

ND (3) 

ND (1) 

110 

240,000 

ND (1) 

ND (2) 

ND (10) 

75 

48 

8,200 

180,000 

ND (2) 

ND (5) 

ND (5) 

ND (5) 

ND (5) 

390 

760,000 

ND (2) 

ND (6) 

ND (3) 

6 

9 

2,300 

71,000 

ND (0.8) 

ND (0.5) 

ND (1.2) 

ND (1.2) 

ND (1.2) 

28 

1,600 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

ND (1) 

ND (1) 

ND (1) 

35 

ND (5) 

6 

ND (5) 

33 

ND (15) 

18,000 

ND (2) 

ND (1) 

ND (1) 

ND (860) 

ND (860) 

ND (680) 

ND (680) 

240,000 

ND (100) 

200,000 

ND (1) 

ND (3) 

ND (3) 

ND (4) 

ND (4) 

ND (4) 

ND (4) 

36 

ND (15) 

50,000 

ND (2) 

ND (5) 

ND (5) 

5,600 

ND (600) 

ND (600) 

ND (600) 

230,000 

ND (400) 

110,000 

ND (1.6) 

ND (0.9) 

ND (0.9) 

ND (20) 

ND (20) 

ND (20) 

ND (20) 

15,000 

ND (20) 

59,000 

Total I-TEQDF a 263 2,874 814 3,065 211 

Total TEQDF-WHO98 a 31 2,532 85 2,903 156 

a

Calculated assuming nondetect values are zero. 

ND = Not detected (value in parenthesis is the minimum detection limit) 

Source: Remmers et al. (1992). 

(a) all wood preservative uses of PCP and its salts were classified as “restricted use” only by 

certified applicators, (b) specific worker protection measures were required, (c) limits were 

placed on the HxCDD content of PCP, and (d) label restrictions for home and farm uses of PCP 

prohibited its application indoors and to wood intended for interior use (with a few exceptions) 

as well as its application in a manner that might result in direct exposure of domestic animals or 

livestock or in the contamination of food, feed, or drinking and irrigation water. 

EPA subsequently amended its Notice on the wood preservative uses to establish reliable 

and enforceable methods for implementing certified limits for HxCDD and 2,3,7,8-TCDD in 

registered wood preservative pesticide products (Federal Register, 1987b). Levels of 2,3,7,8­

TCDD were not allowed to exceed 1 ppb in any product, and after February 2, 1989, any 

manufacturing-use PCP released for shipment could not contain HxCDD levels that exceeded an 

8-55
 

average of 2 ppm over a monthly release or a batch level of 4 ppm (a gradually phased-in 

requirement). On January 21, 1987, EPA prohibited the registration of PCP and its salts for most 

nonwood uses (Federal Register, 1987c). EPA deferred action on several uses (uses in 

pulp/paper mills, oil wells, and cooling towers) pending receipt of additional exposure, use, and 

ecological effects data. On January 8, 1993, EPA issued a press advisory stating that its special 

review of these deferred nonwood uses was being terminated because all of these uses had been 

either voluntarily cancelled by the registrants or cancelled by EPA for failure of the registrants to 

pay the required annual maintenance fees (U.S. EPA, 1993e). 

PCP was one of the most widely used biocides in the United States prior to the regulatory 

actions to cancel and restrict certain of its wood and nonwood preservative uses. PCP was 

registered for use as a herbicide, defoliant, mossicide, and mushroom house biocide. It also 

found use as a biocide in pulp-paper mills, oil wells, and cooling towers. These latter three uses 

were terminated on or before 1993 (U.S. EPA, 1993e). However, the major use (greater than 

80% of consumption) of PCP was and continues to be wood preservation. 

The production of PCP for wood preserving began on an experimental basis in the 1930s. 

In 1947, nearly 3,200 metric tons of PCP were reported to have been used in the United States by 

the commercial wood preserving industry. Use in this industry steadily increased through the 

mid-1970s (American Wood Preservers Institute, 1977). Although domestic consumption 

volumes are not available for all years, it is estimated, on the basis of historical production/export 

data for PCP reported in Mannsville (1983), that 90 to 95% of production volume has typically 

been consumed domestically rather than exported. A reasonable estimate of average annual 

domestic PCP consumption during the period of 1970 to 1995 is about 400,000 metric tons. This 

estimate assumes an average annual consumption rate of 20,000 metric tons/yr during the 1970s, 

15,000 metric tons/yr during the 1980s, and 10,000 metric tons/yr during the 1990s. 

Table 8-10 presents a compilation of published data on the CDD/CDF content of 

technical-grade PCP. The only samples that have been analyzed for all dioxin-like CDDs/CDFs 

were manufactured in the mid to late 1980s. Figure 8-4 presents these data in graphical form. It 

is evident from the figures that the predominant congener groups are OCDD, OCDF, HpCDF, 

and HpCDD and the dominant 2,3,7,8-substituted congeners are OCDD, 1,2,3,4,6,7,8-HpCDD, 

and OCDF. Waddell et al. (1995) tested analytical-grade PCP (from Aldrich Chemical Co.) for 

CDD/CDF content and found the same congener profile; however, the CDD/CDF levels were 

three to four orders of magnitude lower. Table 8-11 presents a similar compilation of published 

data on the CDD/CDF content of PCP-Na. The table shows the same patterns of dominant 

congeners and congener groups reported for PCP. 

8-56
 
2,3,7,8-TCDD

1,2,3,7,8-PeCDD

1,2,3,4,7,8-HxCDD1,2,3,6,7,8-HxCDD

1,2,3,7,8,9-HxCDD

1,2,3,4,6,7,8-HpCDD

OCDD

2,3,7,8-TCDF

1,2,3,7,8-PeCDF2,3,4,7,8-PeCDF

1,2,3,4,7,8-HxCDF

1,2,3,6,7,8-HxCDF

1,2,3,7,8,9-HxCDF2,3,4,6,7,8-HxCDF

1,2,3,4,6,7,8-HpCDF1,2,3,4,7,8,9-HpCDF

OCDF

0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

Percent Total CDDs and CDFs 

Congener

Figure 8-4. Congener profile for technical-grade PCP (developed from data 

in last column in Table 8-10). 

Samples of technical-grade PCP manufactured during the mid to late 1980s contained 

about 1.7 mg TEQDF-WHO98/kg (3 mg I-TEQ/kg), based on the data presented in Table 8-10. No 

published reports could be located that present the results of any congener-specific analyses of 

PCP manufactured since the late 1980s. However, monthly measurements of CDD/CDF 

congener group concentrations in technical PCP manufactured for use in the United States have 

been reported to EPA from 1987 to the present (letter dated March 5, 1997, from Thomas 

Mitchell, KMG-Bernuth, to Matthew Lorber, U.S. EPA; letter dated February 7, 1997, from John 

Wilkinson, Pentachlorophenol Task Force, to Matthew Lorber, U.S. EPA; U.S. EPA, 1999a). 

The average congener group concentrations reported to EPA for the years 1988 (i.e., one year 

after EPA regulations were imposed limiting HxCDD and 2,3,7,8-TCDD concentrations in PCP) 

to 1999 are presented in Table 8-10. In general, the average congener group concentrations 

during the period of 1988 to 1999 are lower by factors of 2 to 4 than those observed in the mid to 

late 1980s’ full congener analysis samples. If it is assumed that the toxic CDD/CDF congeners 

have also been reduced by similar factors, then the TEQ content of PCP manufactured since 1988 

is about 0.6 mg TEQDF-WHO98/kg (1 mg I-TEQ/kg). 

8-57
 
An estimated 12,000 metric tons of PCP were used for wood preservation in the United 

States in 1987 (WHO, 1991). An estimated 8,400 metric tons were used in 1994 (American 

Wood Preservers Institute, 1995); for purposes of this report, it is assumed that an identical 

amount was used in 1995. In 1999, approximately 7,710 metric tons of PCP were produced 

annually in the United States (Council of Great Lakes Industries, 1999); for purposes of this 

report, it is assumed that an identical amount was produced in 2000. Assuming that 95% of the 

production volume was consumed domestically (Mannsville, 1983), and that all of the PCP 

produced in 2000 was used for wood preservation, approximately 7,325 metric tons of PCP was 

used in the United States for wood preservation. Combining these activity level estimates with 

the TEQ concentration estimates presented above indicates that 20,000 μg TEQDF-WHO98 

(36,000 μg I-TEQDF), 4,800 μg TEQDF-WHO98 (8,400 μg I-TEQDF), and 4,175 μg TEQDF-WHO98 

(7,325 μg I-TEQDF) were incorporated into PCP-treated wood products in 1987, 1995, and 2000, 

respectively. These amounts in PCP products are not considered an environmental release and 

therefore are not included in the inventory. As discussed below, there is some evidence that 

releases could occur, but no consistent estimation approach could be found. 

Although the estimates of the mass of TEQ in treated wood are fairly certain, no studies 

are available that provide measured CDD/CDF release rate data from which a reliable estimate 

can be made of the amount of CDDs/CDFs that have or will volatilize or leach from treated 

wood. Several recent field studies, discussed in the following paragraphs, demonstrate that 

CDDs/CDFs do apparently leach into soil from PCP-treated wood, but the studies do not provide 

release rate data. No studies were located that provide any measured CDD/CDF volatilization 

rates from PCP-treated wood. Although CDDs/CDFs have very low vapor pressures, they are 

not bound to, nor do they react with, the wood in any way that would preclude volatilization. 

Several studies, discussed below, have attempted to estimate potential CDD/CDF volatilization 

releases using conservative assumptions or modeling approaches, but these estimates span many 

orders of magnitude. 

Gurprasad et al. (1995) analyzed three PCP-treated utility poles and their surrounding 

surface soils for penta- through octa-CDD content. All three poles showed significant levels of 

HxCDD (0.29 to 0.47 mg/kg), HpCDD (4.69 to 6.63 mg/kg), and OCDD (27.9 to 42.1 mg/kg), 

but no PeCDD. Surface soils collected 2 cm from the poles also had detectable levels of 

HxCDD, HpCDD, and OCDD; however, no consistent pattern was found between the CDD 

concentrations in the poles and those in the adjacent soils. The soil concentrations did, however, 

show the same relative congener group pattern observed in the wood. CDD concentrations in 

soils obtained 20 cm from the poles were an order of magnitude less than those measured at 2 

cm. Soils 26 m from the poles showed nondetect values or values close to the DL of 0.01 to 0.02 

mg/kg. 

8-58
 
In a study of the leaching of PCP from 31 utility poles, the Electric Power Research 

Institute (EPRI, 1995) found patterns of PCP distribution in soils surrounding poles similar to 

those found by Gurprasad et al. (1995) for CDDs. PCP concentrations decreased by as much as 

two orders of magnitude between 7.5 cm from the poles and 20 cm from the poles, with an 

average decrease of slightly more than one order of magnitude over this distance. The EPRI 

study also found no obvious trend between PCP concentration in the wood (eight poles analyzed) 

and the age of the poles (4 to 11 years) or the PCP concentration in the surface soil. On the basis 

of their results and those of the EPRI study, Gurprasad et al. concluded that CDDs probably leach 

from PCP-treated utility poles with the PCP/oil carrier and travel in the soil in a similar manner. 

Wan (1995) and Wan and Van Oostdam (1995) measured CDD/CDF concentrations in 

waters and sediments from ditches surrounding utility poles and railroad ties and demonstrated 

that chlorophenol-treated wood could serve as a source of CDD/CDFs to the aquatic 

environment. Ten samples were collected at each of six utility pole sites and five railroad tie 

sites 1 to 2 days after major rainfall events and then were composited into one sample per site 

prior to analyses. Total CDDs (mean value of 76.7 mg/kg) and total CDFs (mean value of 18.7 

mg/kg) detected in chlorophenol/creosote-treated utility poles were about six to eight times 

greater, respectively, than the CDD and CDF concentrations detected in chlorophenol/creosotetreated railroad ties. 

Total CDDs found in water from railway ditches without utility poles (i.e., only treated 

railroad ties were present) were approximately 20 times higher than the background level found 

in farm ditch water. Total CDDs in railway ditches with utility poles were 4,300 times higher 

than the background levels. Water from railway ditches without utility poles contained total 

CDF levels 13 times higher than background levels, whereas water in ditches adjacent to poles 

had levels 8,500 times higher than background levels. Total CDDs in ditch sediments adjacent to 

and 4 m downstream of utility poles were about 5,900 and 2,200 times higher, respectively, than 

background levels; total CDFs for the same sites were about 8,100 and 1,700 times higher, 

respectively, than background levels. Total CDDs found in ditch sediments of railway and ditch 

sediments adjacent to utility poles were about 5 and 700 times higher, respectively, than 

background levels, and total CDFs were about 9 and 1,800 times higher, respectively, than 

background levels. Both CDDs and CDFs were found in utility ditch sediments 4 m downstream 

of treated power poles, but at levels 200 and 400 times lower, respectively, than those found 

adjacent to poles, indicating that they were transported from point sources of contamination. The 

corresponding values for CDFs were 5,400 and 8,000 times higher, respectively, in 

concentration. 

Bremmer et al. (1994) estimated an annual release of 15 to 125 g I-TEQDF from PCPtreated wood in the Netherlands. The lower estimate was based on three basic assumptions: 

8-59
 

(1) the half-life of PCP in treated wood is 15 years (according to industry sources), (2) the halflife of CDDs/CDFs in treated wood is 10 times that of PCP (i.e., 150 years) because of the lower 

vapor pressures of CDDs/CDFs relative to PCP, and (3) the typical CDD/CDF concentration in 

PCP has been 3,000 μg/kg. The higher estimate was based on an assumed half-life of PCP in 

wood of 15 years and the results of an indoor air study by Papke et al. (1989) conducted at 

several kindergartens where PCP-treated wood had been used. Although Papke et al. found no 

clear correlation between indoor air concentrations of CDDs/CDFs and PCP across the range of 

CDD/CDF concentrations observed in the 20-plus samples (2.6 to 427 pg CDD/CDF/m3), there 

did appear to be a positive correlation at the sites with more elevated CDD/CDF concentrations. 

Bremmer et al. (1994) reported the average ratio of PCP to I-TEQ DF air concentrations at these 

elevated sites to be 1:5 × 10-6 (or about the same ratio as the concentration of I-TEQ DF in 

technical PCP). The results of the Papke et al. (1989) study imply that CDDs/CDFs may be 

released from PCP-treated wood at the same rate as is PCP rather than at a rate 10 times slower. 

Rappe (1995) used the emission factor approach developed by Bremmer et al. (1994) and 

an assumed U.S. usage volume of PCP over the past 50 years (0.5 million metric tons) to 

estimate that as much as 10.5 kg I-TEQDF could volatilize from PCP-treated wood in the United 

States annually. Eitzer and Hites (1987) derived a dramatically different estimate: 3 kg/yr of 

total CDDs/CDFs (or 66 g I-TEQDF per year, assuming an I-TEQDF content in PCP of 3 mg/kg). 

Eitzer and Hites based their estimate on an assumption that 0.1% of the PCP produced annually 

enters the atmosphere and that the CDD/CDF contaminants present in the PCP (assumed to be 

130 mg/kg) are released to the atmosphere at the same rate as the PCP (i.e., 0.1%). The basis for 

the first assumption by Eitzer and Hites is not clear because U.S. EPA (1980), which was cited as 

the source of the 0.1% emission factor, does not appear to address volatilization of PCP from inservice treated wood. The report does, however, estimate that most PCP in treated wood leaches 

relatively rapidly from the wood, presumably to land, within a period of 12 years. 

Eduljee and Dyke (1996) and Douben et al. (1995) estimated that 0.8 g I-TEQDF is 

released to the air annually from PCP-treated wood in the United Kingdom. This estimate was 

based on the assumed emission of 0.1% of the CDDs/CDFs present in PCP-treated wood during 

the first year of the service life of the wood that was assumed by Eitzer and Hites (1987). No 

emissions were assumed for subsequent years of use of the treated wood. 

The California Air Resources Board (Chinkin et al., 1987) generated estimates of 

CDD/CDF volatilization releases at wood treatment facilities from bundles of treated wood that 

remain on site for 1 month prior to shipment. An “adapted” version of a model developed by 

McCord (1981) was used for estimating volatile releases from a constantly filling lagoon. The 

model is primarily driven by chemical-specific vapor pressures and air diffusivity coefficients. 

Chinkin et al. did not provide all model input parameter values used to generate the emission 

estimates. However, running the model with typical dimensions for treated poles yields an I8-60
 

TEQDF emission rate on the order of 6E-12 g/yr-pole, an extremely low number (170 billion poles 

would together emit 1 g TEQ/yr). 

In addition to cancelling some pesticide registrations and establishing product standards, 

EPA’s Office of Pesticide Programs (OPP) issued two DCIs in 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, 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 process, with 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, 1988c). EPA can also require additional data 

from registrants, when necessary, through various mechanisms, including the DCI process. 

The purpose of the first DCI, dated June and October 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, using an analysis of 

raw materials and process chemistry, those pesticides that might contain halogenated dibenzo-pdioxin (HDD) and halogenated dibenzofuran (HDF) contaminants. The 93 pesticides (76 

pesticide active ingredients) to which the DCI applied, along with their corresponding 

Shaughnessey and Chemical Abstract code numbers, are presented in Table 8-29. (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 registrations for these chemicals were subject to the 

requirements of the DCI unless their product qualified for a Generic Data Exemption (i.e., a 

registrant exclusively used a FIFRA-registered pesticide product as a source of an active 

ingredient identified in Table 8-29 in formulating a product). Registrants whose products did not 

meet the Generic Data Exemption were required to submit the types of data listed below to 

enable EPA to assess the potential for formation of tetra- through hepta-HDD or -HDF 

contaminants during manufacture. 

•	 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. Under 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. 

8-61
 
Table 8-29. Status of first pesticide Data Call-In: pesticides suspected of having the potential to be

contaminated with dioxins if synthesized under conditions favoring dioxin formation 

8-62
 


Shaughnessey 

code Pesticide [active ingredient] CAS No. 

Sup

withd

000014 Dichlorodifluoromethane 75-71-8 Y008706 O-(4-Bromo-2,5-dichlorophenyl) O,O-dimethyl phosphorothioate 2104-96-3 Y009105 Dimethylamine 2,3,5-triiodobenzoate 17601-49-9 Y012001 Neburon 555-37-3 Y

012101 Crufomate 299-86-5 Y019201 MCPB, 4-butyric acid [4-(2-Methyl-4-chlorophenoxy)butyric acid] 94-81-5 N

019202 MCPB, Na salt [Sodium 4-(2-methyl-4-chlorophenoxy)butyrate] 6062-26-6 N019401 4-Chlorophenoxyacetic acid 122-88-3 N

025501 Chloroxuron 1982-47-4 Y

027401 Dichlobenil 1194-65-6 N

28201 Propanil [3',4'-Dichloropropionanilide] 709-98-8 N

028601 Dichlofenthion [O-(2,4-Dichlorophenyl) O,O-diethyl phosphorothioate)] 97-17-6 Y029201 DDT [Dichloro diphenyl trichloroethane] 50-29-3 Y

29601 Dichlone [2,3-dichloro-1,4-naphthoquinone] 117-80-6 Y

029902 Ammonium chloramben [3-amino-2,5-dichlorobenzoic acid] 1076-46-6 Y029906 Sodium chloramben [3-amino-2,5-dichlorobenzoic acid] 1954-81-0 Y

030602 Sodium 2-(2,4-dichlorophenoxy)ethyl sulfate 136-78-7 Y031301 DCNA [2,6-Dichloro-4-nitroaniline] 99-30-9 N

031503 Potassium 2-(2-methyl-4-chlorophenoxy)propionate 1929-86-8 Y

031516 MCCP, DEA Salt [Diethanolamine 2-(2-methyl-4-chlorophenoxy)propionate] 1432-14-0 Y031563 MCPP, IOE [Isooctyl 2-(2-methyl-4-chlorophenoxy)propionate] 28473-03-2 N034502 Dicapthon [O-(2-chloro-4-nitrophenyl) O,O-dimethyl phosphorothioate] 2463-84-5 Y035502 Monuron trichloroacetate [3-(4-chlorophenyl)-1,1-dimethylurea trichloroacetate] 140-41-0 Y

35505 Diuron [3-(3,4-dichlorophenyl)-1,1-dimethylurea] 330-54-1 N

35506 Linuron [3-(3,4-dichlorophenyl)-1-methoxy-1-methylurea] 330-55-2 N35901 Metobromuron [3-(p-bromophenyl)-1-methoxy-1-methylurea] 3060-89-7 Y
Table 8-29. Status of first pesticide data call-in: pesticides suspected of having the potential to bec

contaminated with dioxins if synthesized under conditions favoring dioxin formation (continued) 

8-63
 

Shaughnessey 

code Pesticide [active ingredient] CAS No. 

Sup

withd

53501 Methyl parathion [O,O-Dimethyl O-p-nitrophenyl phosphorothioate] 298-00-0 N55001 Dichlorophene [Sodium 2,2'-methylenebis(4-chlorophenate)] 97-23-4 Y

55005 Dichlorophene, sodium salt [Sodium 2,2'-methylenebis(4-chlorophenate)] 10254-48-5 Y

55201 1,2,4,5-Tetrachloro-3-nitrobenzene 117-18-0 Y

57501 Ethyl parathion [O,O-diethyl O-p-nitrophenyl phosphorothioate] 56-38-2 N

58102 Carbophenothion [S-(((p-chlorophenyl)thio)methyl) O,O-diethyl phosphorodithioate] 786-19-6 Y

58301 Ronnel [O,O-dimethyl O-(2,4,5-trichlorophenyl) phosphorothioate] 229-84-3 Y58802 Mitin FF [Sodium 5-chloro-2-(4-chloro-2-(3-(3,4-dichlorophenyl)ureido)phenoxy) 

benzenesulfonate] 

3567-25-7 N

59401 Orthodichlorobenzene 95-50-1 Y61501 Paradichlorobenzene 106-46-7 N62201 Chlorophene [2-Benzyl-4-chlorophenol] 120-32-1 N

62202 Potassium 2-benzyl-4-chlorophenate 35471-49-9 N

62203 Sodium 2-benzyl-4-chlorophenate 3184-65-4 N

62204 2-Chlorophenol 95-57-8 Y62206 2-Chloro-4-phenylphenol 92-04-6 Y62207 Potassium 2-chloro-4-phenylphenate 18128-16-0 Y62208 4-Chloro-2-phenylphenol NA Y

62209 4-Chloro-2-phenylphenol, potassium salt 53404-21-0 Y62210 6-Chloro-2-phenylphenol 85-97-2 Y62211 6-Chloro-2-phenylphenol, potassium salt 18128-17-1 Y62212 4-Chloro-2-phenylphenol, sodium salt 10605-10-4 Y

62213 6-Chloro-2-phenylphenol, sodium salt 10605-11-5 Y

62214 4 and 6-Chloro-2-phenylphenol, diethanolamine salt 53537-63-6 Y62215 2-Chloro-4-phenylphenol, sodium salt 31366-97-9 Y

64202 4-Chloro-2-cyclopentylphenol 13347-42-7 Y
Table 8-29. Status of first pesticide data call-in: pesticides suspected of having the potential to bec

contaminated with dioxins if synthesized under conditions favoring dioxin formation (continued) 

8-64
 


Shaughnessey 

code Pesticide [active ingredient] CAS No. 

Sup

withd

64208 Fentichlor [2,2'-Thiobis(4-chloro-6-methylphenol)] 4418-66-0 Y

64209 Fentichlor [2,2'-Thiobis(4-chlorophenol)] 97-24-5 Y64214 4-Chloro-2-cyclopentylphenol, potassium salt of 35471-38-6 Y64218 4-Chloro-2-cyclopentylphenol, sodium salt 53404-20-9 Y67707 Chlorophacinone 3691-35-8 N

69105 ADBAC [Alkyl* dimethyl benzyl ammonium chloride *(50% C14, 40% C12, 10% 

C16)] 

68424-85-1 N

69144 ADBAC [Alkyl* dimethyl 3,4-dichlorobenzyl ammonium chloride *(61% C12, 23% 

C14, 11% C16, 5% C18)] 

NA N

77401 Niclosamide [2-Aminoethanol salt of 2',5-dichloro-4'-nitrosalicylanilide] 1420-04-8 N

77406 5-Chlorosalicylanilide 4638-48-6 Y78780 2-Methyl-4-isothiazolin-3-one NA Y

79202 Tetradifon [4-chlorophenyl 2,4,5-trichlorophenyl sulfone] 116-29-0 Y79301 Chloranil [tetrachloro-p-benzoquinone] 118-75-2 Y80403 6-Chlorothymol 89-68-9 Y80811 Anilazine [2,4-Dichloro-6-(o-chloroanilino)-s-triazine] 101-05-3 Y

81901 Chlorothalonil [tetrachloroisophthalonitrile] 1897-45-6 N

82602 Sodium 2,3,6-Trichlorophenylacetate 2439-00-1 Y84101 Chlorfenvinphos 470-90-6 Y

84901 O-(2-Chloro-1-(2,5-dichlorophenyl)vinyl) O,O-diethyl phosphorothioate 1757-18-2 Y

86801 PCMX [4-Chloro-3,5-xylenol] 88-04-0 N97003 Piperalin [3-(2-Methylpiperidino)propyl 3,4-dichlorobenzoate] 3478-94-2 N

100601 Fenamiphos NA N101001 p-Chlorophenyl diiodomethyl sulfone 20018-12-6 Y101101 Metribuzin 21087-64-9 N104301 Bifenox [methyl 5-(2,4-dichlorophenoxy)-2-nitrobenzoate] 42576-02-3 Y106001 Methazole [2-(3,4-dichlorophenyl)-4-methyl-1,2,4-oxadiazolidine-3,5-dione] 20354-26-1 Y
Table 8-29. Status of first pesticide data call-in: pesticides suspected of having the potential to bec

contaminated with dioxins if synthesized under conditions favoring dioxin formation (continued) 

8-65
 


Shaughnessey 

code Pesticide [active ingredient] CAS No. 

Sup

withd

108201 Diflubenzuron [N-(((4-chlorophenyl)amino)carbonyl)-2,6-difluorobenzamide] 35367-38-5 N

109001 Oxadiazon [2-tert-butyl-4-(2,4-dichloro-5-isopropoxyphenyl)- delta 2 -1,3,4­

oxadiazoline-5-one] 

19666-30-9 N

109301 Fenvalerate 51630-58-1 N109302 Fluvalinate [N-2-Chloro-4-trifluoromethyl)phenyl-DL-valine (+-)-cyano(3-phenoxyphenyl)methyl ester] 69409-94-5 N109801 Iprodione [3-(3,5-Dichlorophenyl)-N-(1-methylethyl)-2,4-dioxo-1­

imidazolidinecarboxamide (9CA)] 

36734-19-7 N

109901 Triadimefon 

[1-(4-Chlorophenoxy)-3,3-dimethyl-1-(1H-1,2,4-triazol-1-yl)-2-butanone] 

43121-43-3 N

110902 Diclofop - methyl [methyl 2-(4-(2,4-dichlorophenoxy)phenoxy)propanoate] 51338-27-3 N111401 Profenofos [O-(4-Bromo-2-chlorophenyl)-O-ethyl S-propyl phosphorothioate] 41198-08-7 N

111601 Oxyfluorfen [2-chloro-1-(3-ethoxy-4-nitrophenoxy)-4-(trifluoromethyl)benzene] 42874-03-3 N111901 Imazalil [1-(2-(2,4-Dichlorophenyl)-2-(2-propenyloxy)ethyl)-1H-imidazole] 35554-44-0 N112802 Bromothalin [N-Methyl-2,4-dinitro-n-(2,4,6-tribromophenyl)-6­

(trifuloromethyl)benzenamine] 

63333-35-7 N

113201 Vinclozolin [3-(3,5-Dichlorophenyl)-5-ethenyl-5-methyl-2,4-oxazolidinedione (9CA)] 50471-44-8 N119001 Fenridazon [Potassium 1-(p-chlorophenyl)-1,4-dihydro-6-methyl-4-oxo- pyridazine3-carboxylate] 83588-43-6 N123901 Tridiphane [2-(3,5-Dichlorophenyl)-2-(2,2,2-trichloroethyl) oxirane] 58138-08-2 N

125601 Paclobutrazol 76738-62-0 N128838 Linalool 78-70-6 N

206600 Fenarimol [a-(2-chlorophenyl)-a-(4-chlorophenyl)-5-pyrimidinemethanol] 60168-88-9 N

NA = Not available -- = No information given 
•	 Discussion of the formation of impurities. Under the requirements mandated by 40 

CFR 158.120, Subdivision D, EPA required submittal of a detailed discussion and 

assessment of the possible formation of HDDs and HDFs. 

Registrants had the option of voluntarily canceling their product or “reformulate to remove an 

active ingredient” to avoid having to comply with the DCI. 

The second DCI, dated June and October 1987, “Data Call-In for Analytical Chemistry 

Data on Polyhalogenated Dibenzo-p-Dioxins/Dibenzofurans (HDDs and HDFs),” was issued for 

68 pesticides (16 pesticide active ingredients) suspected to be contaminated with CDDs/CDFs 

(see Table 8-30). All registrants supporting registrations for these pesticides were subject to the 

requirements of this DCI unless the product qualified for various exemptions or waivers. 

Pesticides covered by the second DCI were strongly suspected by EPA of containing detectable 

levels of CDDs/CDFs. 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 CDDs or CDFs. Registrants were required to 

develop an analytical method for measuring the HDD/HDF content of their products. 

The DCI established a regimen for defining the precision of the analytical method. 

Target LOQs were established in the DCI for specific CDD/CDF congeners (see 

Table 8-25). 

•	 Certification of limits of CDDs or CDFs. 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 of developing 

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 did not apply, or the application or award of a low-volume, minor-use waiver. 

The data contained in CSFs, as well as any other data generated under 40 CFR 158.120, 

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 

8-66
 

Table 8-30. Status of second pesticide Data Call-In: pesticides suspected of being contaminated wi

8-67
 

Shaughnessey 

code Pesticide [active ingredient] CAS No. 

Suppor

withdraw

29801 Dicamba [3,6-dichloro-o-anisic acid] 1918-00-9 No 

29802 Dicamba dimethylamine [3,6-dichloro-o-anisic acid] 2300-66-5 No 29803 Diethanolamine dicamba [3,6-dichloro-2-anisic acid] 25059-78-3 Yes 

30001 2,4-Dichlorophenoxyacetic acid 94-75-7 No 

30002 Lithium 2,4-dichlorophenoxyacetate 3766-27-6 No 

30003 Potassium 2,4-dichlorophenoxyacetate 14214-89-2 Yes 

30004 Sodium 2,4-dichlorophenoxyacetate 2702-72-9 No 30005 Ammonium 2,4-dichlorophenoxyacetate 2307-55-3 Yes 30010 Alkanol* amine 2,4-dichlorophenoxyacetate *(salts of the ethanol and 

ispropanol series) 

NA Yes 

30011 Alkyl* amine 2,4-dichlorophenoxyacetate *(100% C12) 2212-54-6 Yes 30013 Alkyl* amine 2,4-dichlorophenoxyacetate *(100% C14) 28685-18-9 Yes 30014 Alkyl* amine 2,4-dichlorophenoxyacetate *(as in fatty acids of tall oil) NA Yes 30016 Diethanolamine 2,4-dichlorophenoxyacetate 5742-19-8 No 30017 Diethylamine 2,4-dichlorophenoxyacetate 20940-37-8 Yes 30019 Dimethylamine 2,4-dichlorophenoxyacetate 2008-39-1 No 

30020 N,N-Dimethyloleylamine 2,4-dichlorophenoxyacetate 53535-36-7 Yes 30021 Ethanolamine 2,4-dichlorophenoxyacetate 3599-58-4 Yes 30023 Heptylamine 2,4-dichlorophenoxyacetate 37102-63-9 Yes 

30024 Isopropanolamine 2,4-dichlorophenoxyacetate 6365-72-6 Yes 30025 Isopropylamine 2,4-dichlorophenoxyacetate 5742-17-6 No 30028 Morpholine 2,4-dichlorophenoxyacetate 6365-73-7 Yes 30029 N-Oleyl-1,3-propylenediamine 2,4-dichlorophenoxyacetate 2212-59-1 Yes 

30030 Octylamine 2,4-dichlorophenoxyacetate 2212-53-5 Yes 30033 Triethanolamine 2,4-dichlorophenoxyacetate 2569-01-9 Yes 30034 Triethylamine 2,4-dichlorophenoxyacetate 2646-78-8 No 

30035 Triisopropanolamine 2,4-dichlorophenoxyacetate 32341-80-3 No 
Table 8-30. Status of second pesticide Data Call-In: pesticides suspected of being contaminated wi

(continued) 

8-68
 


Shaughnessey 

code Pesticide [active ingredient] CAS No. 

Suppor

withdraw

30039 N,N-Dimethyl oleyl-linoleyl amine 2,4-dichlorophenoxyacetate 55256-32-1 Yes 30052 Butoxyethoxypropyl 2,4-dichlorophenoxyacetate 1928-57-0 Yes 

30053 Butoxyethyl 2,4-dichlorophenoxyacetate 1929-73-3 No 

30055 Butoxypropyl 2,4-dichlorophenoxyacetate 1928-45-6 Yes 

30056 Butyl 2,4-dichlorophenoxyacetate 94-80-4 Yes 30062 Isobutyl 2,4-dichlorophenoxyacetate 1713-15-1 Yes 

30063 Isooctyl(2-ethylhexyl) 2,4-dichlorophenoxyacetate 1928-43-4 No 

30064 Isooctyl(2-ethyl-4-methylpentyl) 2,4-dichlorophenoxyacetate 25168-26-7 Yes 30065 Isooctyl(2-octyl) 2,4-dichlorophenoxyacetate 1917-97-1 Yes 30066 Isopropyl 2,4-dichlorophenoxyacetate 94-11-1 No 30072 Propylene glycol butyl ether 2,4-dichlorophenoxyacetate 1320-18-9 Yes 30801 4-(2,4-Dichlorophenoxy)butyric acid 94-82-6 No 30804 Sodium 4-(2,4-dichlorophenoxy)butyrate 10433-59-7 No 30819 Dimethylamine 4-(2,4-dichlorophenoxy)butyrate 2758-42-1 No 30853 Butoxyethanol 4-(2,4-dichlorophenoxy)butyrate 32357-46-3 Yes 30856 Butyl 4-(2,4-dichlorophenoxy)butyrate 6753-24-8 Yes 30863 Isooctyl 4-(2,4-dichlorophenoxy)butyrate 1320-15-6 Yes 31401 2-(2,4-Dichlorophenoxy)propionic acid 120-36-5 No 

31419 Dimethylamine 2-(2,4-dichlorophenoxy)propionate 53404-32-3 No 31453 Butoxyethyl 2-(2,4-dichlorophenoxy)propionate 53404-31-2 No 31463 Isooctyl 2-(2,4-dichlorophenoxy)propionate 28631-35-8 No 

31501 MCPP acid [2-(2-Methyl-4-chlorophenoxy)propionic acid] 7085-19-0 No 

31519 MCPP, DMA [Dimethylamine 2-(2-methyl-4-chlorophenoxy)propionate] 32351-70-5 No 35301 Bromoxynil [3,5-dibromo-4-hydroxybenzonitrile] 1689-84-5 No 44901 Hexachlorophene [2,2'-Methylenebis(3,4,6-trichlorophenol)] 70-30-4 Yes 
Table 8-30. Status of second pesticide Data Call-In: pesticides suspected of being contaminated wi

(continued) 

8-69
 


Shaughnessey 

code Pesticide [active ingredient] CAS No. 

Suppor

withdraw

44902 Hexachlorophene, Na salt [Monosodium 

2,2'-methylenebis(3,4,6-trichlorophenate)] 

5736-15-2 Yes 

44904 Hexachlorophene, K salt [Potassium 

2,2'-methylenebis(3,4,6-trichlorophenate)] 

67923-62-0 Yes 

54901 Irgasan [5-Chloro-2-(2,4-dichlorophenoxy)phenol] 3380-34-5 No 63004 Tetrachlorophenols 25167-83-3 Yes 

63005 Tetrachlorophenols, sodium salt 25567-55-9 Yes 

63006 Tetrachlorophenols, alkyl* amine salt*(as in fatty acids of coconut oil) NA Yes 

63007 Tetrachlorophenols, potassium salt 53535-27-6 Yes 64203 Bithionolate sodium [Disodium 2,2'-thiobis(4,6-dichlorophenate)] 6385-58-6 Yes 64212 Phenachlor [2,4,6-Trichlorophenol] 88-06-2 Yes 

64219 Potassium 2,4,6-trichlorophenate 2591-21-1 Yes 64220 2,4,6-Trichlorophenol, sodium salt 3784-03-0 Yes 64501 Phenothiazine 92-84-2 Yes 78701 Dacthal-DCPA [Dimethyl tetrachloroterephthalate] 1861-32-1 No 79401 Endosulfan [hexachlorohexahydromethano-2,4,3-benzodioxathiepin-3-oxide] 115-29-7 No 82501 Silvex [2-(2,4,5-trichlorophenoxy)propionic acid] 93-72-1 Yes 83701 Tetrachlorvinphos [2-Chloro-1-(2,4,5-trichlorophenyl)vinyl dimethyl 

phosphate] 

961-11-5 No 

104101 Edolan [Sodium 1,4',5'-trichloro-2'-(2,4,5-trichlorophenoxy) 

methanesulfonanilide] 

69462-14-2 Yes 

NA = Not available -- = No information given 
based on the trends identified in those data as well as in data made public by EPA are 

summarized below. 

The two DCIs included 161 pesticides. Of these, 92 are no longer supported by 

registrants. Following evaluation of the process chemistry submissions required under the DCIs, 

OPP determined that formation of CDDs/CDFs was not likely during the manufacture of 43 of 

the remaining 69 pesticides; thus, analysis of samples of these 43 pesticides was not required by 

OPP. Evaluation of process chemistry data is ongoing at OPP for an additional 7 pesticides. 

Table 8-30 indicates which pesticides are no longer supported and those for which OPP has 

required further testing to determine whether CDDs/CDFs are present as contaminants (U.S. 

EPA, 1995g). OPP required that analysis of production samples be performed on the remaining 

19 pesticides (see Table 8-31). The status of the analytical data generation/evaluation to date is 

summarized as follows: (a) no detection of CDDs/CDFs above the LOQs in registrant 

submissions for 13 active ingredients, (b) detection of CDDs/CDFs above the LOQs for 2,4-D 

acid (two submissions) and 2,4-D 2-ethyl hexyl acetate (one submission), and (c) ongoing data 

generation or evaluation for four pesticides. 

Table 8-32 presents a summary of results obtained by EPA for CDDs/CDFs in eight 

technical 2,4-D herbicides; these data were extracted from program files in OPP. Because some 

of these files contained CBI, the data in this table were reviewed by OPP staff to ensure that no 

CBI was being disclosed (memorandum dated May 28, 1996, from S. Funk, U.S. EPA, to W. 

Hazel, U.S. EPA). Figure 8-5 presents a congener profile for 2,4-D based on the average 

congener concentrations reported in Table 8-32. 

Schecter et al. (1997) reported the results of analyses of samples of 2,4-D manufactured 

in Europe, Russia, and the United States (see Table 8-33). The total TEQ concentrations 

measured in the European and Russian samples were similar to those measured in the EPA DCI 

samples; however, the levels reported by Schecter et al. for U.S. samples were significantly 

lower. Similarly, Masunaga et al. (2001) reported the analyses of two agrochemical formulations 

containing 2,4-D manufactured in Japan (Table 8-33). The total TEQ concentration measured in 

one of the samples was similar to what Schecter et al. (1997) reported for the U.S. samples; no 

TEQ was detected in the other sample. 

As discussed in Section 11.2.1 (Chapter 11), an estimated 28,100 metric tons of 2,4-D 

were used in the United States in 2000, making it one of the top 10 pesticides in terms of quantity 

used (EPA proprietary data). The pesticide 2,4-D is the only product judged to have the potential 

for environmental release through its agricultural use. However, no estimate of environmental 

release can be made for 2000. Since 1995, the chemical manufacturers of 2,4-D have been 

undertaking voluntary actions to significantly reduce the dioxin content of the product. No 

information is available on the level of dioxin contamination, if any, that may have been present 

8-70
 

Table 8-31. Summary of analytical data submitted to EPA in response to pesticide Data Call-In(s) 

8-71
 


Shaughnessey 

code 

Pesticide 

Common name Chemical name 

019201 MCPB, 4-butyric acid 4-(2-methyl-4-chlorophenoxy)butyric acid 019401 4-CPA 4-Chlorophenoxyacetic acid 

027401 Dichlobenil 2,6-Dichlorobenzonitrile 029801 Dicamba 3,6-Dichloro-o-anisic acid 029802 Dicamba, dimethylamine 3,6-Dichloro-o-anisic acid, dimethylamine salt 030001 2,4-D 2,4-Dichlorophenoxy acetic acid 030063 2,4-D, 2EH Isooctyl(2-ethylhexyl)2,4-dichlorophenoxyacetate 030801 2,4-DB 4-(2,4-Dichlorophenoxy)butyric acid 

031301 DCNA 2,6-Dichloro-4-nitroaniline 031401 2,4-DP 2-(2,4-Dichlorophenoxy)propionic acid 031501 Mecoprop (MCPP) 2-(2-methyl-4-chlorophenoxy)propionic acid 

035301 Bromoxynil 3,5-Dibromo-4-hydroxybenzonitrile 054901 Irgasan 5-Chloro-2-(2,4-dichlorophenoxy)phenol 078701 Dacthal (DCPA) Dimethyl tetrachloroterephthalate 

081901 Chlorothalonil Tetrachloroisophthalonitrile 083701 Tetrachlorvinphos 2-Chloro-1-(2,4,5-trichlorophenyl)vinyl dimethyl phosphate 108201 Diflubenzuron N-(((4-chlorophenyl)amino)carbonyl)-2,6-difluorobenzamide 109001 Oxadiazon 2-Tert-butyl-4(2,4-dichloro-5-isopropoxyphenyl)-delta2-1,3,4-oxadiazoline-5-one 110902 Dichlofop-methyl Methyl-2-(4-(2,4-dichlorophenoxy)phenoxy) propanoate 

aPositive is defined as the detection of any congener at a concentration equal to or exceeding the limits of quantitation listed in Table 8-

Sources: U.S. EPA (1995f); facsimile dated March 27, 1996 from S. Funk, U.S. EPA, to D. Cleverly, U.S. EPA. 
Table 8-32. Summary of results for CDDs/CDFs in technical 2,4-D and 2,4-D 

ester herbicides 

Congener 

EPA 

LOQa 

(μg/kg) 

Total 

no. of 

technicals 

Number of 

technicals 

greater than 

LOQ 

Observed 

maximum 

conc. 

(μg/kg) 

Average 

conc.b 

(μg/kg) 

2,3,7,8-TCDD 0.1 8 2 0.13 0.06 

1,2,3,7,8-PeCDD 0.5 8 3 2.6 0.78 

1,2,3,4,7,8-HxCDD 2.5 8 0 0.81 0.31 

1,2,3,6,7,8-HxCDD 2.5 8 0 0.77 0.39 

1,2,3,7,8,9-HxCDD 2.5 8 0 0.68 0.24 

1,2,3,4,6,7,8-HpCDD 100.0 8 0 1.5 0.21 

OCDD -­ -­ -­ -­ -­

2,3,7,8-TCDF 1 8 0 0.27 0.07 

1,2,3,7,8-PeCDF 5 8 0 0.62 0.38 

2,3,4,7,8-PeCDF 5 7 0 0.73 0.07 

1,2,3,4,7,8-HxCDF 25 8 0 1.6 0.36 

1,2,3,6,7,8-HxCDF 25 8 0 1.2 0.11 

1,2,3,7,8,9-HxCDF 25 8 0 1.4 0.16 

2,3,4,6,7,8-HxCDF 25 8 0 1.1 0.14 

1,2,3,4,6,7,8-HpCDF 1,000 8 0 8.3 2.17 

1,2,3,4,7,8,9-HpCDF 1,000 8 0 1.2 0.18 

OCDF -­ -­ -­ -­ --

TOTALc 5.6 

I-TEQDF 0.7 

TEQDF-WHO98 1.1 

a

LOQ required by EPA in the data call-in. b

Average of the mean results for multiple analyses of four technical 2,4-D and/or 2,4-D ester products for which

 detectable CDD/CDF congener concentrations less than the LOQs were quantified; nondetect values were

 assumed to be zero. c

Total equals the sum of the individual congener averages. 

LOQ = Limit of quantitation 

-- = Analyses not performed 

Source: U.S. EPA Office of Pesticide Programs file. 

in 2,4-D in 2000. An estimated 26,300 and 30,400 metric tons were used during 1995 and 1987, 

respectively (U.S. EPA, 1997e, 1988d). On the basis of the average CDD/CDF congener 

concentrations in 2,4-D presented in Table 8-33 (not including OCDD and OCDF), the 

corresponding TEQDF-WHO98 concentration is 1.1 μg/kg (0.7 μg I-TEQDF/kg). Combining this 

8-72
 

2,3,7,8-TCDD

1,2,3,7,8-PeCDD

1,2,3,4,7,8-HxCDD1,2,3,6,7,8-HxCDD1,2,3,7,8,9-HxCDD

1,2,3,4,6,7,8-HpCDD

OCDD

2,3,7,8-TCDF

1,2,3,7,8-PeCDF2,3,4,7,8-PeCDF

1,2,3,4,7,8-HxCDF1,2,3,6,7,8-HxCDF1,2,3,7,8,9-HxCDF2,3,4,6,7,8-HxCDF

1,2,3,4,6,7,8-HpCDF1,2,3,4,7,8,9-HpCD

OCDF

0%

5%

10%

15%

20%

25%

30%

35%

40%

45%

Percent of Total CDDs and CDFs 

Congener

Figure 8-5. Congener profile for 2,4-D (salts and esters) (based on mean 

concentrations reported in Table 8-34). 

TEQ concentration with the activity level estimates for 1995 and 1987 indicates that 28.9 g 

TEQDF-WHO98 (18.4 g I-TEQDF) were released in 1995 and 33.4 g TEQDF-WHO98 (21.3 g I

TEQDF) in 1987. The release estimates for 1987 and 1995 are assigned a high confidence rating, 

indicating high confidence in both the production and the emission factor estimates. Because no 

estimate can be made for 2000, it is rated as Category E (no estimate of environmental release of 

CDDs/CDFs can be made because of the poor quality of existing information). 

8.4. OTHER CHEMICAL MANUFACTURING AND PROCESSING SOURCES 

8.4.1. Municipal Wastewater Treatment Plants 

8.4.1.1. Sources 

CDDs/CDFs have been measured in nearly all sewage sludges tested, although the 

concentrations and, to some extent, the congener profiles and patterns differ widely. Potential 

sources of the CDDs/CDFs include microbial formation (discussed in Chapter 9), runoff to 

­

8-73
 

Table 8-33. CDD/CDF concentrations (μg/kg) in samples of 2,4-D and pesticide formulations conta

8-74
 


Congener/congener 

group 

Acbar 

Super 

(Gaza 

Citya) 

Amco 

Super 

(Gaza 

Citya) (Bethlehem)a 

Chimprom 

(Russia) 

Dragon 

Lawn Weed 

Killer 

KGRO 

(U.S.) 

Pro Care Premium 

(U.S.) 

Ortho 

Weed-BGone 

(U.S.) 

Sigma 

Co. 

(U.S.) 

Amer

Bra

Chem

Co. (U

2,3,7,8-TCDD ND (0.1) ND (0.1) ND (0.1) ND (0.02) ND (0.001) -­ -­ -­ -­ -

1,2,3,7,8-PeCDD 0.1 ND (0.1) 1.2 0.03 0.0014 -­ -­ -­ -­ -

1,2,3,4,7,8-HxCDD ND (0.1) ND (0.1) ND (0.1) 0.02 ND (0.001) -­ -­ -­ -­ -

1,2,3,6,7,8-HxCDD ND (0.1) 0.2 0.6 0.05 0.0024 -­ -­ -­ -­ -

1,2,3,7,8,9-HxCDD ND (0.1) ND (0.1) 0.4 ND (0.02) 0.001 -­ -­ -­ -­ -

1,2,3,4,6,7,8-HpCDD 0.1 1.2 0.3 0.23 0.0017 -­ -­ -­ -­ -

OCDD 0.1 2.6 0.1 0.85 0.0063 -­ -­ -­ -­ -

2,3,7,8-TCDF 

1,2,3,7,8-/1,2,3,4,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-/1,2,3,4,7,9 -HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.3 

ND (0.1) 

ND (0.1) 

ND (0.1) 

ND (0.1) 

ND (0.1) 

ND (0.1) 

0.1 

ND (0.1) 

0.2 

ND (0.1) 

0.2 

ND (0.1) 

0.1 

ND (0.1) 

ND (0.1) 

ND (0.1) 

0.8 

ND (0.1) 

3.8 

ND (0.1) 

0.7 

0.1 

0.4 

0.1 

ND (0.1) 

0.1 

0.1 

ND (0.1) 

0.4 

ND (0.1) 

1.2 

0.06 

0.08 

0.11 

ND (0.02) 

0.05 

0.24 

0.02 

0.46 

0.0036 

0.001 

0.0011 

0.0013 

ND (0.001) 

ND (0.001) 

0.0011 

0.0016 

ND (0.001) 

0.0039 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-

-

-

-

-

-

-

-

-

-

Total 2,3,7,8-CDD (nondetect = 0) 0.3 4 2.6 1.18 0.0128 0.0144 0.0143 0.0091 0.127 0.0Total 2,3,7,8-CDF  (nondetect = 0) 0.6 4.9 1.9 2.22 0.0136 0.1628 0.4253 0.1095 3.0507 0.0Total I-TEQDF 

 (nondetect = 0)b 0.082 0.066 0.85 0.142 0.0023 0.0009 0.0012 0.0014 0.0013 0.0

Total TEQDF-WHO98 

 (nondetect = 0)b 0.134 0.061 1.449 0.156 0.003 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-

-

-

-

-

-

-

-

-

-

Total CDD/CDF -­ -­ -­ -­ -­ -­ -­ -­ -­ -
Table 8-33. CDD/CDF concentrations (μg/kg) in samples of 2,4-D and pesticide formulations conta

(continued) 

a2,4-D manufactured in Europe and packaged in Palestine. bCalculated assuming nondetect values are zero. 

ND = not detected (value in parenthesis is the detection limit) -- = No information given Sources: Schecter et al. (1997); Masunaga et al. (2001). 

8-75
 

sewers from lands or urban surfaces contaminated by product uses or deposition of previous 

emissions to air (discussed in Chapter 11, Section 11.2.1), household wastewater, industrial 

wastewater, and chlorination operations within the wastewater treatment facility (Rappe, 1992b; 

Rappe et al., 1994; Horstmann et al., 1992; Sewart et al., 1995; Cramer et al., 1995; Horstmann 

and McLachlan, 1995). 

The major source(s) for a given publicly owned treatment works (POTW) is likely to be 

site specific, particularly in industrialized areas. For example, Rieger and Ballschmiter (1992) 

traced the origin of CDDs/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. However, a series of studies by Horstmann et al. (1992, 

1993a, b) and Horstmann and McLachlan (1994a, b, 1995) demonstrated that wastewater 

generated by laundering and bathing could be the major source at many, if not all, POTWs that 

serve primarily residential populations. Although runoff from streets during precipitation events, 

particularly from streets with high traffic density, was reported by these researchers as 

contributing measurably, the total contribution of TEQ from household wastewater was eight 

times greater than that from surface runoff at the study city. 

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 CDDs/CDFs 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 hexaCDDs. Because of the similar concentrations and congener profiles found, the authors concluded 

that the presence of CDDs/CDFs in washing machine wastewater is widespread. A simple mass 

balance performed using the results (Horstmann and McLachlan, 1994a) showed that the 

CDDs/CDFs found in the four washing machine wastewater samples could account for 27 to 

94% of the total CDDs/CDFs measured in the sludge of the local wastewater treatment plant. 

Horstmann et al. (1993a) performed additional experiments that showed that detergents, 

commonly used bleaching agents, and the washing cycle process itself were not responsible for 

the observed CDDs/CDFs. To determine whether the textile fabric or fabric finishing processes 

could account for the observed CDDs/CDFs, Horstmann et al. (1993b), Horstmann and 

McLachlan (1994a, b), and Klasmeier and McLachlan (1995) analyzed the CDD/CDF content of 

raw cotton cloth, white synthetic materials, and more than 100 new textile finished products. 

Low concentrations were found in most products (less than 50 ng/kg of total CDDs/CDFs), but a 

small percentage contained high concentrations, up to 290 μg/kg of total CDDs/CDFs. On the 

basis of the concentrations and patterns found, the authors concluded that neither unfinished new 

fabrics nor common cotton finishing processes could explain the CDD/CDF levels found in 

8-76
 

wastewater; rather, the use of CDD-/CDF-containing textile dyes and pigments and the use in 

some developing countries of PCP to treat unfinished cotton appeared to be the sources of the 

detected CDDs/CDFs. 

Horstmann and McLachlan (1994a, b, 1995) reported the results of additional 

experiments showing that the small percentage of clothing items with high CDD/CDF levels 

could be responsible for the quantity of CDDs/CDFs observed in household wastewater and 

sewage sludge. They demonstrated that the CDDs/CDFs can be gradually removed from the 

fabric during washing; they can be transferred to the skin, subsequently transferred back to other 

textiles, and then washed out, or they can be transferred to other textiles during washing and then 

removed during subsequent washing. 

8.4.1.2. Releases to Water 

8.4.1.2.1. Emissions data. The presence of CDDs/CDFs in sewage sludge suggests that 

CDDs/CDFs may also be present in the wastewater effluent discharges of POTWs; however, few 

studies reporting the results of effluent analyses for CDDs/CDFs have been published. 

Rappe et al. (1989a) tested the effluent from two Swedish POTWs for all 2,3,7,8­

substituted CDD/CDF congeners. OCDD was detected in the effluents from both facilities at 

concentrations ranging from 14 to 39 pg/L. The investigators detected 1,2,3,4,6,7,8-HpCDD and 

1,2,3,4,6,7,8-HpCDF in the effluent of one facility at concentrations of 2.8 and 2 pg/L, 

respectively. No 2,3,7,8-substituted tetra-, penta-, and hexa-CDDs or CDFs were detected (DLs 

of 0.2 to 20 pg/L). 

Ho and Clement (1990) reported the results of sampling during the late 1980s of 37 

POTWs in Ontario, Canada, for each of the five CDD/CDF congener groups with four to eight 

chlorines. The sampled facilities included 27 secondary treatment facilities, seven primary 

treatment facilities, one tertiary plant, and two lagoons. The facilities accounted for about 73% 

of the sewage discharged by POTWs in Ontario. No CDDs/CDFs were detected (DL in the low 

ng/L range) in the effluents from the lagoons and the tertiary treatment facility. Only OCDD and 

TCDF were detected in the effluents from the primary treatment facilities (two and one effluent 

samples, respectively). HpCDD, OCDD, TCDF, and OCDF were detected in the effluents from 

the secondary treatment facilities (detected in four or fewer samples at levels ranging from 0.1 to 

11 ng/L). 

Gobran et al. (1995) analyzed the raw sewage and final effluent of an Ontario, Canada, 

wastewater treatment plant for CDD/CDF congeners over a 5-day period. Although HpCDD, 

OCDD, HpCDF, and OCDF were detected in the raw sewage (12 to 2,300 pg/L), no CDDs/CDFs 

were detected in the final effluent at congener-specific DLs ranging from 3 to 20 pg/L. 

8-77
 
The California Regional Water Quality Control Board reported the results of effluent 

testing at nine POTWs in the San Francisco area (memorandum dated March 21, 1996, from Lila 

Tang, California Regional Water Quality Control Board, to David Cleverly, U.S. EPA). A total 

of 30 samples were collected between 1992 and 1995, and 1 to 6 samples were analyzed for each 

POTW. The mean concentrations of CDDs/CDFs are shown in Table 8-34. The overall mean 

TEQ concentration was 0.27 pg TEQDF-WHO98/L (0.29 pg I-TEQDF/L). With the exception of 

OCDD, most 2,3,7,8-substituted CDD/CDF congeners were seldom detected. 

Rappe et al. (1998) analyzed effluent samples from 17 POTWs in Mississippi, 10 of 

which receive input from industrial facilities. Treatment processes at the facilities include the 

use of one or more of the following: lagoons, activated sludge, aerated digestion, wetlands, 

oxidative ditch, and trickling filter. Additionally, 12 of the facilities use chlorine gas in the 

treatment process. The wastewater flows at the facilities range from 0.11 to 39.75 million L/day; 

however, wastewater flow rates were not known for two facilities. Table 8-35 presents the 

concentrations of dioxins measured in the effluent samples for each facility and total TEQ 

emission factors. Concentrations were congener-specific for only 2,3,7,8-TCDD; 2,3,7,8-TCDF; 

1,2,3,7,8-PeCDD; 2,3,4,7,8-PeCDF; OCDD; and OCDF. Also provided were concentrations for 

total HxCDD and total HpCDD. The total I-TEQ concentrations reported by Rappe et al. 

(assuming nondetect values equal to one-half the DL) ranged from 0.274 to 3.84 pg I-TEQDF/L 

(average of 0.81 pg/I-TEQDF/L). Because concentrations for all congeners were not provided, 

emission factors could not be calculated in TEQDF-WHO98. 

The CRWQCB data (memorandum dated March 21, 1996, from Lila Tang, California 

Regional Water Quality Control Board, to David Cleverly, U.S. EPA) were collected to provide 

representative effluent concentrations for the San Francisco area. These data cannot be 

considered to be representative of CDD/CDF effluent concentrations at the 16,000-plus POTWs 

nationwide. Therefore, the data can be used only to generate a preliminary estimate of the 

potential mass of CDD/CDF TEQ that may be released annually by U.S. POTWs. 

8.4.1.2.2. Activity level information. Based on the results of the 1996 and 2000 Clean Water 

Needs Surveys, an estimated 122 billion L and 148 billion L of wastewater were treated daily by 

POTWs in the United States in 1996 and 2000, respectively (U.S. EPA, 1997c, 2004). 

Wastewater treatment data were not available for 1987; however, an estimate was 

developed using the population of the United States as a surrogate. In 2000, the population of the 

United States was approximately 281 million people. Using the estimate of water treated daily 

by POTWs in 2000, approximately 527 L/person of wastewater were treated daily by POTWs. In 

1990, the population of the United States was approximately 249 million people. Assuming the 

8-78
 

Table 8-34. Mean CDD/CDF concentrations and range of detection limits 

(DLs) (pg/L) in effluents from nine U.S. publicly owned treatment works 

(POTWs) 

Congener/congener 

group 

No. of 

detections/

 samples 

Range of 

DLs 

Range of detected 

concentrations 

(POTW mean basis) 

Overall mean 

concentrationsa 

Minimum Maximum 

Nondetect 

set to 0 

Nondetect 

set to 1⁄2 DL 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0/30 

0/30 

0/30 

0/30 

0/30 

3/30 

13/30 

0.31–8.8 

0.45–15 

0.43–9.8 

0.81–10 

0.42–9.7 

0.75–18 

6.2–57 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

5 

99.75 

0 

0 

0 

0 

0 

1.06 

29.51 

0.98 

1.32 

1.38 

1.42 

1.31 

3.61 

37.95 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

1/27 

1/30 

1/30 

1/30 

1/30 

1/30 

1/30 

2/30 

0/30 

1/30 

0.74–4.4 

0.64–9.4 

0.61–14 

0.25–6.8 

0.23–6.8 

0.57–10 

0.25–7.9 

0.36–6.9 

0.19–11 

0.86–28 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

1.3 

2 

2.8 

2.4 

1.5 

2 

ND 

4.6 

ND 

3.2 

0.14 

0.22 

0.31 

0.27 

0.17 

0.22 

0 

0.68 

0 

0.36 

0.98 

1.58 

1.68 

1.22 

0.97 

1.72 

0.93 

1.83 

1.18 

3.4 

Total 2,3,7,8-CDD 

Total 2,3,7,8-CDF 

Total I-TEQDF 

Total TEQDF-WHO98 

-­ --

ND 

ND 

ND 

ND 

104.75 

19.8 

2.42 

2.33 

30.57 

2.37 

0.29 

0.27 

47.97 

15.49 

3.66 

4.28 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

4/27 

0/27 

1/30 

3/30 

13/30 

2/30 

1/30 

1/30 

2/30 

1/30 

1.2–8.8 

0.62–200 

0.84–11 

0.75–18 

6.2–57 

0.39–6.8 

0.64–25 

0.93–17 

0.36–19 

0.86–28 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

ND 

9.7 

ND 

1.7 

8.4 

99.75 

25 

20 

13 

4.6 

3.2 

1.23 

0 

0.19 

1.83 

29.51 

6.61 

2.22 

1.44 

0.68 

0.36 

2.61 

6.27 

1.93 

4.77 

37.95 

7.7 

4.72 

3.43 

2.41 

3.4 

Total CDD/CDF ND 185.35 44.07 75.19 

a

The overall means are the means of the individual POTW mean concentrations rather than the means of the

 individual sample concentrations. 

DL = Detection limit 

ND = Not detected 

-- = No information given 

Source: Memorandum dated March 21, 1996, from Lila Tang, California Regional Water Quality Control Board, to 

David Cleverly, U.S. EPA. 

8-79
 
Table 8-35. Effluent concentrations (pg/L) of CDDs/CDFs from publicly owned treatment works in 

8-80
 


Facility 

2,3,7,8­

TCDD 

2,3,7,8­

TCDF 

1,2,3,7,8­

PeCDD 

2,3,4,7,8­

PeCDF 

Total 

HxCDD 

Total 

HpCDD OCDD 

Waynesboro ND (0.17) 0.18 ND (0.2) ND (0.1) ND 3.5 13 Meridian 0.18 0.12 ND (0.16) ND (0.09) 1.3 7.6 58 

Pascagoula ND (0.13) 0.15 ND (0.15) 0.11 ND 0.82 3.6 

W. Biloxi 0.18 0.24 ND (0.15) 0.082 ND 0.9 4 NGulfport 0.16 0.24 ND (0.15) 0.094 ND 2.3 9.9 Laurel ND (0.18) 0.15 ND (0.23) ND (0.12) ND 2.9 38 NBrookhaven ND (0.18) 0.54 0.45 0.16 0.85 3.2 28 Natchez ND (0.16) 0.41 0.6 0.34 2.5 2.4 9.1 Picayune ND (0.22) 0.56 ND (0.27) ND (0.14) 6.5 38 120 

Picayunea ND (0.13) 0.54 ND (0.12) ND (0.07) 6 30 53 

Waveland ND (0.18) 17 0.22 0.66 ND 3 14 Corinth ND (0.15) 0.17 ND (0.16) ND (0.09) 0.77 2.7 18 New Augusta ND (0.1) 1.3 0.28 0.085 21 120 2,500 

Beaumont ND (0.1) 0.14 ND (0.13) 0.088 0.64 2.4 11 

Leaksville ND (0.12) 0.72 0.25 0.15 8.9 46 780 

McLain ND (0.06) ND (0.05) ND (0.10) ND (0.06) 2.5 14 200 

Hattiesburg South ND (0.16) ND (0.24) ND (0.24) ND (0.11) 1.2 4.5 59 

Hattiesburg North ND (0.19) 0.18 ND (0.26) ND (0.13) 0.96 9.1 73 

AVERAGE 0.17 1.42 0.36 0.2 4.43 16.3 221.76 

aBlind double. 
 

ND = Not detected (value in parenthesis is the detection limit). When calculating the average, NDs were not included.
 Source: Rappe et al. (1998).
 
population did not change drastically between 1987 and 1990, and assuming that the daily 

domestic wastewater treatment per person remained constant between 1987 and 2000, EPA 

estimates that approximately 131 billion L of wastewater were treated daily at POTWs in 1987. 

8.4.1.2.3. Emission Estimates. Multiplying the amount of wastewater treated by 365 days/yr by 

the overall mean TEQ concentrations reported by CRWQCB (memorandum dated March 21, 

1996, from Lila Tang, California Regional Water Quality Control Board, to David Cleverly, U.S. 

EPA) (0.27 pg TEQDF-WHO98/L [0.29 pg I-TEQDF/L]) yields annual TEQ release estimates of 

12.9 g TEQDF-WHO98 (13.9 g I-TEQDF), 12 g TEQDF-WHO98 (13 g I-TEQDF), and 14.6 g TEQDFWHO98 (15.7 g I-TEQDF) for 1987, 1995, and 2000, respectively. These estimates should be 

regarded as preliminary indications of possible emissions from this source. 

8.4.1.3. Sewage Sludge Land Disposal 

Sewage sludge is the solid, semi-solid, or liquid residue generated during the treatment of 

wastewater. During wastewater treatment, nutrients, pathogens, inorganic compounds (metals 

and trace elements), and organic compounds (CDDs/CDFs, PCBs, and surfactants) from the 

incoming wastewater are partitioned to the resulting sewage sludge (NRC, 2002). The sludge is 

either disposed of through methods such as incineration or landfill/surface disposal or 

beneficially used through methods such as land application. 

Sewage sludge that is applied to land is referred to as biosolids. In order to be applied to 

land, the biosolids must be treated to meet land application regulatory requirements (Federal 

Register, 1993a). Biosolids are often used for crop production, gardening, forestry, turf growth, 

and landscaping. Some other uses include strip mine and gravel pit reclamation and wetland 

restoration. Land application of biosolids is beneficial because it improves the physical and 

chemical properties of the soil needed for plant growth, it reduces the need for other disposal 

methods, and it reduces or eliminates the need for commercial fertilizers. Commercial fertilizers 

often have higher nutrient contents than do biosolids; therefore, the application of biosolids to 

land in lieu of commercial fertilizers may reduce the impacts of high levels of excess nutrients 

entering the environment (U.S. EPA, 1999b). 

8.4.1.3.1. Emissions data. EPA conducted the National Sewage Sludge Survey in 1988 and 

1989 to obtain national data on sewage sludge quality and management. As part of this survey, 

EPA analyzed sludges from 174 POTWs that employed at least secondary wastewater treatment 

for more than 400 analytes, including CDDs/CDFs. Although sludges from only 16% of the 

POTWs had detectable levels of 2,3,7,8-TCDD, all sludges had detectable levels of at least one 

CDD/CDF congener (U.S. EPA, 1996e). I-TEQDF concentrations as high as 1,820 ng/kg dry 

8-81
 

weight were measured. The congener-specific results of the survey are presented in Table 8-36. 

If all nondetect values found in the study are assumed to be zero, then the mean and median ITEQDF concentrations of the sludges from the 174 POTWs are 50 and 11.2 ng/kg (dry-weight 

basis), respectively. If the nondetect values are set equal to the DL, then the mean and median 

I-TEQDF concentrations are 86 and 50.4 ng/kg, respectively (U.S. EPA, 1996e; Rubin and White, 

1992). 

Table 8-36. CDD/CDF concentrations (ng/kg) measured in the 1998/1999 

National Sewage Sludge Surveya 

Congener 

Percent 

detected 

Maximum 

concentration 

detected 

Median concentration Mean concentration 

Nondetect 

set to 

detection 

limit 

Nondetect 

set to 

zero 

Nondetect 

set to 

detection 

limit 

Nondetect 

set to 

zero 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

16 

18 

25 

49 

39 

98 

100 

116 

736 

737 

737 

737 

52,500 

905,000 

6.86 

9.84 

22.5 

27.3 

28 

335 

3,320 

0 

0 

0 

0 

0 

335 

3,320 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

65 

22 

26 

43 

35 

16 

27 

71 

26 

80 

337 

736 

736 

1,500 

737 

1,260 

737 

7,100 

842 

69,500 

17 

9.6 

10.4 

28 

18 

18 

18 

57 

23 

110 

3.9 

0 

0 

0 

0 

0 

0 

36 

0 

80 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

Total I-TEQDF 1,820 50.4 11.2 86b 50b 

Total 2,3,7,8-CDD/CDF -­ -­ -­ -­ -­

a

For publicly owned treatment works with multiple samples, the pollutant concentrations were averaged before the 

summary statistics presented in the table were calculated. b

Values presented by Rubin and White (1992) for 175 rather than 174 publicly owned treatment works. 

-- = No information given 

Source: U.S. EPA (1996e). 

8-82
 

Green et al. (1995) and Cramer et al. (1995) reported the results of analyses of 99 samples 

of sewage sludge collected from wastewater treatment plants across the United States during the 

summer of 1994 as part of the 1994/1995 Association of Metropolitan Sewerage Agencies 

(AMSA) survey. These data are summarized in Table 8-37. To calculate average results in units 

of TEQ, Green et al. averaged the results from all samples collected from the same facility to 

ensure that the results were not biased toward the concentrations found at facilities from which 

more than one sample was collected. Also, eight samples were excluded from the calculation of 

the overall TEQ averages because it was unclear whether they were duplicate samples from other 

POTWs. POTW average TEQ concentrations were calculated for 74 POTWs. If all nondetect 

values are assumed to be zero, then the overall study mean and median I-TEQDF concentrations 

were 47.7 and 33.4 ng I-TEQDF/kg (dry weight basis), respectively (standard deviation of 44.7 ng 

I-TEQDF/kg). The corresponding mean and median TEQDF-WHO98 concentrations were 36.3 and 

25.5 ng/kg, respectively (standard deviation, 38.6). 

The mean and median results reported by Green et al. (1995) and Cramer et al. (1995) are 

very similar in terms of total TEQ to those reported by EPA for samples collected 5 years earlier 

(U.S. EPA, 1996e; Rubin and White, 1992). The predominant congeners in both data sets are the 

octa- and hepta-CDDs and CDFs. Although not present at high concentrations, 2,3,7,8-TCDF 

was commonly detected. 

In addition to effluents, Rappe et al. (1998) analyzed the levels of CDDs/CDFs in 

municipal sewage sludge from the 17 POTWs in Mississippi. Table 8-38 presents the 

concentrations of dioxins measured in the sewage sludge samples and total TEQ emission factors 

reported by Rappe et al. Concentrations were only congener specific for 2,3,7,8-TCDD; 2,3,7,8­

TCDF; 1,2,3,7,8-PeCDD; 2,3,4,7,8-PeCDF; OCDD; and OCDF. Also provided were 

concentrations for total HxCDD and total HpCDD. The TEQ emission factors (assuming 

nondetects equal to one-half the DL) ranged from 2.26 to 1,270 ng I-TEQDF/kg. The 

predominant congeners in all samples were the octa- and hepta-CDDs. The sludge with the 

highest concentrations of octa- and hepta-CDDs was from the Picayune POTW, which receives 

industrial inputs, including effluents from wood treatment facilities that likely contain PCP. In 

general, the sludges with the lowest TEQ values were from the facilities that do not receive 

effluent from industrial facilities. Additionally, the samples with the two lowest TEQ values 

were from facilities that do not use free chlorine as a disinfectant. 

In 1999, sewage sludge samples from a POTW in Ohio were collected and analyzed for 

CDDs/CDFs (U.S. EPA, 2000b). The facility, which accepts both domestic and industrial 

wastewater, employs secondary wastewater technology. Assuming nondetects are zero, the mean 

TEQ emission factor was 21.9 ng TEQDF-WHO98/kg (dry-weight basis). These results are 

presented in Table 8-39. 

8-83
 
Table 8-37. CDD/CDF concentrations (ng/kg) measured in 99 sludges collected from U.S. publicly 

treatment works (POTWs) during 1994 

8-84
 


Congener 

Percent 

detected 

Maximum 

concentration 

detected 

Median concentration Mean conc

Nondetect 

set to 

detection 

limit 

Nondetect 

set to 

zero 

Nondetect 

set to 

detection limita 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

40 

23 34 

87 

64 

98 

99 

12.3 

37.5 45.6 

130 

88.8 

5,380 

65,500 

1.95 

8.23 5.25 

25.6 

12.3 

642 

6,630 

0 

0 0 

24.7 

9.48 

642 

6,630 

2.72 (2.4) 

10.9 (7.8) 11.1 (8.13) 

33.8 (27.6) 

20.2 (17.7) 

981 (977) 

11,890 (12,540) 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-HpCDF 

OCDF 

76 

21 

42 

48 

17 

4 

35 

64 31 

93 

156 

60.3 

155 

170 

200 

115 

356 

1,460 

213 

11,200 

7.53 

7.91 

9.7 

11.5 

14 

7.53 

9.85 

91.7 11.7 

286 

6.28 

0 

0 

0 

0 

0 

0 

31.8 

0 

281 

12.8 (19.6) 

10.7 (11.3) 

15.7 (19.8) 

20.4 (25.3) 

30.4 (53.6) 

11.1 (13.6) 

21.8 (40.4) 

223 (271) 

27.1 (34.8) 

786 (1,503) 

Average I-TEQDF (facility basis)b 246 49.6 33.4 64.5 (50.1) 

Total 2,3,7,8-CDD/CDF 85,279.5 7,782.8 7,625.3 14,108.7 (15,573) 

Average TEQDF-WHO98 (facility basis)b 44.6 25.5 57.2 (44.4) 

aValue in parenthesis is the standard deviation. b For POTWs with multiple samples, the sample TEQ concentrations were averaged to POTW averages before calculation of the total T

values presented in the table. A total of 74 POTW average concentrations were used in the calculations. In addition, the following samnot included in the averaging because, according to Green et al. (1995), it was not possible to determine whether they were duplicate oother POTWs: 87, 88, 89, 90, 91, 97, 98, and 106. 

Sources: Green et al. (1995); Cramer et al. (1995). 
Table 8-38. Concentrations of CDDs/CDFs (ng/kg dry matter) in sewage sludge from publicly own

works in Mississippi 

8-85
 


Facility 

2,3,7,8­

TCDD 

2,3,7,8­

TCDF 

1,2,3,7,8­

PeCDD 

2,3,4,7,8­

PeCDF 

Total 

HxCDD 

Total 

HpCDD OCDD 

Waynesboro 2.1 2.9 3.5 1.4 85 920 7,400 

Meridian ND (0.06) 2.1 6.4 2.8 10 100 7,400 

Pascagoula 2 3.6 5.3 3.5 170 970 4,300 

W. Biloxi 0.84 2.4 3.2 1.3 78 280 1,800 

Gulfport 1.9 9.1 9.5 3.4 200 1,100 7,700 Laurel 0.17 0.3 0.37 0.25 22 160 2,700 Brookhaven 2 2.5 11 2.5 130 1,400 9,300 

Natchez ND (0.58) 8.3 8.4 ND (1.5) 270 1,100 6,800 Picayune 5.3 69 74 24 17,000 250,000 480,000 Picayunea 4.1 66 60 17 16,000 210,000 420,000 

Waveland 1.6 2.6 5.1 1.9 130 580 3,500 Corinth 0.3 1.8 0.97 0.93 42 230 3,300 New Augusta ND (0.13) 0.17 0.15 0.094 21 140 1,400 

Beaumont 0.17 0.67 0.78 0.37 59 470 1,900 Leaksville ND (0.051) 0.14 0.32 0.11 16 92 560 

McLain 0.076 0.17 0.11 0.031 39 140 2,600 

Hattiesburg S 1 1.1 9.1 2.2 170 1,3000 4,400 Hattiesburg N ND (0.035) 1.7 4 2 310 3,600 27,000 

Average 1.7 9.7 11.2 3.8 1,930.7 26,904.6 55,114.4 

aBlind double. 
 

ND = Not detected (value in parenthesis is the detection limit). When calculating the average, NDs were not included.
 Source: Rappe et al. (1998).
 
Table 8-39. CDD/CDF median concentrations (ng/kg) measured in 1999 

from a publicly owned treatment works facility in Ohio 

Congener Nondetect set to zero 

Nondetect set to 

1⁄2 detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND 

ND 

2.67 

21.33 

30.33 

298 

2,963 

0.0018 

0.0082 

2.67 

21.33 

30.33 

298 

2,963 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

26.67 

4.33 

10 

21 

5.33 

ND 

9 

171 

ND 

364.67 

26.67 

4.34 

10 

21 

5.33 

0.0033 

9 

171 

0.01 

364.67 

Average total TEQDF-WHO98 21.87 21.88 

ND = Not detected 

Source: U.S. EPA (2000b). 

In 2000 and 2001, AMSA conducted another survey of dioxin-like compounds in sewage 

sludge (Alvarado et al., 2001). A total of 200 sewage sludge samples were collected from 171 

POTWs located in 31 states. Assuming nondetects are zero, TEQ emission factors range from 

0.08 to 3,578.61 ng TEQDF-WHO98/kg. The mean and median TEQ emission factors were 34.5 

and 11.79 ng TEQDF-WHO98/kg, respectively. 

EPA conducted another National Sewage Sludge Survey to characterize the dioxin and 

dioxin-like equivalence levels in biosolids produced by 6,857 POTWs operating in the United 

States in 2001 (U.S. EPA, 2002d). Samples were collected from 94 POTWs using secondary or 

higher treatment practices. All facilities had been sampled previously as part of the 1988/1989 

National Sewage Sludge Survey. The overall mean and median TEQDF-WHO98 concentrations 

were 75 and 15 ng/kg, respectively. However, when the data were weighted using the daily 

influent wastewater flow rates (i.e., the number of facilities with wastewater flow rate 

8-86
 

>100 Mg/day, >10 but #100 Mg/day, >1 but #10 Mg/day, and #1 Mg/day), the overall mean and 

median TEQDF-WHO98 concentrations were 21.7 and 15.5 ng/kg, respectively. These data are 

summarized in Table 8-40. 

Table 8-40. CDD/CDF mean concentrations (ng/kg) measured in the 2001 

National Sewage Sludge Survey 

Congener Nondetect set to zero 

Nondetect set to

 1⁄2 detection limit 

2,3,7,8-TCDD 

1,2,3,7,8-PeCD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

1.41 

5.76 

11.8 

21.3 

3.6 

492 

6,780 

1.1 

4.57 

7.49 

15.1 

2.22 

273 

2,730 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

3.11 

2.61 

6.03 

1.37 

0.27 

5.21 

5.5 

9.13 

167 

802 

2.3 

1.5 

2.8 

1 

0 

2.6 

3.36 

2.8 

88.2 

279 

Average total TEQDF-WHO98 21.7 15.5 

Source: U.S. EPA (2002d). 

The CDD/CDF concentrations and congener group patterns observed in the U.S. surveys 

are similar to those reported for sewage sludges in several other Western countries. Stuart et al. 

(1993) reported mean CDD/CDF concentrations of 23.3 ng I-TEQDF/kg (dry weight) for three 

sludges from rural areas, 42.3 ng I-TEQDF/kg for six sludges from light industry/domestic areas, 

and 52.8 ng I-TEQDF/kg for six sludges from industrial/domestic areas collected during 

1991–1992 in England and Wales. Näf et al. (1990) reported concentrations ranging from 31 to 

40 ng I-TEQDF/kg (dry weight) in primary and digested sludges collected from the POTW in 

Stockholm, Sweden, during 1989. Gobran et al. (1995) reported an average concentration of 

15.7 ng I-TEQDF/kg in anaerobically digested sludges from an industrial/domestic POTW in 

Ontario, Canada. In all three studies, the congener group concentrations increased with 

8-87
 

2,3,7,8-TCDD

1,2,3,7,8-PeCDD

1,2,3,4,7,8-HxCDD

1,2,3,6,7,8-HxCDD1,2,3,7,8,9-HxCDD

1,2,3,4,6,7,8-HpCDD

OCDD

2,3,7,8-TCDF

1,2,3,7,8-PeCDF2,3,4,7,8-PeCDF

1,2,3,4,7,8-HxCDF1,2,3,6,7,8-HxCDF

1,2,3,7,8,9-HxCDF

2,3,4,6,7,8-HxCDF

1,2,3,4,6,7,8-HpCDF

1,2,3,4,7,8,9-HpCDF

OCDF

0%

10%

20%

30%

40%

50%

60%

70%

80%

90%

100%

Percent of Total CDDs and CDFs 

Congener

Figure 8-6. Congener profiles for sewage sludge (created from data in Table 

8-38). 

increasing degrees of chlorination, with OCDD the dominant congener. Figure 8-6 presents 

congener profiles, using the mean concentrations reported by Green et al. (1995). 

Because the mean I-TEQDF concentration values reported in the 1988/1989 sewage sludge 

survey (U.S. EPA, 1996e) and the 1995 survey (Green et al., 1995; Cramer et al., 1995) were 

very similar, the estimated amounts of TEQs that may have been present in sewage sludge and 

released to the environment in 1987 and 1995 were assumed to be the same. These values were 

estimated using the average (49 ng I-TEQDF/kg) of the mean I-TEQDF concentration values 

(nondetects equal to DLs) reported by U.S. EPA (1996e) (50 ng I-TEQDF/kg) and by Green et al. 

(1995) and Cramer et al. (1995) (36.3 ng TEQDF-WHO98/kg [47.7 ng I-TEQDF/kg]). Therefore, 

the overall average mean emission factor for reference years 1987 and 1995 is 36.3 ng TEQDFWHO98/kg (48.9 ng I-TEQDF/kg). The emission factor of 21.7 ng TEQDF-WHO98/kg, as 

calculated from the 2001 survey, appears to be the most reasonable TEQ emission factor estimate 

for reference year 2000 because this estimate is nationally weighted on the basis of wastewater 

flow rates of POTWs operating in the United States in 2001. 

8.4.1.3.2. Activity level information. According to the results of its 1988/1989 National 

Sewage Sludge Survey, EPA estimated that approximately 5.4 million dry metric tons of sewage 

8-88
 

sludge were generated in 1989 (Federal Register, 1993a). EPA also used the results of the 1984 

to 1996 Clean Water Needs Surveys to estimate that 6.3 million dry metric tons of sewage sludge 

were generated in 1998. Because estimates for reference years 1987 and 1995 are not available, 

the 1989 and 1998 activity level estimates were used for 1987 and 1995, respectively. Tables 

8-41 and 8-42 list the volumes, by use and disposal practices, of sludge disposed of annually for 

reference years 1987 and 1995, respectively. 

U.S. EPA (1999b) estimated that 6.6 million dry metric tons of sewage sludge would be 

generated in 2000. Table 8-43 lists the volumes, by use and disposal practices, of sludge 

disposed of annually for reference year 2000. Similarly, the National Research Council (NRC) 

analyzed the amount of biosolids being applied to land in 2002 (NRC, 2002). Citing 2001 data 

(unpublished) from the Wisconsin Department of National Resources, the NRC estimated that 

approximately 8,650 of the 16,000 POTWs operating in the United States generated sewage 

sludge requiring use or disposal. Using data from 37 states, the investigators estimated that 

5,900 of these sewage sludge generators either land applied or publicly distributed more than 3.4 

million dry tons of biosolids annually. The volume of biosolids, by use and disposal practices, is 

presented in Table 8-44. The volume of biosolids and their distribution among the various 

categories estimated by the NRC are very close to those estimated by EPA. 

8.4.1.3.3. Emission estimates. Annual potential releases of CDDs/CDFs were determined by 

multiplying the mean total TEQ concentrations by the sludge volumes generated. The mean total 

TEQ concentrations in sewage sludges for reference years 1987, 1995, and 2000 are reported in 

Tables 8-41, 8-42, and 8-43, respectively. For 1987, the total annual potential release from 

nonincinerated sludges was 151 g TEQDF-WHO98. Of this amount, 2.6 g TEQDF-WHO98 (3.5 g ITEQDF) entered commerce as a product for distribution and marketing and 76.6 g TEQDF-WHO98 

(103 g I-TEQDF) was applied to land. The remaining 71.8 g TEQDF-WHO98 did not result in an 

environmental release because it was sent to RCRA Subtitle D landfills or disposal sites. For 

1995, the total annual potential release from nonincinerated sludges was 178 g TEQDF-WHO98. 

Of this amount, 3 g TEQDF-WHO98 (4 g I-TEQDF) entered commerce as a product for distribution 

and marketing and 116.1 g TEQDF-WHO98 (156.5 g I-TEQDF) was applied to land. The remaining 

58.9 g TEQDF-WHO98 did not result in an environmental release because it was sent to RCRA 

Subtitle D landfills or disposal sites. For 2000, the total annual release of nonincinerated sludges 

was 111 g TEQDF-WHO98. Of this amount, 1.9 g TEQDF-WHO98 (1.9 g I-TEQDF) entered 

commerce as a product for distribution and marketing and 78.2 g TEQDF-WHO98 (78.2 g ITEQDF) was applied to land. The remaining 30.9 g TEQDF-WHO98 did not result in an 

environmental release because it was sent to RCRA Subtitle D landfills or disposal sites. 

8-89
 
Table 8-41. Quantity of sewage sludge disposed of and potential dioxin TEQ 

releases from primary, secondary, and advanced treatment publicly owned 

treatment works for reference year 1987 

Use/disposal practice 

Volume 

disposed of 

(1,000 dry 

metric 

tons/yr) 

Percent of 

total volume 

Potential dioxin releasea 

(g TEQ/yr) 

I-TEQDF 

TEQDF ­

WHO98 

Land application 1,714 32b 84 62.2 

Distribution and marketing 71 1.3 3.5 2.6 

Surface disposal site/other 396 7.4 19.4 14.4 

Sewage sludge landfill 157 2.9 7.7 5.7 

Co-disposal landfillsc 1,819 33.9 89.1 66 

Sludge incinerators and 

co-incineratorsd 

865 16.1 e e 

Ocean disposal (336)f (6.3)f f f 

TOTAL 5,358 100 203.7 150.9 a

Potential dioxin TEQ release for nonincinerated sludges was estimated by multiplying the sludge volume generated 

(second column) by the average of the mean I-TEQDF concentrations in sludge reported by Rubin and White (1992) 

(i.e., 50 ng/kg dry weight) and Green et al. (1995) and Cramer et al. (1995) (47.7 ng/kg). The calculations of 

TEQDF-WHO98 used the mean concentration of 36.3 ng TEQDF-WHO98/kg for the results reported by Green et al. 

(1995) and Cramer et al. (1995). b

Includes 21.9% applied to agricultural land, 2.8% applied as compost, 0.6% applied to forestry land, 3.1% applied 

to “public contact” land, 1.2% applied to reclamation sites, and 2.4% applied in undefined settings. c

Landfills used for disposal of sewage sludge and solid waste residuals. d

Co-incinerators treat sewage sludge in combination with other combustible waste materials. e

See Section 3.5 for estimates of CDD/CDF releases to air from sewage sludge incinerators. f

The Ocean Dumping Ban Act of 1988 generally prohibited the dumping of sewage sludge into the ocean after 

December 31, 1991. Ocean dumping of sewage sludge ended in June 1992 (Federal Register, 1993a). The current 

method of disposal of the 336,000 metric tons of sewage sludge that were disposed of in the oceans in 1988 has not 

been determined. 

These release estimates are assigned a high confidence rating for both the production and 

the emission factor estimates. The high rating is based on the judgment that the 174 facilities 

tested as part of the 1988/1989 National Sewage Sludge Survey (U.S. EPA, 1996e), the 74 

facilities tested as part of the 1994/1995 AMSA Survey (Green et al., 1995 and Cramer et al., 

1995), and the 94 facilities tested as part of the 2001 National Sewage Sludge Survey (U.S. EPA, 

2002d) were reasonably representative of the variability in POTW technologies and sewage 

characteristics nationwide. 

8-90
 

Table 8-42. Quantity of sewage sludge disposed of and potential dioxin TEQ 

releases from primary, secondary, and advanced treatment publicly owned 

treatment works for reference year 1995 

Use/disposal practice 

Volume disposed 

of (1,000 dry 

metric tons/yr) 

Percent of 

total volume 

Potential dioxin releasea 

(g TEQ/yr) 

I-TEQDF 

TEQDF ­

WHO98 

Land applicationb 2,500 40 122.3 90.7 

Advanced treatmentc 700 11 34.2 25.4 

Other beneficial used 500 8 24.5 18.2 

Surface disposal/landfill 1,100 17 53.8 39.9 

Incineration 1,400 22 e e 

Other disposal method 100 2 4.9 3.6 

TOTAL 6,300 100 239.7 177.8 a

Potential dioxin TEQ release for nonincinerated sludges was estimated by multiplying the sludge volume generated
 


(second column) by the average of the mean I-TEQDF concentrations in sludge reported by Rubin and White (1992)
 


(50 ng/kg dry weight) and Green et al. (1995) and Cramer et al. (1995) (47.7 ng/kg). The calculations of TEQDFWHO98 used the mean concentration of 36.3 ng TEQDF-WHO98/kg for the results reported by Green et al. (1995)
 


and Cramer et al. (1995).
 
 b

Without further processing or stabilization, such as composting.
 
 c

Such as composting.
 
 d

EPA assumes that this category includes distribution and marketing (i.e., sale or give-away of sludge for use in
 


home gardens). Based on the 1988 National Sewage Sludge Survey and 1988 Needs Survey, approximately 1.3%
 


of the total volume of sewage disposed is distributed and marketed (Federal Register, 1993a). Therefore, it is
 


estimated that 3 g TEQDF-WHO98 (4 g I-TEQDF) were released through distribution and marketing in 1995.
 
 e

See Section 3.5 for estimates of CDD/CDF releases to air from sewage sludge incinerators. 

Sources: Federal Register (1990, 1993a). 

8.4.2. Drinking Water Treatment Plants 

There is no strong evidence that chlorination of water for drinking purposes results in the 

formation of CDDs/CDFs. Few surveys of CDD/CDF content in finished drinking water have 

been conducted. Those that have been published have only rarely reported the presence of any 

CDDs/CDFs, even at low picogram per liter DLs, and in those cases, CDDs/CDFs were also 

present in the untreated water. 

Rappe et al. (1989b) reported the formation of tetra- through octa-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 I-TEQDF/L, respectively. 

No CDDs were detected at DLs ranging from 1 to 5 pg/L. However, the water samples were 

8-91
 

Table 8-43. Quantity of sewage sludge disposed of and potential dioxin TEQ 

releases from primary, secondary, and advanced treatment publicly owned 

treatment works for reference year 2000 

Use/disposal practice 

Volume disposed of 

(1,000 dry metric 

tons/yr) 

Percent of 

total volume 

Potential dioxin releasea 

(g TEQ/yr) 

TEQDF-WHO98 

Land applicationb 2,800 42 60.80 

Advanced treatmentc 800 12 17.40 

Other beneficial used 500 8 10.90 

Surface disposal/landfill 900 14 19.50 

Incineration 1,500 23 e 

Other disposal method 100 2 2.17 

TOTAL 6,600 100 111.00 a

Potential dioxin TEQ release for nonincinerated sludges was estimated by multiplying the sludge volume generated 

(second column) by the average of the mean TEQDF-WHO98 concentrations in sludge reported in U.S. EPA (2002d). b

Without further processing or stabilization, such as composting. c

Such as composting. d

EPA assumes that this category includes distribution and marketing (sale or give-away of sludge for use in home 

gardens). Based on the 1988 National Sewage Sludge Survey and 1988 Needs Survey, approximately 1.3% of the 

total volume of sewage disposed is distributed and marketed (Federal Register, 1993a). Therefore, it is estimated 

that 1.9 g TEQDF-WHO98 were released through distribution and marketing in 2000. e

See Section 3.5 for estimates of CDD/CDF releases to air from sewage sludge incinerators. 

Table 8-44. Biosolids disposal practices for reference year 2000 

Use/disposal practice 

Volume disposed of (1,000 dry 

metric tons/yr) 

Percent of 

total volume 

Land application 3,100 61 

Surface disposal/landfill 940 18 

Incineration 1,000 20 

Other 64 1 

TOTAL 5,104 100 

Source: NRC (2002). 

chlorinated at a dosage rate of 300 mg/L, which is considerably higher (by one to two orders of 

magnitude) than the range of dosage rates typically used to disinfect drinking water. The authors 

hypothesized that the CDFs or their precursors were present in chlorine gas. 

8-92
 
Rappe et al. (1990c) analyzed a 1,500-L sample of drinking water from a municipal 

drinking water treatment plant in Sweden. Although the untreated water was not analyzed, a 

sludge sample from the same facility was analyzed. The large sample volume enabled DLs on 

the order of 0.001 pg/L. The TEQ contents of the water and the sludge were 0.0029 pg ITEQDF/L and 1.4 ng/kg, respectively. The congener patterns of the drinking water and sludge 

sample were very similar, suggesting that the CDDs/CDFs detected in the finished water were 

present in the untreated water. 

8.4.3. Soaps and Detergents 

As discussed in Section 8.4.1, CDDs/CDFs were detected in nearly all sewage sludges 

tested, whether the sludges were obtained from industrialized areas or from rural areas. Because 

of the ubiquitous presence of CDDs/CDFs in sewage sludge, several studies have been conducted 

to determine their source(s). A logical category of products to test, because of their widespread 

use, is detergents, particularly those that contain or release chlorine during use (i.e., hypochloritecontaining and dichloroisocyanuric acid-containing detergents). The results of studies conducted 

to date, summarized below, indicate that CDDs/CDFs are not formed during use of chlorine-free 

detergents, chlorine-containing or chlorine-releasing detergents, or chlorine bleach during 

household bleaching operations. 

Sweden’s Office of Nature Conservancy (1991) reported that the results of a preliminary 

study conducted at one household indicated that CDDs/CDFs may be formed during use of 

dichloroisocyanurate-containing dishwasher detergents. A more extensive main study with 

multiple runs was then conducted using standardized food, dishes, cutlery, and other household 

items. Testing of laundry washing and fabric bleaching and actual testing of the CDD/CDF 

content of detergents was also performed. The study examined (1) hypochlorite- and 

dichloroisocyanurate-containing dishwasher detergents, (2) sodium hypochlorite-based bleach 

(4.4% NaOCl) in various combinations with and without laundry detergent, and (3) sodium 

hypochlorite-based bleach, used at a high enough concentration to effect bleaching of a pair of 

imported blue jeans. 

CDDs/CDFs were not detected in either the chlorine-free detergent or the detergent with 

hypochlorite; 0.6 pg TEQ/g were detected in the detergent containing dichloroisocyanurate. The 

results of all dishwasher and laundry washing machine tests showed very low levels of 

CDDs/CDFs, often nondetected values. There was no significant difference between the controls 

and the test samples; in fact, the control samples had a higher TEQ content than did some of the 

test samples. The drainwater from the dishwasher tests contained <1 to <3 pg I-TEQDF/L (the 

water-only control sample contained <2.8 pg I-TEQDF/L). The CDD/CDF content of the laundry 

8-93
 

drainwater samples ranged from <1.1 to <4.6 pg I-TEQDF/L (the water-only control sample 

contained <4.4 pg I-TEQDF/L). 

Thus, under the study’s test conditions, CDDs/CDFs were not formed during dishwashing 

or laundry washing or during bleaching with hypochlorite-containing bleach. No definitive 

reason could be found for the difference in results between the preliminary study and the main 

study for dishwashing with dichloroisocyanurate-containing detergents. The authors of the study 

suggested that differences in the foods used and the prewashing procedures employed in the two 

studies were the likely causes of the variation in the results. 

Rappe et al. (1990b) analyzed a sample of a Swedish commercial soft soap, a sample of 

tall oil, and a sample of tall resin for CDD/CDF content. Tall oil and tall resin, by-products of 

the pulping industry, are the starting materials for the production of soft liquid soap. Crude tall 

oil, collected after the Kraft pulping process, is distilled under reduced pressure at temperatures 

of up to 280 to 290°C, yielding tall oil and tall resin. The measured TEQ content of the liquid 

soap was found to be 0.647 ng TEQDF-WHO98/L (0.447 ng I-TEQDF/L). PeCDDs were the 

dominant congener group, followed by HpCDDs, HxCDDs, PeCDFs, and OCDD, with some 

tetra-CDFs and CDDs also present. The TEQ contents of the tall oil (12 ng TEQDF-WHO98/kg 

[9.4 ng I-TEQDF/kg]) and tall resin (196 ng TEQDF-WHO98/kg [200 ng I-TEFDF/kg]) were 

significantly higher than the level found in the liquid soap. The tall oil contained primarily tetraand penta-CDDs and CDFs, whereas the tall resin contained primarily HpCDDs, HxCDDs, and 

OCDD. The investigators compared the congener patterns of the three samples and noted that 

although the absolute values for the tetra- and penta-CDDs and CDFs differed among the tall oil, 

tall resin, and liquid soap samples, the same congeners were present. The congener patterns for 

the higher-chlorinated congeners were very similar. Table 8-45 presents the results of the study. 

In 1987, 118 million L of liquid household soaps were shipped in the United States (U.S. 

DOC, 1990b); shipment quantity data are not available in the 1992 U.S. Economic Census (U.S. 

DOC, 1996). Because only one sample of liquid soap has been analyzed for CDD/CDF content 

(Rappe et al., 1990b), no estimate of environmental release can be made. 

8.4.4. Textile Manufacturing and Dry Cleaning 

As discussed in Section 8.4.1, CDDs/CDFs have been detected in almost all sewage 

sludges tested, whether they were obtained from industrialized areas or rural areas. To determine 

whether textile fabric or fabric finishing processes could account for the observed CDDs/CDFs, 

several studies were conducted in Germany. These studies, summarized in the following 

paragraphs, indicate that some finished textile products do contain detectable levels of 

CDDs/CDFs and that they can be released from the textile during laundering or dry cleaning; 

8-94
 

Table 8-45. CDD/CDF concentrations in Swedish liquid soap, tall oil, and 

tall resin 

Congener/congener group 

Liquid soap 

(ng/L) 

Tall oil 

(ng/kg) 

Tall resin 

(ng/kg) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

ND (0.009) 

0.4 

ND (0.02) 

0.32 

0.18 

1.9 

1 

3.6 

5.3 

ND (2) 

ND (2) 

ND (2) 

ND (1) 

5.3 

ND (1) 

3.1 

ND (4) 

810 

500 

5,900 

6,000 

2,3,7,8-TCDF 

1,2,3,4,8-/1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8/9-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.62 

0.29 

0.2 

0.013 

ND (0.004) 

ND (0.004) 

ND (0.004) 

ND (0.005) 

ND (0.01) 

NA 

17 

4.2 

1.9 

1.4 

0.7 

ND (0.7) 

ND (0.5) 

ND (0.8) 

ND (2) 

NA 

ND (2) 

ND (0.4) 

ND (0.5) 

24 

-­

ND (1) 

ND (0.7) 

10 

9.0 

NA 

Total 2,3,7,8-CDDa 

Total 2,3,7,8-CDFa 

Total I-TEQDF 

a 

Total TEQDF-WHO98 

a 

3.8 

1.123 

0.447 

0.647 

14.2 

25.2 

9.4 

12 

13,213.1 

43 

200 

196 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

0.12 

15 

3.4 

3.6 

1 

1 

1.3 

0.15 

ND (0.01) 

NA 

31 

380 

3.3 

ND (1) 

5.3 

26 

41 

4.9 

ND (2) 

NA 

ND (1) 

25 

6,800 

11,000 

6,000 

ND (2) 

ND (0.5) 

56 

19 

NA 

Total CDD/CDFa 25.57 491.5 23,900 a

Calculations assume nondetect values were zero. 

NA = Not analyzed 

ND = Not detected (value in parenthesis is the detection limit) 

-- = No information given 

Source: Rappe et al. (1990b). 

8-95
 

however, textile finishing processes are typically not sources of CDD/CDF formation. Rather, 

the use of CDD/CDF-containing dyes and pigments and the use in some countries of PCP to treat 

unfinished cotton appear to be the sources of the detected CDDs/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 that subjected 

one of the cotton materials 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. On 

the basis of the concentrations found, the authors concluded that neither unfinished new fabrics 

nor common cotton finishing processes could explain the CDD/CDF levels found in laundry 

wastewater. 

Fuchs et al. (1990) reported that the dry-cleaning solvent redistillation residues collected 

from 12 commercial and industrial dry-cleaning operations contained considerable amounts of 

CDDs/CDFs. The reported I-TEQDF content ranged from 131 to 2,834 ng/kg, with the dominant 

congeners being OCDD and HpCDD. Towara et al. (1992) demonstrated that neither the use of 

chlorine-free solvents nor variation of the dry-cleaning process parameters lowered the 

CDD/CDF content of the residues. 

Umlauf et al. (1993) conducted a study to characterize the mass balance of CDDs/CDFs 

in the dry-cleaning process. The soiled clothes (containing 16 pg total CDDs/CDFs per kg) 

accounted for 99.996% of the CDD/CDF input. Input of CDDs/CDFs from indoor air containing 

0.194 pg/m3 accounted for the remainder (0.004%). The dry-cleaning process removed 82.435% 

of the CDDs/CDFs in the soiled clothing. Most of the input CDDs/CDFs (82.264%) were found 

in the solvent distillation residues. Air emissions (at 0.041 pg/m3) accounted for 0.0008% of the 

total input, which was less than the input from indoor air. The fluff (at a concentration of 

36 ng/kg) accounted for 0.1697%, and water effluent (at a concentration of 0.07 pg/L) accounted 

for 0.0000054%. 

Horstmann and McLachlan (1994a, b, 1995) analyzed 35 new textile samples (primarily 

cotton products) obtained in Germany for CDDs/CDFs. Low levels were found in most cases 

(total CDD/CDF <50 ng/kg). The dominant congeners were OCDD and HpCDD. 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 

8-96
 
CDDs/CDFs was not a textile finishing process, because a process source would have resulted in 

a more consistent level of contamination. Klasmeier and McLachlan (1995) subsequently 

analyzed 68 new textile products obtained in Germany for OCDD and OCDF. Most samples had 

nondetectable levels (42 samples <60 ng/kg). Only four samples had levels exceeding 500 ng/kg. 

Horstmann and McLachlan (1994a, b) reported finding two different congener group 

patterns in the more contaminated of the 35 textile products. One pattern agreed with the 

congener pattern for PCP reported by Hagenmaier and Brunner (1987), whereas the other pattern 

was similar to that reported by Remmers et al. (1992) for chloranil-based dyes. The authors 

hypothesized that the use of PCP to preserve cotton, particularly when it is randomly strewn on 

bales of cotton as a preservative during sea transport, was the likely source of the high levels 

occasionally observed. Although the use of PCP for nonwood uses was prohibited in the United 

States in 1987 (see Section 8.3.8), PCP is still used in developing countries, especially to 

preserve cotton during sea transport (Horstmann and McLachlan, 1994a). 

Horstmann and McLachlan (1994a, b) conducted additional experiments that 

demonstrated that the small percentage of clothing items with high CDD/CDF levels could be 

responsible for the quantity of CDDs/CDFs observed in household wastewater (see Section 

8.4.1.1). 

8-97
 

9. INDICATIONS OF POSSIBLE NATURAL SOURCES OF CDDs/CDFs 

Numerous laboratory and field research studies have demonstrated that biochemical and 

photolytic formation of CDDs/CDFs from chlorophenol precursors is possible. In addition, 

under certain conditions, some CDDs/CDFs can be biodegraded to form less chlorinated (and 

possibly more toxic) CDDs/CDFs. Both of these mechanisms are discussed in this chapter; 

however, the extent to which CDDs/CDFs are formed by either mechanism in the environment is 

not known at present. The potential for releases of CDDs/CDFs from the application of animal 

manure to farmland and the mining and use of ball clay is also discussed. 

9.1. BIOLOGICAL SOURCES OF CDDs/CDFs 

9.1.1. Biotransformation of Chlorophenols 

Biochemical formation of CDDs/CDFs—particularly the higher-chlorinated congeners— 

from chlorophenol precursors is possible, as indicated in laboratory studies with solutions of 

trichlorophenols and pentachlorophenol (PCP) in the presence of peroxidase enzymes and 

hydrogen peroxide (Svenson et al., 1989; Oberg et al., 1990; Wagner et al., 1990; Oberg and 

Rappe, 1992; Morimoto and Kenji, 1995) and with sewage sludge spiked with PCP (Oberg et al., 

1992). However, the extent to which CDDs/CDFs are formed in the environment via this 

mechanism cannot be estimated at this time. 

In 1991, Lahl et al. (1991) reported finding CDDs/CDFs in all 22 samples of the various 

types of composts analyzed. The hepta- and octa-substituted CDDs/CDFs were typically the 

dominant congener groups found. The I-TEQDF content of the composts ranged from 0.8 to 35.7 

ng I-TEQDF/kg. The CDDs/CDFs found in compost may primarily result from atmospheric 

deposition onto plants that are subsequently composted, but they may also be caused by uptake of 

CDDs/CDFs from air by the active compost (Krauss et al., 1994). CDDs/CDFs are also 

frequently detected in sewage sludges, and they may come primarily from the sources identified 

in Section 8.4.1. 

Peroxidases are common enzymes in nature. For example, the initial degradation of the 

lignin polymer by white- and brown-rot fungi is peroxidase catalyzed (Wagner et al., 1990). The 

conversion efficiency of chlorinated phenols to CDDs/CDFs that has been observed is low. In 

the solution studies, Oberg and Rappe (1992) and Morimoto and Kenji (1995) reported 

conversion efficiencies of PCP to OCDD of about 0.01% and 0.8%, respectively, and Wagner et 

al. (1990) reported a conversion efficiency of trichlorophenol to HpCDD of about 0.001%. 

Oberg et al. (1990) reported a conversion efficiency of trichlorophenols to CDDs/CDFs of about 

0.001%. In their sewage sludge study, Oberg et al. (1992) reported a conversion efficiency of 

PCP to total CDDs of 0.0002 to 0.0004%. 

9-1
 
Several researchers have conducted both laboratory and field studies in an attempt to 

better understand the extent of, and factors affecting, the fate or formation of CDDs/CDFs in 

composts and sewage sludges. The findings of several of these studies are discussed in the 

following paragraphs. These findings are not always consistent because the congener profiles 

and patterns detected—and the extent of CDD/CDF formation, if any—may vary with compost 

materials studied, differences in experimental or field composting design, and duration of the 

studies. 

Harrad et al. (1991) analyzed finished composts and active compost windrows from a 

municipally operated yard waste composting facility in Long Island, NY. Concentrations 

measured in 12 finished composts ranged from 14 to 41 ng I-TEQDF/kg (mean, 3 ng I-TEQDF/kg). 

The concentrations in the five active compost samples (1 to 30 days in age) ranged from 7.7 to 54 

ng I-TEQDF/kg (mean, 21 ng I-TEQDF/kg). The authors observed that CDD/CDF concentrations 

measured in two soil samples from the immediate vicinity of the composting facility were 

significantly lower (1 and 1.3 ng I-TEQDF/kg) than the levels found in the composts, suggesting 

that the source(s) of CDDs/CDFs in the composts was different from the source(s) affecting local 

soils. 

Harrad et al. (1991) also noted a strong similarity between the congener profiles observed 

in the composts and the congener profile of a PCP formulation (i.e., predominance of 1,2,4,6,8,9­

HxCDF and 1,2,3,4,6,8,9-HpCDF in their respective congener groups), which seemed to indicate 

that leaching of CDDs/CDFs from PCP-treated wood in the compost piles was the likely source 

of the observed CDDs/CDFs. The levels of PCP in the 12 finished composts ranged from 7 to 

190 μg/kg (mean, 33 μg/kg), and the PCP levels in the active compost samples ranged from 17 to 

210 μg/kg (mean, 68 μg/kg). The PCP level in both soil samples was 1.5 μg/kg. 

Goldfarb et al. (1992) and Malloy et al. (1993) reported the results of testing composts at 

three municipal yard waste composting facilities (5 to 91 ng I-TEQDF/kg; mean, 30 ng I

TEQDF/kg), two municipal solid waste composting facilities (19 to 96 ng I-TEQDF/kg; mean, 48 

ng I-TEQDF/kg), and one municipal facility composting solid waste and dewatered sewage sludge 

(37 to 87 ng I-TEQDF/kg; mean, 56 ng I-TEQDF/kg). All facilities were located in the United 

States. Two general trends were observed for the three types of composts: a progressive increase 

in analyte levels, with an increasing degree of chlorination for each compound type (CDDs, 

CDFs, chlorophenols, and chlorobenzenes), and a progressive increase in concentration of each 

congener or homologue group from yard waste to solid waste to solid waste/sewage sludge 

composts. As noted above, the mean TEQ concentrations showed this same trend, which was 

primarily due to increasing levels of 1,2,3,4,6,7,8-HpCDD and OCDD. The mean PCP 

concentrations in the three compost types were 20 μg/kg (yard waste), 215 μg/kg (solid waste), 

and 615 μg/kg (solid waste/sewage sludge). 

9-2
 

­
Comparison of congener profiles by the authors indicated that the CDD/CDF residue in 

PCP-treated wood in the compost feedstock was a major but not exclusive contributor of the 

observed CDDs/CDFs. The authors postulated that biological formation of HxCDDs, HpCDDs, 

and OCDD from chlorophenols (tri, tetra, and penta) in the compost could be responsible for the 

elevated levels of these congener groups relative to their presence in PCP. 

Oberg et al. (1993) measured the extent of CDD/CDF formation in three conventional 

garden composts; two were spiked with PCP and one was spiked with hexachlorobenzene. One 

PCP-spiked compost was monitored for 55 days and the other for 286 days. A significant 

increase in the concentrations of the more highly chlorinated congeners, particularly the 

HpCDDs, OCDD, and, to a lesser extent, OCDF, were observed. Similar results were reported 

for the hexachlorobenzene-spiked compost, which was monitored for 49 days. Oberg et al. stated 

that for a “typical” composting event, a two- to threefold increase in TEQ content corresponded 

with an elevation of 0.2 to 0.5 ng I-TEQDF/kg dry weight. 

Weber et al. (1995) subjected sewage sludges from two German communities to 

anaerobic digestion in laboratory reactors for 60 days. The two sludges were spiked with 2,3,5­

trichlorophenol (10 to 25 mg/kg), a mixture of 2,3,5-trichlorophenol and dichlorophenols (2.5 to 

25 mg/kg), or a mixture of di-, tri-, and tetrachlorobenzenes (4 to 40 mg/kg). The initial 

CDD/CDF concentrations in the two sludges were 9 and 20 ng I-TEQDF/kg. In nearly all of the 

digestion experiments, the addition of the precursors did not lead to any significant changes in 

concentrations. The only exceptions were increased 2,3,7,8-TCDF concentrations in the mixed 

chlorophenol experiments and decreased 2,3,7,8-TCDF concentrations in the mixed 

chlorobenzene experiments. However, the same increases or decreases for this congener were 

also observed in the controls (i.e., no precursors added). 

Researchers at the U.S. Department of Agriculture (USDA) (Fries et al., 1997) reported 

that dairy cows that were fed PCP-treated wood excreted amounts of OCDD almost four times 

greater than the amounts ingested. Feil and Tiernan (1997) reported that rats fed technical PCP 

had liver concentrations of HxCDD, HpCDD, HpCDF, OCDD, and OCDF two to three orders of 

magnitude higher than those of rats fed purified PCP. These results suggest the in vivo 

formation of CDDs/CDFs from pre-dioxins (i.e., chlorinated phenoxy phenols present as 

contaminants in the PCP). A follow-up USDA study (Huwe et al., 1998) investigated the 

metabolic conversion of a pre-dioxin (monochloro-2-phenoxyphenol) to OCDD in a feeding 

study with rats. The results of the study demonstrated the formation of OCDD from the predioxin, although the conversion was estimated to be less than 2%. Interestingly, the study noted 

that the presence of added OCDD in the feed material increased the percentage of pre-dioxin 

conversion. 

9-3
 
Wittsiepe et al. (1998) demonstrated that CDDs/CDFs can be formed through reaction of 

chlorophenols with myeloperoxidase (a component of neutrophile granulocytes, a subgroup of 

human leucocytes). The CDDs/CDFs formed showed different homologue patterns and 

formation rates depending on the degree of chlorination of the chlorophenol substrate. The 

formation rates ranged from 1 to 16 μmol of CDD/CDF per mol of chlorophenol substrate. 

9.1.2. Biotransformation of Higher CDDs/CDFs 

Results of several studies that examined the fate of a range of CDD/CDF congeners in 

pure cultures, sediments, and sludges indicate that under certain conditions some CDD/CDF 

congeners will undergo biodegradation to form lower-chlorinated (and possibly more toxic) 

CDDs/CDFs. However, the extent to which more toxic CDDs/CDFs are formed in the 

environment via this mechanism cannot be estimated at this time. 

Several reports indicate that CDDs and CDFs may undergo microbial dechlorination in 

anaerobic sediments. Adriaens and Grbic-Galic (1992, 1993) and Adriaens et al. (1995) reported 

the results of a series of microcosm studies using Hudson River sediment (contaminated with 

Aroclor 1242) and aquifer material (contaminated with CDDs) from Pensacola, FL. Both types 

of substrates were spiked with several CDDs (1,2,3,4,6,7,8-HpCDD; 1,2,3,4,7,8-HxCDD; and 

1,2,4,6,8,9-/1,2,4,6,7,9-HxCDD) and CDFs (1,2,3,4,6,7,8-HpCDF and 1,2,4,6,8-PeCDF) and 

monitored over a 16-month period at an incubation temperature of 30°C. The Hudson River 

sediment was spiked with 144 μg/kg of each congener, and the Pensacola aquifer material was 

spiked with 63 μg/kg of each congener. 

All of the congeners, with the exception of 1,2,3,4,6,7,8-HpCDF, showed a slow decrease 

in concentration over time, which was attributed to biologically mediated reductive 

dechlorination, with net disappearance rates ranging from 0.0031 wk-1 to 0.0175 wk-1 (i.e., halflives of approximately 1 to 4 yr). However, Adriaens et al. concluded that actual half-lives may 

be orders of magnitude higher. The experiment with 1,2,3,4,6,7,8-HpCDD yielded formation of 

1,2,3,4,7,8-HxCDD and 1,2,3,6,7,8-HxCDD. Thus, removal of the peri-substituted (1,4,6,9) 

chlorines was favored, with enrichment of 2,3,7,8-substituted congeners. No lower-chlorinated 

congeners were identified from incubation with the other tested congeners. 1,2,4,6,8-PeCDF was 

also examined in dichlorophenol-enriched cultures. After 6 months of incubation, several 

TCDFs were identified, which also indicated that peri-dechlorination was the preferred route of 

reduction. 

Barkovskii and Adriaens (1995, 1996) reported that 2,3,7,8-TCDD extracted from Passaic 

River sediments was susceptible to reductive dechlorination when incubated at 30°C under 

methanogenic conditions in a mixture of aliphatic and organic acids inoculated with 

microorganisms obtained from the sediments. The initial concentration of 2,3,7,8-TCDD (20 ± 4 

9-4
 

μg/L) decreased by 30% to 14 ± 2 μg/L over a period of 7 months, with the consecutive 

appearance and disappearance of tri-, di-, and mono-CDDs. Experiments were also conducted by 

spiking the sediment with HxCDDs, HpCDDs, and OCDD. Up to 10% of the spiked OCDD was 

converted to hepta-, hexa-, penta-, tetra-, tri-, di-, and mono-chlorinated isomers, but the reaction 

stoichiometry was not determined. Two distinct pathways of dechlorination were observed: the 

peri-dechlorination pathway of 2,3,7,8-substituted hepta- to penta-CDDs, resulting in the 

production of 2,3,7,8-TCDD, and the peri-lateral dechlorination pathway of non-2,3,7,8­

substituted congeners. 

Several studies have reported that CDDs/CDFs can be formed during composting 

operations through biological action on chlorophenols present in the compost feed material. The 

results of studies that specify likely involvement of chlorophenols are described in Section 9.1. 

Another possible formation mechanism was suggested by Vikelsoe et al. (1994), who reported 

that higher-chlorinated CDD/CDF congeners are formed when humic acid is reacted with a 

peroxidase enzyme, hydrogen peroxide, and sodium chloride. It is expected that some organic 

material in compost and sewage sludge has a humic-like structure. Several additional studies are 

described below in which the potential involvement of chlorophenols could not be assessed 

because chlorophenol concentrations in the composts were not reported. 

Schäfer et al. (1993) monitored the seasonal changes in the CDD/CDF content, as well as 

the extent of CDD/CDF formation, in composts from a vegetable and garden waste composting 

operation in Germany. Finished compost samples were collected and analyzed every 2 months 

for 1 yr. An annual cycle was observed in TEQ concentrations, with peak concentrations in the 

summer (approximately 8.5 ng I-TEQDF/kg) being 2.5 times higher than the lowest concentrations 

observed in the winter (approximately 3.5 ng I-TEQDF/kg). No seasonal source was apparent that 

could explain the observed differences in seasonal levels. 

The CDD/CDF content of the starting waste materials for two compost cycles (March and 

September) was measured to monitor the extent of CDD/CDF formation during composting. For 

the March cycle sample, most 2,3,7,8-substituted CDD/CDF congeners decreased in 

concentration during composting. Four CDF congeners showed a slight increase in concentration 

(less than 10%). For the September cycle sample, OCDD and HpCDD concentrations increased 

300% during composting. Increases of less than 10% were observed for HxCDDs and OCDF; all 

other 2,3,7,8-substituted CDD/CDF congeners showed decreases in concentrations during 

composting. 

Krauss et al. (1994) measured the extent of CDD/CDF formation during the composting 

of household waste using a laboratory compost reactor. After 11 wk, the TEQ content of the 

compost increased from 3 to 4.5 ng. The largest increases in mass content were observed for 

HpCDD (primarily 1,2,3,4,6,7,8-HpCDD) and OCDD. TCDD, PeCDD, and HxCDD showed no 

9-5
 

change in mass content. All CDF congener groups showed decreases in mass content; however, 

the concentrations in both the starting and the finished compost were close to the analytical 

detection limits. 

Oberg et al. (1994) reported the results of monitoring two household waste composts and 

two garden composts. The total CDD/CDF content of both household waste composts decreased 

over the 12-wk test period. Total CDD content and PCB content decreased, but total CDF 

content increased, in contrast to the findings of Krauss et al. (1994). However, a small increase 

in OCDD content in both composts was observed. The two garden composts were monitored for 

a 60-wk period. Total CDD/CDF concentration increased, with the largest increases observed for 

OCDD and HpCDDs. The lower-chlorinated CDFs decreased in concentration. 

As a follow-up to a preliminary study (Hengstmann et al., 1990) that indicated CDD/CDF 

concentrations may increase and congener profiles may change during anaerobic digestion of 

sewage sludge, Weber et al. (1995) subjected sewage sludges from two German communities to 

anaerobic digestion and aerobic digestion in laboratory reactors for 60 days and 20 days, 

respectively. The initial average I-TEQDF concentrations in the raw sludges were 20 and 200 ng 

I-TEQDF/kg. No significant increase or decrease in total CDD/CDF content or congener group 

content was observed with either sludge. In contrast, a significant decrease in CDD/CDF content 

was observed in the aerobic digestion experiments with both sludges. The greatest percentage 

decreases in congener group concentrations (greater than 40%) were observed for TCDF, 

PeCDF, HxCDF, TCDD, and PeCDD in the sludge initially containing 20 ng I-TEQDF/kg and for 

TCDF, TCDD, HpCDD, and OCDD in the initially high-content sludge. The greatest percentage 

decreases in congener concentrations (greater than 40%) were observed for non-2,3,7,8­

substituted congeners. 

The data presented in this section and in Section 9.1.1 do not provide a basis for making a 

release estimate via biotransformation; therefore, biotransformation releases are classified as 

Category E (not quantifiable). 

9.1.3. Dioxin-Like Compounds in Animal Manure 

In 2000, approximately 9 billion individual livestock and poultry animals were raised on 

commercial farms in the United States (U.S. Census Bureau, 2001). It is estimated that beef 

animals, dairy cows, chickens, turkeys, and pigs, combined, produced in excess of 190 billion kg 

(dry weight) of manure in 2000 (Table 9-1). Because livestock and poultry manure can provide 

valuable organic material and nutrients for crop and pasture growth, most of the animal manure 

generated at commercial farms and animal feed lots is applied to farmland as fertilizer. To the 

extent dioxin-like compounds may contaminate animal manures, the practice of land-spreading 

animal waste may result in releases of CDDs/CDFs to the open and circulating environment. 

9-6
 

Table 9-1. Estimated quantity of animal manure produced in the United 

States in 2000 

Species 

Numbers of 

individuals 

on farms in 

2000a 

Average 

weight of 

animal 

(lbs)

b 

Total live 

weight on 

farms (lbs) 

Manure 

generation rate 

factor 

(dry weight lb/lb 

live unit 

weight/day)c 

Manure 

generated 

(lb/yr dry 

weight) 

Manure 

produced 

(kg/yr dry 

weight) 

Swine 6.73e+07 135 9.09e+09 8.2e!03 2.72e+10 1.23e+10 

Layer 4.35e+08 4 1.74e+09 1.6e!02 1.02e+10 4.61e+09 

Broiler 8.26e+09 2 1.65e+10 2.1e!02 1.27e+11 5.74e+10 

Turkey 2.7e+08 15 4.05e+09 1.2e!02 1.77e+10 8.04e+09 

Beef 9.73e+07 800 7.78e+10 6.9e!03 1.96e+11 8.89e+10 

Dairy cow 9.21e+06 1,400 1.29e+10 1e!02 4.71e+10 2.13e+10 

Total 4.25e+11 1.93e+11 a

Source: U.S. Census Bureau (2001). b

Source: U.S. EPA (2001d). c

Source: Stevens and Jones (2003). 

Stevens and Jones (2003) published results of CDD and CDF detection in animal manure applied 

to farmland in the United Kingdom. Manure from six milking dairy cows was sampled at six 

farms in the northern United Kingdom. In addition, single samples of sheep, chicken, and pig 

manure were collected from other farms in the region. The samples were shipped to a laboratory 

for trace chemical analysis. Samples were analyzed using high-resolution gas chromatography 

coupled with high-resolution mass spectrometry and a capillary column for the identification of 

CDD/CDF congeners. Recoveries of the internal standard ranged from 51 to 94%, with a mean 

of 74% for CDD/CDF congeners. Table 9-2 summarizes the results of the study. The pig and 

chicken manure contained approximately 0.2 ng WHO-TEQ/kg, and the cow manure averaged 

3.6 ng WHO-TEQ/kg in concentration. 

This study provides extremely limited data on the possible levels and occurrences of 

dioxin-like compounds in farm animal manure, and, therefore, these data are clearly not 

representative of national releases of dioxin-like compounds from the land application of all farm 

animal manure in the United States. Accordingly, EPA currently considers this source to be 

unquantifiable (Category E) in terms of dioxin emissions. 

9-7
 

Table 9-2. CDD and CDF concentrations (ng/kg dry weight) in samples of 

animal manure in the United Kingdom 

Congener 

Cows (n = 6) 

(mean) Sheep (n = 1) Pig (n = 1) 

Chicken 

(n = 1) 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

0.17 

0.46 

2.4 

4.5 

2.6 

120 

460 

0.11 

0.41 

0.9 

0.86 

0.56 

9.4 

53 

0.01 

0.07 

0.26 

0.1 

0.07 

0.8 

11 

0.01 

0.04 

0.03 

0.09 

0.12 

1.4 

14 

Total 2,3,7,8-CDD 590.1 65.2 12.3 15.7 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.3 

0.3 

0.28 

0.6 

0.51 

1.9 

0.4 

7.6 

12 

35 

1.2 

1.1 

1.2 

1.4 

1.1 

0.15 

1.4 

5.2 

0.56 

5 

0.03 

0.04 

0.06 

0.05 

0.06 

0.04 

0.06 

0.48 

0.04 

0.73 

0.03 

0.09 

0.12 

0.15 

0.07 

0.05 

0.14 

0.37 

0.09 

0.8 

Total 2,3,7,8-CDF 58.9 18.3 1.6 1.9 

Total CDD/CDF 649 83.5 13.9 17.6 

WHO-TEQ 3.6 2.1 0.2 0.2 

Source: Stevens and Jones (2003). 

9.2. PHOTOTRANSFORMATION OF CHLOROPHENOLS 

Several researchers have demonstrated that CDDs/CDFs can be formed via photolysis of 

PCP under laboratory conditions. However, the extent to which CDDs/CDFs are formed in the 

environment via this mechanism cannot be estimated at this time. 

Lamparski et al. (1980) conducted laboratory studies to determine the effect of simulated 

summer sunlight on the formation of OCDD, HpCDDs, and HxCDDs in wood pressure-treated 

in the laboratory with PCP. In the first set of experiments, wood veneers (southern pine) treated 

with purified PCP or Dowicide EC-7, using methylene chloride as the PCP carrier, were exposed 

to light for 70 days. The PCP concentration in the treated wood was 5% by weight, which 

approximates the concentration in the outer layer of PCP-treated wood utility poles. Photolytic 

condensation of PCP to form OCDD was observed, with the OCDD concentration increasing by 

a maximum factor of 3,000 for the purified PCP and by a factor of 20 for EC-7 at about day 20 

9-8
 

before leveling off. HpCDD and HxCDD were also formed, apparently by photolytic 

degradation of OCDD rather than by condensation of PCP and tetrachlorophenols. The HxCDD 

concentration increased by a factor of 760 for the purified PCP and by a factor of 50 for EC-7 

over the 70-day exposure period. The predominant HpCDD congener formed was 1,2,3,4,6,7,8­

HpCDD as a result of an apparently preferential loss of chlorine at the peri position (positions 1, 

4, 6, and 9). 

In a second set of experiments conducted by Lamparski et al. (1980), a hydrocarbon oil 

(P-9 oil) was used as the carrier to treat the wood. The increases observed in the OCDD, 

HpCDD, and HxCDD were reported to be much lower relative to the increases observed in the 

first set of experiments, which used methylene chloride as the carrier. Results were reported only 

for OCDD. The OCDD concentration increased by a maximum factor of 1.5 for both EC-7 and 

technical PCP and by a factor of 88 for purified PCP. The authors concluded that the oil either 

reduced condensation of PCP to OCDD or accelerated degradation to other species by providing 

a hydrocarbon trap for free-radical species. 

Vollmuth et al. (1994) studied the effect of irradiating laboratory water and landfill 

seepage water that contained PCP under conditions simulating those used to purify water with 

ultraviolet (UV) radiation (5-hr exposure to 254 nm radiation from low-pressure mercury lamps). 

Before irradiation, the three solutions tested contained approximately 1 mg/L of PCP or PCP-Na, 

but the CDD/CDF content of one solution varied dramatically from those of the other two (1.5 

vs. 2,066 and 2,071 pg I-TEQDF/L). Irradiation resulted in nearly total destruction of PCP 

(greater than 99% loss) in all three experiments. An overall net increase in I-TEQDF-content was 

observed in the initially low I-TEQDF-content water, but a net decrease was observed for the two 

initially high I-TEQDF-content waters. 

Irradiation of laboratory water containing purified PCP showed an increase in I-TEQDF 

concentration from 1.5 pg/L to 214.5 pg/L. The increase was due entirely to the formation of 

1,2,3,4,6,7,8-HpCDD, OCDD, and 1,2,3,4,6,7,8-HpCDF. Formation of non-2,3,7,8-substituted 

HpCDDs and HpCDFs was also observed. The ratios of the concentrations of these non-2,3,7,8­

congeners to the concentrations of the 2,3,7,8-congeners were 0.6 for HpCDDs and 5 for 

HpCDFs. The HpCDD and HpCDF congeners formed indicate that the operative mechanism 

was photoinduced dechlorination of OCDD at a peri position and dechlorination of OCDF at 

only the 1 and 9 peri positions. 

Irradiation of water containing technical PCP-Na (Dowicide-G) resulted in a net loss in ITEQDF content, from 2,065.5 pg/L to 112.7 pg/L. The only 2,3,7,8-substituted congener showing 

an increased concentration was 1,2,3,6,7,8-HxCDD. The other congeners originally present in 

the technical PCP-Na showed reductions of 80.6 to 100%. 

9-9
 
The I-TEQDF content of seepage water from a landfill (2,071 pg I-TEQDF/L) was reduced 

by a factor of 2, to 1,088 pg I-TEQDF/L. However, several 2,3,7,8-substituted congeners did 

increase in concentration (1,2,3,6,7,8-HxCDD; 1,2,3,4,6,7,8-HpCDD; 1,2,3,4,6,7,8-HpCDF; and 

OCDF). 

Waddell et al. (1995) also studied the effect of irradiating distilled laboratory water 

containing PCP under conditions simulating those used to purify water with UV radiation. The 

results obtained were similar to those of Vollmuth et al. (1994). Analytical-grade PCP at a 

concentration of 10 mg/L was exposed for 12 min to 200 to 300 nm radiation from a mediumpressure mercury lamp. All CDD/CDF congener groups increased in concentration over the 12­

min exposure period, with the greatest increases observed for OCDD (75-fold increase) and 

HpCDDs (34-fold increase). The I-TEQ ­ DF content of the solution increased from 4.2 pg I

TEQDF/L to 137 pg I-TEQDF/L over the 12-min period. The dominant congeners formed, in terms 

of both concentration and contribution to I-TEQDF, were 1,2,3,4,6,7,8-HpCDD, OCDD, and 

1,2,3,7,8,9-HxCDD. 

9.2.1. Photolysis of Higher CDDs/CDFs 

Photolysis appears to be one of the few environmentally significant degradation 

mechanisms for CDDs/CDFs in water, air, and soil. Although good mass balances were not 

obtained and the photolytic pathways for CDDs/CDFs were not fully identified in most studies, a 

major photolysis pathway appears to be photodechlorination, resulting in formation of lowerchlorinated CDDs/CDFs. A preferential loss of chlorines from the peri positions (1, 4, 6, and 9) 

rather than from the lateral positions (2, 3, 7, and 8) was reported for some congener groups 

when irradiated as dry films and sorbed to soil and in gas-phase CDDs/CDFs (Choudhry and 

Webster, 1989; Kieatiwong et al., 1990; Sivils et al., 1994, 1995; Tysklind et al., 1992). Several 

researchers reported that carbon-oxygen cleavage and other mechanisms may be similarly or 

more important pathways for CDDs/CDFs containing four or fewer chlorines. 

Because of the difficulties inherent in controlling experimental variables for nonvolatile 

and highly lipophilic compounds such as CDDs/CDFs, few photolysis studies have been 

performed on natural waters, soils, atmospheric particulates, and atmospheric gases to examine 

the rates and products of photolysis under environmentally relevant conditions. Thus, it is not 

possible at this time to quantitatively estimate the mass of various CDD/CDF congeners formed 

in the environment annually via photolytic mechanisms. The following sections summarize the 

key findings of environmentally significant studies for water, soil, vegetation, and air. 

9-10
 

9.2.2. Photolysis in Water 

Numerous studies have demonstrated that CDDs/CDFs will undergo photodechlorination 

following first-order kinetics in organic solution, with preferential loss of chlorine from the 

lateral positions. Photolysis is slow in pure water, but it increases dramatically when solvents 

serving as hydrogen donors such as hexane, benzene, methanol, acetonitrile, hexadecane, ethyl 

oleate, dioxane, and isooctane are present. However, only a few studies have examined the 

photolysis of CDDs/CDFs using natural waters and sunlight. 

Choudhry and Webster (1989) experimentally determined the sunlight photolysis half-life 

of 1,3,6,8-TCDD in pond water to be 3.5 days (more than 10 times greater than the half-life 

predicted by laboratory experiments using a water/acetonitrile solution). The authors attributed 

this significant difference in photolysis rates to the light screening/quenching effects of dissolved 

organic matter. 

Friesen et al. (1990) examined the photolytic behavior of 1,2,3,4,7-PeCDD and 

1,2,3,4,6,7,8-HpCDD in water:acetonitrile (2:3, v/v) and in pond water under sunlight at 50 

degrees North latitude. The observed half-lives of these two compounds in the water:acetonitrile 

solution were 12 and 37 days, respectively, but were much shorter in pond water, 0.94 and 2.5 

days, respectively. Similarly, Friesen et al. (1993) studied the photodegradation of 2,3,7,8-TCDF 

and 2,3,4,7,8-PeCDF by sunlight using water:acetonitrile (2:3, v/v) and lake water. The observed 

half-lives were 6.5 and 46 days, respectively, in the water:acetonitrile solution and 1.2 and 0.19 

days, respectively, in lake water. The significant differences between the natural water and the 

water:acetonitrile solution results were attributed to indirect or sensitized photolysis due to the 

presence of naturally occurring components in the lake and pond water. 

Dung and O’Keefe (1992), in an investigation of aqueous photolysis of 2,3,7,8-TCDF and 

1,2,7,8-TCDF, reported findings similar to those of Friesen et al. (1993). The photolysis rates of 

the two TCDF congeners observed in river and lake water (half-lives of about 4 to 6 hr) were 

double those observed in pure water (half-lives of about 8 to 11 hr). The authors attributed the 

difference in rates to the presence of natural organics in the river and lake water that may act as 

sensitizers. 

9.2.3. Photolysis on Soil 

Photolysis of CDDs/CDFs on soil has not been well characterized. According to the data 

generated to date, however, photolysis is an operative degradation process only in the nearsurface soil where UV light penetrates (the top few millimeters or less of soil), and 

dechlorination of peri-substituted chlorines appears to occur preferentially. 

Miller et al. (1989) studied the CDD degradation products resulting from irradiation of 

13C-labeled OCDD on two soil types using sunlamps. Approximately 38 to 42% of the OCDD 

9-11
 

was degraded by day 5 of the experiment; no significant further loss of OCDD was observed over 

the following 10 days. Although the authors determined that photodechlorination was not the 

dominant photolysis pathway, it was observed in both soils; approximately 10 to 30% of the 

lower-chlorinated congeners were produced from the immediate higher-chlorinated congeners. 

The HpCDD and HxCDD congeners observed as degradation products were present in 

proportions similar to the number of congeners in each congener group. However, the 

investigators observed greater yields of 2,3,7,8-TCDD and 1,2,3,7,8-PeCDD than would be 

expected on the basis of the number of potential TCDD and PeCDD congeners. One-fifth to 

one-third of the total yield of PeCDDs was 1,2,3,7,8-PeCDD, and one-half of the total yield of 

TCDDs was 2,3,7,8-TCDD. 

Kieatiwong et al. (1990) performed experiments similar to those of Miller et al. (1989) 

using natural sunlight rather than sunlamps for irradiation of 13C-labeled OCDD on soils. 

Photodechlorination was estimated to account for approximately 10% of the loss of OCDD. 

One-third to one-half of the total yield of PeCDDs was 1,2,3,7,8-PeCDD, and one-half of the 

total yield of TCDDs was 2,3,7,8-TCDD. These findings, along with those of Miller et al., 

indicate that the 2,3,7,8-substituted TCDD and PeCDD congeners were either preferentially 

formed or were photochemically less reactive than the other congeners that were formed. 

Tysklind et al. (1992) studied the sunlight photolysis of OCDD on soil and reported 

results similar to those of Miller et al. (1989) and Kieatiwong et al. (1990). Photodechlorination 

was observed with production of HpCDDs, HxCDDs, PeCDDs, and TCDDs over the 16-day 

irradiation period. Photodechlorination at the peri-substituted positions was the preferred 

photodechlorination mechanism; the proportions of 2,3,7,8-substituted congeners present in the 

soils after 16 days for each congener group were as follows: HxCDD, 65%; PeCDD, 40%; and 

TCDD, 75%. Tysklind et al. (1992) also studied the sunlight photolysis of OCDF on soil. 

Photodechlorination was observed; however, unlike the case with OCDD, photodechlorination of 

the lateral-substituted positions was found to be the dominant photodechlorination mechanism, 

resulting in a relative decreasing proportion of 2,3,7,8-substituted congeners during the 

irradiation period. 2,3,7,8-TCDF was not observed in any of the irradiated samples. 

9.2.4. Photolysis on Vegetation 

Photolysis of CDDs/CDFs sorbed on the surface of vegetation has not been well 

characterized, and the findings to date are somewhat contradictory. McCrady and Maggard 

(1993) reported that 2,3,7,8-TCDD sorbed on the surface of reed canary grass (Phalaris 

arundinacea L.) underwent photolytic degradation, with a half-life of 44 hr in natural sunlight. 

In contrast, Welsch-Pausch et al. (1995) found little difference in the CDD/CDF congener 

9-12
 

patterns between grass (Lolium multiflorum) grown on an outdoor plot and grass grown in a 

greenhouse (i.e., UV light transmission blocked). 

In an attempt to clarify this contradiction, Welsch-Pausch and McLachlan (1995) studied 

the photodegradation of CDDs/CDFs on pasture grass (Arrhenatherion elatioris) during two 

growing cycles (summer and autumn) using two greenhouses. One greenhouse was constructed 

of glass that blocks UV transmission and the other was constructed of plexiglass (4 mm) with a 

UV light transmission of greater than 50% in the 280 to 320 nm range. In both the summer and 

the autumn exposure periods, the concentrations of CDDs/CDFs (on a congener-group basis) 

were similar in the grass exposed to UV light and the grass that was not exposed. The authors 

concluded that if photodegradation was occurring, it was a relatively insignificant factor in the 

accumulation of CDDs/CDFs in pasture grass. 

9.2.5. Photolysis in Air 

Photolysis of CDDs/CDFs in the atmosphere has not been well characterized. On the 

basis of data generated to date, however, photolysis appears to be a significant mechanism for 

degradation (principally, dechlorination of the peri-substituted chlorines) of those CDDs/CDFs 

present in the atmosphere in the gas phase. For airborne CDDs/CDFs sorbed to particulates, 

photolysis appears to proceed very slowly, if at all. Because of the low volatility of CDDs/CDFs, 

few studies have been attempted to measure actual rates of photodegradation of gas-phase 

CDD/CDF, and only recently have studies examined the relative importance of photolysis to 

particulate-bound CDDs/CDFs. 

Sivils et al. (1994, 1995) studied the gas-phase photolysis of several CDDs (2,3,7­

TrCDD; 2,3,7,8-TCDD; 1,2,3,4-TCDD; 1,2,3,7,8-PeCDD, and 1,2,4,7,8-PeCDD) by irradiating 

the effluent from a gas chromatograph with broadband radiation in the UV/visible region for 

periods of up to 20 min. The irradiated sample was then introduced into a second gas 

chromatograph to measure the extent of dechlorination. The results showed that degradation 

followed first-order kinetics and that an inverse relationship existed between the degree of 

chlorination and the rate of disappearance. Although the lack of photoproducts prevented an 

independent confirmation of the preferential loss mechanism, the results indicate that laterally 

substituted congeners (chlorines at the 2, 3, 7, and 8 positions) degrade at a slower rate than do 

the peri-substituted congeners (chlorines at the 1, 4, 6, and 9 positions). Although Sivils et al. 

(1994) did not present the rate constants, the degradation rate for 2,3,7,8-TCDD (30% loss in 

20 min) was reported to be slower than the rates for all other tested CDDs. Also, 1,2,4,7,8­

PeCDD (with two perichlorines) degraded significantly faster than did 1,2,3,7,8-PeCDD (with 

only one perichlorine). 

9-13
 
Mill et al. (1987) studied the photolysis of 2,3,7,8-TCDD sorbed onto small-diameter fly 

ash particulates suspended in air. The results indicated that fly ash confers photostability on 

2,3,7,8-TCDD. Little (8%) to no loss was observed on the two fly ash samples after 40 hr of 

illumination. Tysklind and Rappe (1991) and Koester and Hites (1992) reported similar results 

of photolysis studies with fly ash. Tysklind and Rappe subjected fly ash from two German 

incinerators to various simulated environmental conditions. The fraction of photolytically 

degradable CDD/CDF after 288 hr of exposure was in the range of 20 to 40% of the extractable 

CDD/CDF. However, a 10 to 20% reduction was also observed in the darkened control samples. 

With the exception of HpCDD and HpCDF, the concentration of all other congener groups either 

increased or stayed the same during the exposure period from hour 144 to hour 288. 

Koester and Hites (1992) studied the photodegradation of CDDs/CDFs naturally adsorbed 

to fly ash collected from five electrostatic precipitators. They observed no significant 

degradation in 11 photodegradation experiments performed on the ash for periods ranging from 2 

to 6 days. The authors concluded that (a) the absence of photodegradation was not due to the 

absence of a hydrogen-donor organic substance; (b) other molecules on the ash, as determined by 

a photolysis experiment with an ash extract, inhibited photodegradation, either by absorbing light 

and dissipating energy or by quenching the excited states of the CDDs/CDFs; and (c) the surface 

of the ash itself may have hindered photolysis by shielding the CDDs/CDFs from light. 

9.3. CDDs/CDFs IN BALL CLAY 

9.3.1. Initial Discovery of CDD/CDF Contamination of Ball Clay 

The presence of dioxin-like compounds in ball clay was discovered in 1996 as a result of 

an investigation to determine the sources of relatively high levels of dioxin found in two chicken 

fat samples during a national survey of poultry. The survey was conducted jointly by USDA, the 

U.S. Food and Drug Administration (FDA), and EPA to assess the national prevalence and 

concentrations of CDDs, CDFs, and coplanar PCBs in poultry (Ferrario et al., 1997). 

The results of the investigation indicated that soybean meal added to chicken feed was the 

source of dioxin contamination (Ferrario et al., 2000). Further investigation showed that the 

CDD contamination came from the ball clay added to the soymeal as an anticaking agent. The 

ball clay was added at approximately 0.3 to 0.5% of the soybean meal. Samples of raw ball clay 

were subsequently taken at the mine of origin in Mississippi. Analysis of the samples showed 

elevated levels of CDDs with a congener profile similar to the CDD profiles found in the 

soymeal, chicken feed, and immature chickens. 

9-14
 

9.3.2. Characteristics of Mississippi Embayment Ball Clays 

The ball clays from the mine discussed above are part of a larger ball clay resource that 

spans portions of western Kentucky, Tennessee, and Mississippi. These clays were deposited 

along the shores of the Mississippi embayment during the early to middle Eocene epoch, which 

occurred approximately 40 to 45 million years ago. The Mississippi embayment ball clays are 

secondary clays composed mainly of poorly defined crystalline kaolinite. Other minerals present 

include illite, smectite, and chlorite. Quartz sand is the major nonclay mineral. These deposits 

of ball clay occur in lenses surrounded by layers of sand, silt, and lignite. The clays can have a 

gray appearance caused by the presence of finely divided carbonaceous particles. It is not 

uncommon to find black carbonized imprints of fossil leaves and other plant debris in the clay 

(Patterson and Murray, 1984). 

The plasticity of ball clay makes it an important natural resource for the ceramic industry. 

The breakdown of the ceramic uses of ball clay is 33% for floor and wall tile, 24% for sanitary 

ware, 11% for pottery, and 32% for other industrial and commercial uses (Virta, 2000). A minor 

use of ball clay was as an anticaking agent in animal feeds, but this use has been banned by the 

FDA (Headrick et al., 1999). Total mining of ball clay in 1999 was 1.14 million metric tons 

(Virta, 2000). 

9.3.3. Levels of Dioxin-Like Compounds in Ball Clay 

The joint EPA/FDA and USDA investigation of ball clay as a source of dioxin 

contamination in animal feeds resulted in sampling the clay at an operational mine in western 

Mississippi. Eight samples of raw (unprocessed) ball clay were collected from an open mining 

pit at a depth of about 10 to 15 m. Samples were prepared and analyzed by EPA using EPA 

Method 1613 (Ferrario et al., 2000). The concentrations of the CDDs/CDFs present in the raw 

ball clay samples from the one mine are shown in Table 9-3. The ratio of the limits of detection 

to the limits of quantification for the CDDs/CDFs in the clay samples were 0.5:1 pg/g (ppt, dry 

weight) for the tetras; 1:2 pg/g for the pentas, hexas, and heptas; and 5:10 pg/g for the octas. The 

mean concentrations of all of the CDDs exceeded 100 ppt (dry weight). 

OCDD was found at the highest concentration in all of the samples, followed by either 

1,2,3,4,6,7,8-HpCDD or 1,2,3,7,8,9-HxCDD. The maximum OCDD concentration in the eight 

samples was approximately 59,000 pg/g. The most toxic tetra and penta congeners were present 

at unusually high concentrations in all of the samples, with average concentrations of 711 pg/g 

and 508 pg/g for 2,3,7,8-TCDD and 1,2,3,7,8-PeCDD, respectively. Although the ball clays 

showed elevated levels of 2,3,7,8-substituted CDDs, they showed very low levels of 2,3,7,8­

substituted CDFs. In addition, there was a consistent ratio within the HxCDD congener 

9-15
 

Table 9-3. Concentrations of CDDs (pg/g, dry weight) in eight ball clay 

samples in the United States 

Congener Mean Median Minimum Maximum TEQDF-WHO98 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

711 

508 

131 

456 

2,093 

2,383 

20,640 

617 

492 

134 

421 

1,880 

2,073 

4,099 

253 

254 

62 

254 

1,252 

1,493 

8,076 

1,259 

924 

193 

752 

3,683 

3,346 

58,766 

711 

508 

13 

46 

209 

24 

2 

Total TEQ 1,513 

Source: Ferrario et al. (2000). 

distribution across all samples (1,2,3,7,8,9-HxCDD was present at higher concentrations than the 

other 2,3,7,8-substituted HxCDD congeners). The average percent distribution of the three 

individual 2,3,7,8-hexa congeners was 5, 17, and 78%, respectively. This congener pattern was 

observed in all the raw ball clay samples analyzed. 

The mean total TEQDF-WHO98 for the raw ball clay was determined to be 1,513 pg/g dry 

weight; 2,3,7,8-TCDD accounted for 47% of the TEQDF-WHO98, followed by 1,2,3,7,8-PeCDD 

at 34%. As expected, even though present at the highest concentration, OCDD contributed less 

than 1% percent of the total TEQDF-WHO98 due to its relatively small WHO-TEF. In 

comparison, the typical ranges of background TEQDF-WHO98 concentrations in North American 

urban and rural surface soil samples were found to be 2 to 21 pg/g and 0.1 to 6 pg/g, respectively 

(U.S. EPA, 2000c). In soil samples, all 2,3,7,8-CDD/CDF congeners were detected, and 2,3,7,8­

TCDD represented less than 1% of total CDD/CDF present. The most prevalent congeners in 

soils were OCDD, followed by OCDF. Table 9-4 compares the mean CDD/CDF congener group 

concentrations in ball clay with those in rural and urban background soils. This comparison 

indicates there are few similarities between the ball clay and soils in the congener group 

distributions. 

9.3.4. Evidence for Ball Clay as a Natural Source 

Several lines of evidence suggest that dioxin-like compounds in ball clay are of natural 

origin. The clay samples were obtained from undisturbed deposits. It is unknown how human 

activity could have contaminated these deposits without disturbing them. The EPA laboratory in 

Athens, GA, analyzed the Mississippi mine clays using a broad screen for anthropogenic 

contaminants and no compounds were found outside of the normal range (Ferrario et al., 2000). 

9-16
 

Table 9-4. Comparison of the mean CDD/CDF congener group 

concentrations in ball clay with those in urban and rural soils in North 

America (pg/g, dry weight) 

Congener group 

Mean concentration 

Raw ball clay Urban background soil Rural background soil 

TCDD 

TCDF 

3,729 

6 

36.1 

23.5 

2.3 

6.8 

PeCDD 

PeCDF 

4,798 

2 

18.1 

40.8 

4.1 

12.7 

HxCDD 

HxCDF 

6,609 

6 

31.7 

23.5 

22.7 

21.9 

HpCDD 

HpCDF 

6,194 

9 

194.4 

46.4 

114.7 

37.3 

OCDD 

OCDF 

11,222 

11 

2,596 

40.2 

565.1 

33.5 

Total CDD/CDF 32,586 3,050.7 821.1 

Sources: Adapted from U.S. EPA (2000c); Ferrario et al. (2000) 

All known anthropogenic sources of dioxin have associated with them a wide variety of other 

contaminants. The absence of elevated levels of other compounds is strong evidence that the 

dioxins found in the clay were not the result of waste disposal. 

The congener profiles of ball clay do not match those of known anthropogenic sources. 

Cleverly et al. (1997) reported on the congener profiles that are typical of known anthropogenic 

sources of dioxin-like compounds in the United States. The results of this study, presented 

below, were used as a basis of comparison for the profiles of raw ball clay. 

The congener pattern characteristic of waste combustion sources differs significantly 

from the ball clay profile in several aspects. In combustion source emissions, all 2,3,7,8­

substituted CDD and CDF congeners are measured, and 2,3,7,8-TCDD is usually 0.1 to 1% of 

total CDD/CDF mass emitted. In ball clay, 2,3,7,8-TCDD is approximately 5% of total mass of 

dioxins present. As with the ball clay, the most prevalent 2,3,7,8-Cl-substituted CDD congeners 

in most incinerator emissions are OCDD and 1,2,3,4,6,7,8-HpCDD; however, combustion 

emissions contain appreciable amounts of CDFs, of which the 1,2,3,4,6,7,8-HpCDF, OCDF, 

1,2,3,4,7,8-HxCDF, 2,3,7,8-TCDF, and 2,3,4,6,7,8-HxCDF congeners dominate. 

The combustion of wood generates a congener profile not unlike that of waste 

combustion (i.e., the ratio of CDD:CDF is <1), and all laterally substituted congeners can be 

detected in emissions. The combustion of tree bark produces a congener profile in which the 

9-17
 

CDD:CDF ratio is >1, showing only minimal and barely detectable levels of CDFs in the smoke, 

the exception being that 2,3,7,8-TCDF is present at approximately 2% of total mass. The 

dominant congener in tree bark combustion emissions is OCDD (>30% total CDD/CDF mass), 

followed by 1,2,3,4,6,7,8-HpCDD and 1,2,3,7,8,9-HxCDD. 

The congener profile of 2,4-D salts and esters seems to mimic a combustion source 

profile in the number of congeners represented and in the minimal amount of 2,3,7,8-TCDD 

relative to all 2,3,7,8-Cl-substituted congeners. Nevertheless, unlike the combustion source 

profile, the 1,2,3,7,8-PeCDD and the 1,2,3,4,6,7,8-HpCDF constitute major fractions of total 

CDD/CDF contamination present in 2,4-D. The congener profile of technical-grade PCP is 

clearly dominated by OCDD and 1,2,3,4,6,7,8-HpCDD; however, only trace amounts of 2,3,7,8­

TCDD are detected in PCP, and 1,2,3,4,6,7,8-HpCDF and OCDF constitute roughly 15% of 

typical formulations. 

Metal smelting and refining processes, such as secondary aluminum, copper, and lead 

smelting, also have all the 2,3,7,8-Cl-substituted CDD/CDF congeners in stack emissions. In 

secondary aluminum smelting, 2,3,7,8-TCDD is less than 0.1% of total CDDs/CDFs, whereas 

PeCDF is nearly 25% of total emissions of dioxin-like compounds, and the CDD/CDF ratio is 

<1. Secondary copper operations show a similar pattern of CDD/CDF emissions, but with six 

compounds dominating emissions: 1,2,3,4,7,8-HxCDF; 1,2,3,6,7,8-HxCDF; 1,2,3,4,6,7,8­

HpCDF; OCDF; OCDD; and 1,2,3,4,6,7,8-HpCDD. In iron ore sintering, the dominant congener 

in emissions of 2,3,7,8-Cl-substituted compounds is 2,3,7,8-TCDF. 

A number of studies have shown that natural processes can produce chlorinated aromatic 

compounds, including dioxin-like compounds. Gribble (1994) reviewed the biological 

production of a wide variety of halogenated organic compounds in nature. The Mississippi salt 

marsh grass “needlerush” (Juncus roemerianus) contains the aromatic compound 1,2,3,4­

tetrachlorobenzene, and the blue-green alga Anacystis marina naturally contains chlorophenol. 

The soil fungus Penicillium sp. produces 2,4-dichlorophenol, and the common grasshopper is 

known to secrete 2,5-dichlorophenol. 

Urhahn and Ballschmiter (1998) also provide a good review of the chemistry of the 

biosynthesis of chlorinated organic compounds under natural conditions. It has been 

hypothesized that CDDs, CDFs, and other chlorinated aromatic compounds can be naturally 

formed from halogenated humic substances, and halomethanes can be formed through 

chloroperoxidase-mediated reactions in undisturbed peat bogs (Silk et al., 1997). A similar 

chloroperoxidase-mediated biochemical formation of CDDs/CDFs from chlorophenols was 

achieved under laboratory conditions by Oberg and Rappe (1992). 

It has been observed that chlorophenols can be biosynthesized (Gribble, 1994; Silk et al., 

1997), and that chorophenols are readily adsorbed into peat-bentonite mixtures (Viraraghavan 

9-18
 

and Slough, 1999). Hoekstra et al. (1999) offers the hypothesis that 2,3,7,8-TCDD, 1,2,3,7,8­

PeCDD, and 1,2,3,7,8,9-HxCDD can be naturally formed in soils of coniferous forests from 

chlorinated phenol. These same congeners are also the predominant congeners in the ball clay 

from the Mississippi embayment. Although none of these natural processes can be directly 

connected with the presence of dioxin in ball clay, the existence of such mechanisms lends 

plausibility to a hypothesis that they are of natural origin. 

CDDs/CDFs have been found in other clays quite distant from Mississippi embayment 

ball clay deposits. No evidence of anthropogenic sources have been discovered in these areas 

either. The presence of CDDs has been discovered in kaolinitic clay mined in Germany (Jobst 

and Aldag, 2000). Because no anthropogenic source could be determined to explain the presence 

and levels of CDDs in the ball clay, the authors speculated that they were the result of an 

unknown geologic process. In addition, the German clay also has a congener profile similar to 

that observed in the Mississippi ball clay, with an absence of CDFs at comparable concentrations 

and the predominance of 1,2,3,7,8,9-HxCDD among the toxic hexa-CDDs. The similarity in the 

congener profiles in ball clay mined in the United States and Germany suggests a common origin 

to the CDDs present in these clays (Ferrario et al., 2000). 

In summary, no anthropogenic sources have been identified that explain the levels and 

profiles of CDDs/CDFs present in ball clay. On the other hand, no definitive scientific evidence 

has been brought forward that identifies the principal chemical and physical mechanism involved 

in the selective chemical synthesis of CDDs under the conditions inherent in the formation of 

ball clays some 40 million years ago. 

9.3.5. Environmental Releases of Dioxin-Like Compounds from the Mining and Processing 

of Ball Clay 

In 1995, approximately 993 million kg of ball clay was mined in the United States (Virta, 

2000). Multiplication of the mean TEQDF-WHO98 concentration in mined ball clay by the total 

amount of ball clay mined in 1995 gives an estimate of 1,502 g TEQDF-WHO98 contained in all 

the ball clay mined in 1995. It is unknown whether any of these CDDs are released to the 

environment during the mining, initial refining, and product handling. As discussed above, most 

ball clay is used to produce ceramics through a process of high-temperature vitrification. The 

temperatures found in ceramic kilns are well above the levels needed for both volatilization and 

destruction of CDDs. Despite these high temperatures, it is unclear whether some release occurs, 

and no stack measurements have yet been made. Therefore, insufficient evidence is available to 

make even a preliminary estimate of releases, and this activity is classified as a Category E 

source. 

9-19
 
10. SOURCES OF DIOXIN-LIKE POLYCHLORINATED BIPHENYLS (PCBs) 

The purpose of this chapter is twofold: (1) to identify sources that release dioxin-like 

PCB congeners into the environment and (2) to derive national estimates for releases from these 

sources in the United States. PCBs have been found in all media and in all parts of the world. 

PCBs were manufactured in relatively large quantities for use in commercial products such as 

dielectrics, hydraulic fluids, plastics, coatings and paints, and although PCBs are no longer 

commercially produced in the United States, they continue to be released to the environment 

through the use and disposal of these products. PCBs may also be inadvertently produced as byproducts during the manufacture of certain organic chemicals and also as products of the 

incomplete combustion of some waste materials. 

10.1. GENERAL FINDINGS OF THE EMISSIONS INVENTORY 

Table 10-1 provides a compilation of known or suspected dioxin-like PCB-emitting 

source categories in the United States for which emission measurements of dioxin-like PCB 

congeners, Aroclors, or PCB congener groups have been reported in government, industry, and 

trade association reports; conference proceedings and journal articles; and comments submitted 

to EPA on previous versions of this document. The intent of Table 10-1 is to clearly identify 

those source categories and media (air, water, land, and products) for which the available data are 

adequate for reliably quantifying emissions of dioxin-like PCBs and those for which the data are 

inadequate. 

Nationwide emission estimates for the United States inventory are presented in 

Table 10-2 (emissions to air, water, land, and product) for those source categories for which 

estimates can be reliably quantified (the category has been assigned a confidence rating of A, B, 

or C) (see Section 1.2.3 for details on confidence ratings). Table 10-2 also lists preliminary 

estimates of the potential magnitude of emissions from “unquantified” sources (i.e., sources 

assigned a confidence rating of D) in reference year 2000. Because of large uncertainties for 

these Category D estimates, they are not included in the quantitative inventory. 

Currently, no significant releases of newly formed dioxin-like PCBs are occurring in the 

United States. Unlike CDDs/CDFs, PCBs were intentionally manufactured in the United States 

in large quantities from 1929 until production was banned in 1977. Releases to the environment 

of “old” dioxin-like PCBs (dioxin-like PCBs manufactured prior to the production ban) can 

occur from ongoing use and disposal practices. Prior to regulations enacted beginning in the late 

1970s that limited the manufacture/use/disposal of PCBs, significant quantities were released to 

the environment in association with (a) the manufacture of PCBs, (b) the manufacture of 

products containing PCBs, and (c) the use and disposal of products containing PCBs as well as 

10-1
 

Table 10-1. Confidence rating classes for 2000 for releases from all known and 

suspected source categories of dioxin-like PCBs

a 

Source category Air Land Water 

Approved PCB disposal E 

Accidental PCB releasees E E E 

Municipal wastewater treatment sludge A 

Municipal waste combustion E 

Industrial wood combustion E 

Medical waste incineration E 

Tire combustion E 

Cigarette combustion D 

Sewage sludge incineration C 

Backyard barrel burning E 

Petroleum refining catalyst regeneration E a

 Blank cells mean not applicable or no data. 

A= Characterization of the source category judged to be adequate for quantitative estimation with high confidence in 

the emission factor and high confidence in the activity level. 

C= Characterization of the source category judged to be adequate for quantitative estimation with low confidence in 

the emission factor and/or the activity level. 

D= These are preliminary indications of the potential magnitude of emissions from “unquantified” sources in 

Reference Year 1995. These estimates were assigned a “confidence category” rating of D and are not included in 

the Inventory. 

E= Not quantifiable. 

materials that may have been contaminated with trace levels of PCBs from prior PCB use or 

disposal. Following the ban on PCB production, releases from these first two categories ceased. 

The third type of releases, those associated with product use and disposal, will continue in at 

least four ways: 

1.	 Disposal of products containing greater than 2 lb of PCBs (e.g., dielectric fluids in 

transformers and large capacitors), which is controlled by disposal regulations that 

have minimized environmental releases; 

2.	 Disposal of products containing small quantities of PCBs (e.g., small capacitors, 

fluorescent lighting fixtures) or trace quantities of PCBs (e.g., wastepapers), which is 

subject to disposal as municipal solid waste but which may result in some release to 

the general environment; 

10-2
 
Table 10-2. Inventory of contemporary releases of dioxin-like PCBs in the United States for 1987, 1

2000 and preliminary release estimates of dioxin-like PCBs for 2000 (g TEQp-WHO98/yr) 

10-3
 


Emissions source category 2000 Inventory 1995 Inventory 1987 Inventory 

A B C A B C A B 

Releases to air Combustion sources 

Cigarettes Sewage sludge incineration 0.7 1.1 

Total quantified releases to air 0.7 1.1 Releases to land Municipal sludge (land application and farming) 18.8 77.4 51.1 Total quantified releases to land 18.8 77.4 51.1 

Releases to products Municipal sludge as soil ammendment 0.5 2.0 1.7 Total quantified releases to products 

0.5 2.0 1.7 

A= Characterization of the source category judged to be adequate for quantitative estimation with high confidence in the emission factor 

the activity level. 

C= Characterization of the source category judged to be adequate for quantitative estimation with low confidence in the emission factor D= These are preliminary indications of the potential magnitude of emissions from “unquantified” sources in Reference Year 1995. The

assigned a “confidence category” rating of D and are not included in the Inventory. 

E= Not quantifiable. 
3. Leaks and spills of still-in-service PCBs; and 

4. Illegal disposal of PCBs. 

Although it has been demonstrated that small quantities of dioxin-like PCBs can be 

emitted into the air during waste combustion, no strong evidence exists that they are emitted in 

significant quantities as by-products during combustion. The widespread occurrence of dioxinlike PCBs in the U.S. environment most likely reflects past releases associated with PCB 

production, use, and disposal. Further support for this finding is based on observations of 

reductions since the 1980s in PCB concentrations in Great Lakes sediment and in other areas. 

10.2. RELEASES OF COMMERCIAL PCBs 

PCBs were commercially manufactured by the direct batch chlorination of molten 

biphenyl with anhydrous chlorine in the presence of a catalyst, followed by separation and 

purification of the desired chlorinated biphenyl fractions. The degree of chlorination was 

controlled by the chlorine contact time in the reactor. Commercial PCB production is believed to 

have been confined to 10 countries. Total PCBs produced worldwide since 1929 (the first year 

of known production) has been estimated at 1.5 million metric tons. 

Initially, PCBs were used primarily as dielectric fluids in transformers. After World War 

II, PCBs found steadily increasing use as dielectric fluids in capacitors, as heat-conducting fluids 

in heat exchangers, and as heat-resistant hydraulic fluids in mining equipment and vacuum 

pumps. PCBs also were used in a variety of “open” applications (i.e, uses from which PCBs 

cannot be recollected), including plasticizers, carbonless copy paper, lubricants, inks, laminating 

agents, impregnating agents, paints, adhesives, waxes, additives in cement and plaster, casting 

agents, dedusting agents, sealing liquids, fire retardants, immersion oils, and pesticides (DeVoogt 

and Brinkman, 1989). 

U.S. production peaked in 1970, with a volume of 39,000 metric tons. In 1971, 

Monsanto Corporation, the major U.S. producer, voluntarily restricted the sale of PCBs for all 

applications, with the exception of “closed electrical systems.” Annual production fell to 18,000 

metric tons in 1974. Monsanto ceased PCB manufacture in mid-1977 and shipped the last 

inventory in October of that year. Regulations issued by EPA beginning in 1977, principally 

under the Toxic Substances Control Act (TSCA) (40 CFR 761), have strictly limited the 

production, import, use, and disposal of PCBs. The estimated cumulative production and 

consumption volumes of PCBs in the United States from 1930 to 1975 were 635.03 million kg 

produced, 1.36 million kg imported (primarily from Japan, Italy, and France), 568.35 million kg 

10-4
 

sold in the United States, and 68.04 million kg exported (ATSDR, 1993; DeVoogt and 

Brinkman, 1989). The reliability of these values is +5% and –20% (Versar, Inc., 1976). 

Monsanto Corporation marketed technical-grade mixtures of PCBs primarily under the 

trade name Aroclor. The Aroclor mixtures are identified by a four-digit numbering code in 

which the last two digits indicate the chlorine content by weight percent. The exception to this 

coding scheme is Aroclor 1016, which contains only mono- through hexachlorinated congeners 

with an average chlorine content of 41%. From 1957 until 1972, Monsanto also manufactured 

several blends of PCBs and polychlorinated terphenyls (PCTs) under the trade names Aroclor 

2565 and Aroclor 4465; manufacture and sales volumes are not available for these blends. Listed 

below are the percentages of total Aroclor production during the years 1957 to 1977 by Aroclor 

mixture, as reported by Brown (1994).

 1957–1977 

U.S. production

 Aroclor 	 (%) 

1016 12.88 

1221 0.96 

1232 0.24 

1242 51.76 

1248 6.76 

1254 15.73 

1260 10.61 

1262 0.83 

1268 0.33 

The trade names of the major commercial PCB technical-grade mixtures manufactured in 

other countries included Clophen (Germany), Fenclor and Apirolio (Italy), Kanechlor (Japan), 

Phenoclor and Pyralene (France), Sovtel (USSR), Delor and Delorene (Czechoslovakia), and 

Orophene (German Democratic Republic) (DeVoogt and Brinkman, 1989). The mixtures 

marketed under these trade names had similar chlorine content (by weight percent and average 

number of chlorines per molecule) to those of various Aroclor mixtures. Listed below are 

comparable mixtures in terms of chlorine content marketed under several trade names.

 Aroclor Clophen Pyralene Phenoclor Fenclor Kanechlor

 1232 2000 200

 1242 A-30 3000 DP-3 42 300

 1248 A-40 DP-4 400

 1254 A-50 DP-5 54 500

 1260 A-60 DP-6 64 600 

10-5
 
Major advances in analytical separation and resolution techniques beginning in the 1970s 

enabled various researchers to identify and quantify PCB congeners present in Aroclors, 

Clophens, and Kanechlors (Jensen et al., 1974; Albro and Parker, 1979; Huckins et al., 1980; 

Albro et al., 1981; Duinker and Hillebrand, 1983; Kannan et al., 1987; Tanabe et al., 1987; 

Duinker et al., 1988; Schulz et al., 1989; Himberg and Sippola, 1990; Larsen et al., 1992; deBoer 

et al., 1993; Schwartz et al., 1993; Frame et al., 1996a, b; Frame, 1997). Schulz et al. (1989) 

were the first to identify and quantify all PCB congeners present in a series of Aroclors and 

Clophens. Frame (1995) reported preliminary results of a nearly completed round robin study, 

one goal of which was to determine the distribution of all PCB congeners above 0.05 weight 

percent in various Aroclors (1221, 1016, 1242, 1260, and 1262) using 18 state-of-the-art gas 

chromatography/mass spectrometry (GC/MS) or electron capture detector (GC/ECD) systems. 

Table 10-3 presents mean summary statistics on the concentrations of the dioxin-like 

PCBs in each mixture group (e.g., Aroclor 1248, Clophen A-40, and Kanechlor 400 are in one 

mixture group) reported by these researchers. Table 10-3 also presents the mean TEQ 

concentration of each congener in each mixture group as well as the total mean TEQ 

concentration in the mixture group. Because of the wide variability in the reported results, the 

uncertainty associated with these mean concentrations is very large. 

For each mixture group, the congeners detected were generally similar. There was, 

however, wide variability in the concentrations reported by some researchers for some congeners. 

Brown et al. (1995) compiled similar statistics using a somewhat different set of studies and 

derived significantly lower mean concentrations of some congeners in several Aroclors. Frame 

(1995) and Larsen (1995) attributed such differences to either potential limitations in the GC 

columns used by various researchers to separate similar eluting congeners or actual differences in 

the congener concentrations in the Aroclor, Clophen, and Kanechlor lots analyzed by various 

research groups. 

The congener distributions also vary among the different mixtures. Therefore, the 

calculated TEQs also vary. The congener distributions for various lots of Aroclor 1254, and the 

corresponding TEQs, are presented in another study (Frame, 1999) in which the relative TEQs 

for late production lots were reported to be much higher than those for the earlier production lots; 

however, the late production lots were estimated to account for only about 1% of the total 

production volume of Aroclor 1254. Therefore, the data for the later production lots were not 

included in the average TEQ calculation for Aroclor 1254 in Table 10-3. 

In the environment, PCBs also occur as mixtures of congeners, but their composition 

differs from those of the commercial mixtures because after release to the environment the 

mixtures change over time through partitioning, chemical transformation, and preferential 

bioaccumulation (U.S. EPA, 1996f). Dioxin-like PCB congeners differ by up to one to two 

10-6
 

Table 10-3. Weight percent concentrations of dioxin-like PCBs in Aroclors, Clophens, and Kanech

10-7
 


Dioxin-like PCB congener 

IUPAC 

number 

No. of 

samples 

analyzed 

No. of 

detections 

Mean conc. 

(nondetect set to 

zero) (g/kg) 

TEQP-WHO98 conc. 

(nondetect set to 

zero) (mg/kg) 

Mean conc.a 

(nondetect set to 

detection limi

(g/kg) 

Aroclor 1016 3,3',4,4'-TCB 

 3,4,4',5-TCB 

 2,3,3',4,4'-PeCB 

 2,3,4,4',5-PeCB 

 2,3',4,4',5-PeCB 

 2',3,4,4',5-PeCB 

 3,3',4,4',5-PeCB 

 2,3,3',4,4',5-HxCB 

 2,3,3',4,4',5'-HxCB 

 2,3',4,4',5,5'-HxCB 

 3,3',4,4',5,5'-HxCB 

 2,2',3,3',4,4',5-HpCB 

 2,2',3,4,4',5,5'-HpCB 

 2,3,3',4,4',5,5'-HpCB 

77 

81 

105 

114 

118 

123 

126 

156 

157 

167 

169 

170 

180 

189 

5 

3 

4 

4 

4 

4 

4 

4 

4 

4 

5 

4 

4 

4 

0 

0 

1 

0 

1 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0.0375 

0 

0.0125 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0.00375 

0 

0.00125 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0.109 

0 

0.091 

0 

0 

0 

0 

0 

0 

0 

0 

0 

Total TEQP-WHO98 0.005 

Total TEQP-WHO94 0.005 Aroclor 1221 3,3',4,4'-TCB 

 3,4,4',5-TCB 

 2,3,3',4,4'-PeCB 

 2,3,4,4',5-PeCB 

 2,3',4,4',5-PeCB 

 2',3,4,4',5-PeCB 

 3,3',4,4',5-PeCB 

 2,3,3',4,4',5-HxCB 

 2,3,3',4,4',5'-HxCB 

 2,3',4,4',5,5'-HxCB 

 3,3',4,4',5,5'-HxCB 

 2,2',3,3',4,4',5-HpCB 

 2,2',3,4,4',5,5'-HpCB 

 2,3,3',4,4',5,5'-HpCB 

77 

81 

105 

114 

118 

123 

126 

156 

157 

167 

169 

170 

180 

189 

4 

4 

4 

4 

4 

4 

4 

4 

4 

4 

4 

3 

3 

4 

4 

1 

3 

0 

4 

0 

0 

0 

0 

0 

0 

0 

0 

0 

1.075 

0.0875 

0.3875 

0 

1.725 

0 

0 

0 

0 

0 

0 

0 

0 

0 

0.1075 

0.00875 

0.03875 

0 

0.1725 

0 

0 

0 

0 

0 

0 

0 

0 

0 

1.078 

0.116 

0.4 

0 

1.725 

0 

0 

0 

0 

0 

0 

0 

0 

0 

Total TEQP-WHO98 0.328 

Total TEQP-WHO94 0.749 
Table 10-3. Weight percent concentrations of dioxin-like PCBs in Aroclors, Clophens, and Kanechl

10-8
 


Dioxin-like PCB congener 

IUPAC 

number 

No. of 

samples 

analyzed 

No. of 

detections 

Mean conc. 

(nondetect set to 

zero) (g/kg) 

TEQP-WHO98 conc. 

(nondetect set to 

zero) (mg/kg) 

Mean conc.a 

(nondetect set to 

detection limi

(g/kg) 

Aroclor 1242, Clophen A-30, and Kanechlor 300 3,3',4,4'-TCB 77 15 15 3.3 0.33 3.301 

 3,4,4',5-TCB 8 7 6 1.09 0.11 1.089 

 2,3,3',4,4'-PeCB 105 11 11 4.02 0.4 4.024  2,3,4,4',5-PeCB 114 8 5 1.13 0.57 1.201 

 2,3',4,4',5-PeCB 118 9 9 8.04 0.8 8.044 

 2',3,4,4',5-PeCB 123 9 7 1.12 0.11 1.157 

 3,3',4,4',5-PeCB 126 14 8 0.049 4.94 0.094  2,3,3',4,4',5-HxCB 156 9 8 0.39 0.2 0.424 

 2,3,3',4,4',5'-HxCB 157 8 2 0.021 0.011 0.096 

 2,3',4,4',5,5'-HxCB 167 8 2 0.021 0.00021 0.096 

 3,3',4,4',5,5'-HxCB 169 14 2 0.000013 0.00013 0.048  2,2',3,3',4,4',5-HpCB 170 6 2 0.19 0 0.244 

 2,2',3,4,4',5,5'-HpCB 180 5 2 0.16 0 0.218 

 2,3,3',4,4',5,5'-HpCB 189 7 0 0 0 0 

Total TEQP-WHO98 7.47 

Total TEQP-WHO94 8.70 Aroclor 1248, Clophen A-40, and Kanechlor 400  3,3',4,4'-TCB 

 3,4,4',5-TCB  2,3,3',4,4'-PeCB 

 2,3,4,4',5-PeCB  2,3',4,4',5-PeCB 

 2',3,4,4',5-PeCB  3,3',4,4',5-PeCB  2,3,3',4,4',5-HxCB 

 2,3,3',4,4',5'-HxCB 

 2,3',4,4',5,5'-HxCB 

 3,3',4,4',5,5'-HxCB  2,2',3,3',4,4',5-HpCB 

 2,2',3,4,4',5,5'-HpCB 

 2,3,3',4,4',5,5'-HpCB 

77 

81 

105 

114 118 

123 126 156 

157 

167 

169 170 

180 

189 

13 

6 9 

7 8 

7 

11 

8 

7 

7 

12 

5 

4 

6 

13 

4 8 

6 8 

7 

6 

8 

3 

3 3 

4 

4 

1 

4.36 

1.76 

10.12 

3.39 

20.98 

1.48 

0.11 

1.13 

0.19 

0.16 

0.01 0.96 

1.24 

0.0018 

0.44 

0.18 1.01 

1.69 2.1 

0.15 

10.55 

0.56 

0.09 

0.0016 

0.1006 0 

0 

0.0001833 

4.36 

1.77 

10.12 

3.4 

20.98 

1.48 0.14 1.13 

0.2 

0.16 

0.041 0.97 

1.24 

0.06 

Total TEQP-WHO98 16.87 

Total TEQP-WHO94 18.55 
Table 10-3. Weight percent concentrations of dioxin-like PCBs in Aroclors, Clophens, and Kanechl

10-9
 


Dioxin-like PCB congener 

IUPAC 

number 

No. of 

samples 

analyzed 

No. of 

detections 

Mean conc. 

(nondetect set to 

zero) (g/kg) 

TEQP-WHO98 conc. 

(nondetect set to 

zero) (mg/kg) 

Mean conc.a 

(nondetect set to 

detection limi

(g/kg) 

Aroclor 1254, Clophen A-50, and Kanechlor 500  3,3',4,4'-TCB 

 3,4,4',5-TCB  2,3,3',4,4'-PeCB  2,3,4,4',5-PeCB  2,3',4,4',5-PeCB  2',3,4,4',5-PeCB  3,3',4,4',5-PeCB  2,3,3',4,4',5-HxCB  2,3,3',4,4',5'-HxCB 

 2,3',4,4',5,5'-HxCB  3,3',4,4',5,5'-HxCB  2,2',3,3',4,4',5-HpCB 

 2,2',3,4,4',5,5'-HpCB 

 2,3,3',4,4',5,5'-HpCB 

77 

81 

105 114 118 123 126 156 157 

167 169 170 

180 

189 

15 

6 

12 

9 

11 

8 

14 10 

9 

10 14 

8 

7 

7 

12 

1 

11 

6 

11 

8 

12 10 

8 9 6 

8 

7 

2 

0.8 

7.85 

35.83 12.17 81.65 

4.59 

0.99 

11.08 

1.91 

2.74 0.08 5.06 

5.79 

0.045 

0.0795 

0.79 3.58 6.08 8.17 0.46 

99.46 

5.54 0.95 

0.0274 0.8 0 

0 

0.0045429 

0.83 

7.86 

35.83 12.23 81.65 

4.59 1.02 

11.08 

1.93 

2.74 0.12 5.06 

5.79 

0.13 

Total TEQP-WHO98 125.94 

Total TEQP-WHO94 126.04 Aroclor 1260, Clophen A-60, and Kanechlor 600 3,3',4,4'-TCB 

 3,4,4',5-TCB 

 2,3,3',4,4'-PeCB  2,3,4,4',5-PeCB 

 2,3',4,4',5-PeCB  2',3,4,4',5-PeCB 

 3,3',4,4',5-PeCB  2,3,3',4,4',5-HxCB  2,3,3',4,4',5'-HxCB 

 2,3',4,4',5,5'-HxCB  3,3',4,4',5,5'-HxCB  2,2',3,3',4,4',5-HpCB  2,2',3,4,4',5,5'-HpCB  2,3,3',4,4',5,5'-HpCB 

77 

81 

105 114 

118 123 

126 156 157 

167 169 170 180 189 

15 

6 

11 

9 

11 

8 

14 11 

8 

10 14 

8 7 8 

6 

1 

10 

4 

10 

1 7 

11 

8 9 5 

8 7 8 

0.13 

0.08 1.59 

0.71 9.51 

0.0005 1.81 6.89 

1.59 2.87 0.16 

32.94 82.61 

1.74 

0.01256 

0.0075 0.16 

0.35 0.95 

0.00005 

180.89 

3.45 

0.79 0.03 1.64 0 0 

0.1739792 

0.17 

0.1 1.59 

0.77 9.51 

0.08 1.84 6.89 

1.59 2.87 0.19 

32.94 82.61 

1.74 

Total TEQP-WHO98 188.45 

Total TEQP-WHO94 192.62 
Table 10-3. Weight percent concentrations of dioxin-like PCBs in Aroclors, Clophens, and Kanechl

aCalculated for a congener only when at least one sample contained detectable levels of that congener. 

Sources: Adapted from Schulz et al. (1989); Duinker and Hillebrand (1983; deBoer et al. (1993); Schwartz et al. (1993); Larsen, et al. (1987); Huckins et al. (1980); Albro and Parker (1979; Jensen et al. (1974); Albro et al. (1981); Duinker et al. (1988); Tanabe et al. (19Sippola (1990); Frame et al. (1996a, b); Frame (1997). 

10-10
 

orders of magnitude in their water solubility, vapor pressure, Kow value, and Henry's Law 

constant. Thus, although all the dioxin-like PCB congeners are poorly soluble in water and have 

very low vapor pressures, they will volatilize and leach at different rates. Similarly, because the 

congeners differ somewhat in their rates of biodegradation, bioaccumulation, and 

photodegradation, the congener patterns found in environmental media and biota will vary from 

those found in commercial mixtures. 

Although environmental mixtures are often characterized in terms of Aroclors, this 

characterization can be both imprecise and inappropriate. Qualitative and quantitative errors can 

arise from judgements in comparing GC/MS peaks for a sample with the characteristic peak 

patterns for different Aroclors, particularly for environmentally altered patterns (U.S. EPA, 

1996f). For the same reason, it can be both imprecise and inappropriate to infer concentrations 

of dioxin-like PCB congeners in an environmental sample on the basis of characterization of the 

sample’s Aroclor content and knowledge of the dioxin-like congener content in the commercial 

Aroclor. Safe (1994) wrote, “Regulatory agencies and environmental scientists have recognized 

that the composition of PCBs in most environmental extracts does not resemble the compositions 

of the commercial product.” Similarly, ATSDR (1993) stated, “It is important to recognize that 

the PCBs to which people may be exposed are likely to be different from the original PCB source 

because of changes in congener and impurity composition resulting from differential partitioning 

and transformation in the environment and differential metabolism and retention.” 

10.2.1. Approved PCB Disposal/Destruction Methods 

In 1978, EPA began regulating the disposal of PCBs and PCB-contaminated waste under 

TSCA, PL 94-469. The disposal regulations, published in the Code of Federal Regulations, 40 

CFR, Part 761, state that the preferred disposal method is incineration at 1,200°C or higher. If 

the waste contains material that cannot be destroyed by incineration, EPA clearance must be 

obtained to dispose of the waste in a chemical waste landfill or by another approved manner. 

The PCB disposal regulations describe disposal of three distinct types of PCB waste: 

PCBs, PCB articles (items containing PCBs), and PCB containers. Within these categories, 

further distinctions are made on the basis of the PCB concentration in the waste, with the 

acceptable disposal methods being based on the concentrations in the specific waste to be 

destroyed. The acceptable disposal methods are Annex I incinerators, high-efficiency boilers, 

Annex II chemical waste landfills, and other approved methods. The following paragraphs and 

Table 10-4 provide brief descriptions of these disposal methods. More complete descriptions of 

the specific methodologies are provided in 40 CFR, Part 761. 

10-11
 
Table 10-4. Disposal requirements for PCBs and PCB items 

10-12
 


PCBs/items Waste characterization Disposal requi

PCBs Mineral oil dielectric fluids from 

PCB transformers 

Those analyzing >500 ppm PCB 

Annex I incineratora 

Mineral oil dielectric fluids from PCB-contaminated transformers 

Those analyzing 50–500 ppm PCB 

Annex I incinerator High-efficiency boiler (40 CFR Other approved incineratorb 

Annex II chemical waste landf

PCB liquid wastes other than mineral oil dielectric fluid 

Those analyzing >500 ppm PCB Those analyzing 50–500 ppm PCB 

Annex I incinerator Annex I incinerator High-efficiency boiler (40 CFR Other approved incineratorb 

Annex II chemical waste landf

Nonliquid PCB wastes (e.g., contaminated materials from spills) 

Annex I incinerator Annex II chemical waste landf

Dredged materials and municipal sewage treatment sludges containing PCBs 

Annex I incinerator Annex II chemical waste landfOther approved disposal meth761.10(a)(5)(iii) 

PCB articles Transformers PCB transformers 

PCB contaminated transformers 

Annex I incinerator Drained and rinsed transformein Annex II chemical waste lanDisposal of drained transform

PCB capacitorsd Annex I incinerator 

PCB hydraulic machines Those containing >1,000 

ppm PCB Those containing <1,000 ppm PCB 

Drained and rinsed machines mmunicipal solid waste or salvaDrained machines may be dispsolid waste or salvaged 
Table 10-4. Disposal requirements for PCBs and PCB items (continued) 

10-13
 


PCBs/items Waste characterization Disposal requi

Other PCB articles Those containing PCB fluids 

Those not containing PCB fluids 

Drained machines may be dispAnnex II Annex I incinerator or Annex landfill 

PCB containers Those used to contain only PCBs 

at a concentration <500 ppm 

As municipal solid waste provare drained prior to disposal 

Other PCB containers Annex I incinerator 

Annex II, provided any liquid prior to disposal Decontaminate per Annex IV 

aAnnex I incinerator is defined in 40 CFR 761.40. bRequirements for other approved incinerators are defined in 40 CFR 761.10(e). cAnnex II chemical waste landfills are described in 40 CFR 761.41. Annex II disposal is permitted if the PCB waste contains less than 5

 not ignitable as per 40 CFR Part 761.41(b)(8)(iii). 

dDisposal of containerized capacitors in Annex II landfills was permitted until March 1, 1981; thereafter, only Annex I incineration has 

Source: U.S. EPA (1987d). 
10.2.1.1. Approved Incinerators/High-Efficiency Boilers 

PCB Annex I incinerators must meet the specific technical standards and criteria listed in 

Annex I of EPA’s PCB regulations. The minimum operating requirements for disposal of liquid 

wastes are 2 sec at 1,200°C with 3% excess oxygen (measured in the stack gas) or 1.5 sec at 

1,600°C with 2% excess oxygen (measured in the stack gas). Monitoring requirements, approval 

conditions, and trial burn requirements are prescribed in Annex I. Operators of commercial or 

industrial incinerators who intend to destroy liquid PCB wastes must demonstrate the 

incineration’s compliance with the Annex I requirements through a comprehensive trial burn 

program. Annex I incinerators operating at optimum performance level should destroy 99.997% 

of liquid PCB waste, with a resulting maximum emission factor of 0.03 g/kg. 

Criteria for Annex I incinerators were established for the destruction of liquid PCB 

wastes; however, these incinerators also may be used for disposal of nonliquid PCB items (such 

as capacitors), provided that a destruction and removal efficiency of 99.9999% and a maximum 

emission factor of 0.001 g/kg are met. 

High-efficiency boilers may be used to destroy PCBs and PCB-contaminated waste with 

PCB concentrations not exceeding 500 ppm. Conventional industrial and utility boilers may be 

designated as high-efficiency boilers if they are operated under the prescribed combustion 

conditions defined in the PCB disposal regulations. The PCB regulations do not specify a 

minimum destruction efficiency for high-efficiency boilers; however, EPA-approved boilers 

operated according to the regulations have reported destruction efficiencies in excess of 99.99%, 

with a corresponding maximum emission factor of 0.1 g/kg (U.S. EPA, 1987d). 

10.2.1.2. Approved Chemical Waste Landfills 

Approved chemical waste landfills can be used for the disposal of some but not all PCB 

wastes. PCB-contaminated materials acceptable for land disposal in an approved landfill include 

PCB mixtures (e.g., certain PCB-contaminated soil/solid debris, PCB-contaminated dredged 

materials, and PCB-contaminated municipal sewage sludge), PCB articles that cannot feasibly be 

incinerated (e.g., drained and flushed transformers), and drained PCB containers. Written 

approval must be obtained from EPA in order to landfill PCB articles other than transformers. 

PCB-contaminated materials not acceptable for land disposal in an approved landfill include 

nonliquid PCB mixtures in the form of contaminated soil, rags, or other solid debris, and sealed 

capacitors. Typically, PCBs disposed of in these landfills are placed in sealed containers, thereby 

minimizing any PCB emissions. 

10-14
 

10.2.1.3. Other Approved Disposal Methods 

Other thermal and nonthermal destruction techniques may be approved by EPA Regional 

Administrators if these processes can effect a level of destruction of PCBs equivalent to that of 

incinerators or boilers. After April 29, 1983, all other PCB disposal technologies (thermal and 

nonthermal) used in more than one EPA Region had to be approved by EPA Headquarters. 

Examples of thermal technologies approved for commercial-scale use or for research and 

development projects include a pyrolysis process to treat contaminated soils, a fluid wall reactor, 

a cement kiln, a diesel engine, a steam-stripping operation, an aluminum melting furnace, and a 

molten salt process. Examples of approved nonthermal processes include chemical 

dechlorination processes, physical/chemical extraction techniques, and biological reduction 

methods. The physical/chemical techniques extract the PCBs from transformers or capacitors 

and concentrate them for disposal; they do not destroy the PCBs. 

10.2.2. Emission Estimates 

Tables 10-5 and 10-6 list the amounts of PCBs reported in EPA’s Toxics Release 

Inventory (TRI) as transferred off site for treatment, energy recovery, or disposal and the amounts 

released between 1988 and 2000, respectively. These quantities do not necessarily represent 

entry of PCBs into the environment. If it is assumed that all transferred PCBs are incinerated in 

high-efficiency boilers with a destruction and removal efficiency of 99.99%, then annual 

emissions of PCBs to air during 1988, 1995, and 2000 could have been as high as 264 kg, 31 kg, 

and 15 kg, respectively. Because no stack testing data are available for dioxin-like PCBs, it is 

not possible to estimate what fraction of these potential PCB releases would have been dioxinlike congeners. 

10.2.3. Accidental Releases of Still-in-Service PCBs 

After the 1977 ban on production of PCBs, releases of commercially produced PCBs to 

the environment (aside from minimal releases occurring during approved disposal or destruction) 

have been limited to accidental release of in-service PCBs (U.S. EPA, 1987d). Accidental 

releases are the result of leaks or spills during failure/breakage of an existing piece of PCBcontaining equipment or of incomplete combustion during accidental fires involving PCBcontaining equipment. These two types of accidental releases are discussed in the following 

sections. 

10-15
 

Table 10-5. Off-site transfers of PCBs reported in the Toxics Release 

Inventory (TRI) (1988–2000) 

Year 

No. of TRI 

forms filed 

Reported transfers (kg) 

Transfers to 

POTWs 

Transfers for 

treatment/disposal Total transfers 

2000 NA 102 150,888 150,990 

1999 NA 0 434,666 434,666 

1998 NA 0 386,903 386,903 

1997 NA a 471,319 471,319 

1996 NA 0 160,802 160,802 

1995 NA 0 308,347 308,347 

1994 NA 0 466,948 466,948 

1993 16 120 463,385 463,505 

1992 20 0 766,638 766,638 

1991 26 0 402,535 402,535 

1990 NA 0 1,181,961 1,181,961 

1989 NA 0.5 2,002,237 2,002,237 

1988 122 113 2,642,133 2,642,246 a

Facilities left that particular cell blank on the Form R submissions. 

NA = Not available 

POTWs = Publicly owned treatment works 

Sources: U.S. EPA (1993f, 1995h, 1998b, 2003c). 

10.2.3.1. Leaks and Spills 

PCBs that remain in active service at this time are those contained in “closed systems” 

(i.e., those pieces of electrical equipment that completely enclose the PCBs and do not provide 

direct atmospheric access for the PCBs during normal use). This equipment includes PCB 

transformers, capacitors, voltage regulators, circuit breakers, and reclosures. With the exception 

of PCB transformers—and probably small PCB capacitors—the majority of the PCB-containing 

electrical equipment in service during 1981 was owned by the electrical utility industry. 

Approximately 70% of the estimated 140,000 PCB transformers in service in 1981 were owned 

by nonutilities. No information was available on the relative distribution of small PCB 

capacitors (Versar, Inc., 1988). 

10-16
 
Table 10-6. Releases of PCBs reported in the Toxics Release Inventory (TRI) (1988–2000) 

10-17
 


Reported releases (kg) 

Fugitive or On-site 

No. of TRI nonpoint air Stack or point Surface water Underground releases 

Year forms filed emissions air emissions discharges injection land 

2000 NA 158 2,497 13 0.5 648,128 

1999 NA 0 0 a a 

1998 NA 0 0 0 a 60,854 

1997 NA 0 0 0 a 3,081 1996 NA 2.3 114 0 0 4,179 

1995 NA 0 0 0 0 0 

1994 NA 0 0 0 0 0 

1993 16 0 0 0 0 120 

1992 20 0 0 0 0 0.5 

1991 26 0 0 0 0 0 

1990 NA 2.3 0 0 0 32,372 1989 NA 0 0 120 0 453 1988 122 2.7 0 4.5 0 341 

aFacilities left that particular cell blank on the Form R submissions. 

NA = Not available Sources: U.S. EPA (1993f, 1995h, 1998b, 2003c). 
The number of each of these items owned by the utility industry, the quantity of PCBs 

contained in each, and an estimate of the annual quantity of PCBs leaked and/or spilled were 

investigated by the Edison Electric Institute and the Utility Solid Wastes Activity Group 

(EEI/USWAG) for EPA in 1981. The findings of this investigation, which were reported in a 

proposed modification to the PCB regulations (Federal Register, 1982a), indicated that more than 

99% of the total quantity of PCBs contained in utility-owned electrical equipment in 1981 

(73,700 metric tons) was in 40,000 PCB transformers (those containing >500 ppm of PCBs) and 

large PCB capacitors (those containing >3 lb of PCBs). An upper-bound estimate of the mass of 

PCBs that leached or spilled from this equipment in 1981 was 177 metric tons. Approximately 

95% of the estimated releases were the result of leaks from large PCB capacitors (Federal 

Register, 1982a). Leaks/spills typically occur in transformers when the gasket joining the top to 

the body corrodes, tears, or physically fails. PCBs can then leak past this failed section and 

potentially spill onto the surrounding ground. PCB capacitors typically fail by rupturing, 

exposing the contained PCBs to the environment. Failure is caused by environmental and 

weathering effects (e.g., lightning) or material failures (e.g., metal fatigue). 

As of mid-1988, the total population of in-service PCB transformers and large PCB 

capacitors was estimated to have decreased from 140,000 to 110,000 and from 3.3 million to 1.9 

million, respectively (Versar, Inc., 1988). PCB transformers have normal operating lifetimes of 

30 years and 40 years, respectively. EPA’s PCB Electrical Use Rule (Federal Register, 1982b) 

required the removal of 950 food/feed industry transformers by 1985 and 1.1 million 

unrestricted-access large PCB capacitors by October 1988. In addition, EPA’s PCB Transformer 

Fires Rule (Federal Register, 1985b) required the removal by 1990 of 7,600 480-volt network 

transformers. 

More recent inventories of PCB-containing electrical equipment are not available. 

However, an Information Collection Request submitted by EPA to the Office of Management 

and Budget for information on uses, locations, and conditions of PCB electrical equipment 

estimated that there may be 150,000 owners of PCB-containing transformers used in industry, 

utilities, government buildings, and private buildings (Federal Register, 1997b). It is expected, 

and is demonstrated by the reported PCB transfers in the EPA’s TRI (see Table 10-5), that many 

owners of PCB electrical equipment have removed PCB-containing equipment to eliminate 

potential liability. 

10.2.3.2. Accidental Fires 

The available information is not adequate to support an estimate of potential annual 

releases of dioxin-like PCBs from accidental electrical equipment fires. For fires involving PCB 

transformers or capacitors, the amount of PCBs released is dependent on the extensiveness of the 

10-18
 

fire and the speed at which it is extinguished. A number of these fires are documented. A New 

York fire involving 200 gal of transformer fluid containing some 65% by weight PCBs resulted 

in a release of up to 1,300 lb of PCBs. A capacitor fire that burned uncontrolled for 2 hr in 

Sweden resulted in the destruction of 12 large utility capacitors containing an estimated 25 

pounds each of PCBs, for a total potential release of 300 lb. However, data are incomplete on the 

exact amount of PCBs released as a result of these two fires. 

EPA has imposed reporting requirements to ensure that the National Response Center is 

informed immediately of fires involving PCB transformers (40 CFR 761). The recordkeeping 

requirements are used to document the use, location, and condition of PCB equipment. 

Responses are mandatory, but the submitter may claim them to be confidential information. The 

number of PCB transformer fires is estimated to be approximately 20 per year; the number of 

PCB capacitor fires is unknown (U.S. EPA, 1987d). As these PCB-containing items reach the 

end of their useful lives and are retired, their susceptibility to fires will be eliminated, and the 

overall number of PCB transformer and capacitor fires will be reduced. 

10.2.4. Municipal Wastewater Treatment 

EPA conducted the National Sewage Sludge Survey in 1988 and 1989 to obtain national 

data on sewage sludge quality and management. As part of this survey, EPA tested for more than 

400 analytes, including seven of the Aroclors, in sludges from 175 publicly owned treatment 

works (POTWs) that employed at least secondary wastewater treatment. Sludges from 19% of 

the POTWs had detectable levels of at least one of the following Aroclors: 1248, 1254, or 1260; 

none of the other Aroclors were detected in any sample (the detection limit [DL] was typically 

about 200 μg/kg dry weight) (U.S. EPA, 1996e). Analyses were not performed for dioxin-like 

PCB congeners. The Aroclor-specific results of the survey are presented in Table 10-7. 

Gutenmann et al. (1994) reported similar results in a survey of sludges from 16 large U.S. 

cities for Aroclor 1260 content. At a DL of 250 μg/kg (dry weight), the investigators detected 

Aroclor 1260 (4,600 μg/kg) at only one facility. These results indicate that PCBs are not likely 

to be formed at POTWs, but rather are present because of disposal of PCB products or 

recirculation of previously disposed of PCBs. 

Although PCBs, measured as Aroclors, were not commonly detected in sewage sludge at 

microgram-per-kilogram levels in studies by EPA (U.S. EPA, 1996e) and Gutenmann et al. 

(1994), the presence of dioxin-like PCB congeners at lower concentrations may be more 

common. Green et al. (1995) and Cramer et al. (1995) reported the results of analyses of 99 

samples of sewage sludge for PCB congener numbers 77, 81, 126, and 169. The sludge samples 

were collected from 74 wastewater treatment plants across the United States during the summer 

10-19
 

Table 10-7. Aroclor concentrations (ng/kg) measured in EPA’s National 

Sewage Sludge Surveya 

Aroclor 

Percent 

detected 

Maximum 

concentration 

Median concentration 

Nondetects set to 

detection limit 

Nondetects set 

to zero 

1016 0 -­ -­ 0 

1221 0 -­ -­ 0 

1232 0 -­ -­ 0 

1242 0 -­ -­ 0 

1248 9 5.2 0.209 0 

1254 8 9.35 0.209 0 

1260 10 4.01 0.209 0 

Any Aroclor (total) 19 14.7 1.49 0 a

For publicly owned treatment works with multiple samples, the pollutant concentrations were averaged before the 

summary statistics presented in the table were calculated. 

-- = No information given 

Source: U.S. EPA (1996e). 

of 1994. These data are summarized in Table 10-8. Results from all samples collected from the 

same facility were averaged by Green et al. and Cramer et al. to ensure that results were not 

biased toward the concentrations found at facilities from which more than one sample was 

collected. If all nondetect values were assumed to be zero, then the POTW mean TEQP-WHO94 

and TEQP-WHO98 concentrations were 25.1 and 24.2 ng TEQ/kg (dry-weight basis), respectively. 

If the nondetect values were set equal to the DLs, then the POTW mean TEQP-WHO94 and 

TEQP-WHO98 concentrations were 25.2 and 24.3 ng TEQ/kg, respectively. 

In 1999, sewage sludge samples from a POTW in Ohio were collected and analyzed for 

PCBs (U.S. EPA, 2000b). The facility, which accepts both domestic and industrial wastewater, 

employs secondary wastewater technology. Assuming nondects were zero, the mean TEQ 

emission factor was 141 ng TEQP-WHO98/kg. These results are presented in Table 10-9. 

In 2000 and 2001, the Association of Metropolitan Sewage Agencies conducted a survey 

of dioxin-like PCB compounds in sewage sludge (Alvarado et al., 2001). A total of 200 sewage 

sludge samples were collected from 171 POTWs located in 31 states. Assuming nondetects were 

zero, the mean and median TEQ emission factors were reported as 8.3 and 3.37 ng TEQPWHO98/kg, respectively. 

10-20
 
Table 10-8. Dioxin-like PCB concentrations measured in sludges collected from 74 U.S. publicly ow

treatment works (POTWs) during 1994a,b 

10-21
 


Congener 

IUPAC 

number 

Percent 

detected 

Maximum 

concentration 

(ng/kg) 

Median concentration (ng/kg) Mean co

Nondetect set to 

1⁄2 detection 

limit 

Nondetect set to zero 

Nondetect 

to 1⁄2 detect

limit 

3,3',4,4'-TCB 77 100 22,900 783 783 2,243 

3,4,4',5-TCB 81 86 1,250 27.3 27 65.2 2,3,3',4,4'-PeCB 105 2,3,4,4',5-PeCB 114 2,3',4,4',5-PeCB 118 2',3,4,4',5-PeCB 123 3,3',4,4',5-PeCB 126 99 3,020 91.6 91.6 237 

2,3,3',4,4',5-HxCB 156 2,3,3',4,4',5'-HxCB 157 2,3',4,4',5,5'-HxCB 167 3,3',4,4',5,5'-HxCB 169 22 1,470 8.5 0 32.5 2,2',3,3',4,4',5-HpCB 170 2,2',3,4,4',5,5'-HpCB 180 2,3,3',4,4',5,5'-HpCB 189 Total TEQP-WHO98 9.3 9.2 24.3 

aFor POTWs with multiple samples, the sample concentrations were averaged by Cramer et al. (1995) to POTW averages before calcula

 mean and median values presented in the table. The TEQP-WHO94 and TEQP-WHO98 values were calculated on a facility-level basis. 

bBlank cells indicate that no measurements of these congeners were made. 

Source: Green et al. (1995); Cramer et al. (1995). 
Table 10-9. Dioxin-like PCB concentrations in sewage sludge collected from 

U.S. publicly owned treatment works during 1999 

Congener 

IUPAC 

number 

Mean emission factor (ng/kg) 

Nondetect set to 1⁄2 

detection limit Nondetect set to zero 

3,3',4,4'-TCB 77 42,467 42,467 

2,3,3',4,4'-PeCB 105 7,230 7,230 

2,3,4,4',5-PeCB 114 701 701 

2',3,4,4',5-PeCB 123 249 249 

2,3',4,4',5-PeCB 118 12,867 12,867 

3,3',4,4',5-PeCB 126 1,270 1,270 

2,3,3',4,4',5-HxCB 156 1,843 1,843 

2,3,3',4,4',5'-HxCB 157 524 524 

2,3',4,4',5,5'-HxCB 167 935 935 

3,3',4,4',5,5'-HxCB 169 570 570 

2,2',3,3',4,4',5-HpCB 170 2,627 2,627 

2,2',3,4,4',5,5'-HpCB 180 6,497 6,497 

2,3,3',4,4',5,5'-HpCB 189 199 199 

Total TEQP-WHO98 141 141 

Source: U.S. EPA (2000b). 

For 2001, EPA conducted another National Sewage Sludge Survey to characterize the 

dioxin and dioxin-like equivalence levels in biosolids produced by the 6,857 POTWs operating 

in the United States in 2001 (U.S. EPA, 2002d). Sewage sludge samples were collected from 94 

POTWs that used secondary or higher treatment practices. All the facilities had been sampled as 

part of the 1988/1989 National Sewage Sludge Survey. To determine the mean and median TEQ 

emission estimates of the dioxin-like PCBs, EPA weighted the values on the basis of wastewater 

flow rates of all POTWs in the United States (i.e., number of facilities with wastewater flow rate 

>100 mg/day, >10 but #100 mg/day, >1 but #10 mg/day, and #1 mg/day). The weighted mean 

and median TEQP-WHO98 concentrations of the dioxin-like PCB congeners were 5.22 and 2.05 

ng/kg, respectively. 

According to the results of its 1988/1989 National Sewage Sludge Survey, EPA estimated 

that approximately 5.4 million dry metric tons of sewage sludge were generated in 1989 (Federal 

Register, 1993a). EPA also used the results of the 1984 to 1996 Clean Water Needs Surveys to 

10-22
 

estimate that 6.3 million dry metric tons of sewage sludge were generated in 1998 and 6.6 

million dry metric tons were generated in 2000 (U.S. EPA, 1999b). Because estimates for 1987 

and 1995 are not available, the 1989 and 1998 activity level estimates are used for reference 

years 1987 and 1995, respectively. Tables 10-10, 10-11, and 10-12 list the volume, by use and 

disposal practices, of sludge disposed of annually for reference years 1989, 1995, and 2000. 

Table 10-10. Quantity of sewage sludge disposed of annually in 1989 by 

primary, secondary, or advanced treatment publicly owned treatment works 

(POTWs) and potential dioxin-like PCB TEQ releases 

Use/disposal practice 

Volume 

disposed of 

(1,000 dry 

metric tons/yr) 

Percent 

of 

total 

volume 

Potential TEQP ­

WHO98 releasea 

(g of TEQ/yr) 

Potential TEQP ­

WHO94 releasea 

(g of TEQ/yr) 

Land application 1,714 32b 41.5 43 

Distribution and marketing 71 1.3 1.7 1.8 

Surface disposal site/other 396 7.4 9.6 9.9 

Sewage sludge landfill 157 2.9 4.2 3.9 

Co-disposal landfillsc 1,819 33.9 44 45.6 

Sludge incinerators and coincineratorsd 865 16.1 e e 

Ocean disposalf 336 6.3 0 0 

TOTAL 5,358 100 101 104.2 

a

Potential TEQ release for nonincinerated sludges was estimated by multiplying the sludge volume generated

 (column 2) by the mean dioxin-like PCB TEQ concentration in 74 POTW sludges reported by Green et al. (1995) 

 and Cramer et al. (1995) (i.e., 24.2 ng TEQP-WHO98/kg and 25.1 ng TEQP-WHO94/kg). b

Includes 21.9% applied to agricultural land, 2.8% applied as compost, 0.6% applied to forestry land, 3.1%

 applied to “public contact” land, 1.2% applied to reclamation sites, and 2.4% applied in undefined settings. c

Landfills used for disposal of sewage sludge and solid waste residuals. d

Co-incinerators treat sewage sludge in combination with other combustible waste materials. e

See Section 10.4.6 for a discussion of dioxin-like PCB releases to air from sewage sludge incinerators. f

The Ocean Dumping Ban Act of 1988 generally prohibited the dumping of sewage sludge into the ocean after 

 December 31, 1991. Ocean dumping of sewage sludge ended in June 1992 (Federal Register, 1993a). The

 current method of disposal of the 336,000 metric tons of sewage sludge that were disposed of in the oceans in

 1988 has not been determined. 

Sources: Federal Register (1990, 1993a); Green et al. (1995); Cramer et al. (1995). 

10-23
 

Table 10-11. Quantity of sewage sludge disposed of annually in 1995 by 

primary, secondary, or advanced treatment publicly owned treatment works 

(POTWs) and potential dioxin-like PCB TEQ releases 

Use/disposal practice 

Volume disposed of 

(1,000 dry metric 

tons/yr) 

Percent of 

total volume 

Potential dioxin releasea 

(g TEQ/yr) 

TEQP-WHO98 TEQP-WHO94 

Land applicationb 2,500 39.7 60.5 62.8 

Advanced treatmentc 700 11.1 16.9 17.6 

Other beneficial used 500 7.9 12.1 12.6 

Surface disposal/Landfill 1,100 17.5 26.6 27.6 

Incineration 1,400 22.2 e e 

Other disposal method 100 1.6 2.4 2.5 

TOTAL 6,300 100 118.5 123.1 

a

Potential TEQ release for nonincinerated sludges was estimated by multiplying the sludge volume generated

 (column 2) by the mean dioxin-like PCB TEQ concentration in 74 POTW sludges reported by Green et al. (1995)

 and Cramer et al. (1995) (i.e., 24.2 ng TEQP- WHO98/kg and 25.1 ng TEQP-WHO94/kg).
 
 b

Without further processing or stabilization, such as composting.
 
 c

Such as composting.
 
 d

EPA assumed that this category includes distribution and marketing (i.e., sale or give-away of sludge for use in

 home gardens). Based on the 1988 National Sewage Sludge Survey and 1988 Needs Survey, approximately 1.3%

 of the total volume of sewage disposed was distributed and marketed (Federal Register, 1993a). Therefore, it is

 estimated that 2 g (TEQP-WHO98 and TEQP-WHO94) were released through distribution and marketing in 1995.
 
 e

See Section 3.5 for estimates of CDD/CDF releases to air from sewage sludge incinerators. 

Sources: U.S. EPA (1999b); Green et al. (1995); Cramer et al. (1995). 

These tables also list the estimated amount of dioxin-like PCB TEQs that may be present 

in sewage sludge and potentially released to the environment. For reference years 1987 and 

1995, these values were estimated using the POTW mean TEQP-WHO98 concentration calculated 

from the results reported by Green et al. (1995) and Cramer et al. (1995). For reference year 

2000, they were estimated using the POTW mean TEQP-WHO98 concentration reported by EPA 

(U.S. EPA, 2002d) as part of the 2001 National Sewage Sludge Survey. Multiplying these TEQ 

concentrations by the sludge volumes generated yields annual potential total releases of 101 g 

TEQP-WHO98 (104.2 g TEQP-WHO94) in 1987, 118.5 g TEQP-WHO98 (123.1 g TEQP-WHO94) in 

1995, and 26.6 g TEQP-WHO98 in 2000 for nonincinerated sludges. 

10-24
 

Table 10-12. Quantity of sewage sludge disposed of annually in 2000 by 

primary, secondary, or advanced treatment publicly owned treatment works 

(POTWs) and potential dioxin-like PCB TEQ releases 

Use/disposal practice 

Volume disposed of 

(1,000 dry metric 

tons/yr) 

Percent of 

total volume 

Potential 

TEQDF-WHO98 releasea 

(g TEQ/yr) 

Land applicationb 2,800 42.4 14.6 

Advanced treatmentc 800 12.1 4.2 

Other beneficial used 500 7.6 2.6 

Surface disposal/landfill 900 13.6 4.7 

Incineration 1,500 22.7 e 

Other disposal method 100 1.5 0.5 

TOTAL 6,600 100 26.6 

a

Potential dioxin TEQ release for nonincinerated sludges was estimated by multiplying the sludge volume

 generated (column 2) by the average of the mean TEQDF-WHO98 concentrations in sludge reported by U.S. EPA

 (2002c) (i.e., 5.22 ng TEQDF-WHO98/kg).
 
 b

Without further processing or stabilization, such as composting.
 
 c

Such as composting.
 
 d

EPA assumed that this category includes distribution and marketing (i.e., sale or give-away of sludge for use in

 home gardens). Based on the 1988 National Sewage Sludge Survey and 1988 Needs Survey, approximately 1.3%

 of the total volume of sewage disposed of was distributed and marketed (Federal Register, 1993a). Therefore, it

 is estimated that 0.5 g TEQDF-WHO98 were released through distribution and marketing in 2000.
 
 e

See Section 3.5 for estimates of CDD/CDF releases to air from sewage sludge incinerators. 

Sources: U.S. EPA (1999b, 2002d). 

Of the 101 g TEQP-WHO98 released in 1987, 1.7 g entered commerce as a product for 

distribution and marketing and the remainder was applied to land (41.5 g to land application and 

9.6 g to surface disposal sites) or landfilled (48.2 g). Of the 118.5 g TEQP-WHO98 released in 

1995, 60.5 g were applied to land without further processing or stabilization, 16.9 g underwent 

advanced treatment such as composting, 26.6 g were disposed of on the surface or landfilled, and 

the remainder was either used or disposed of in other ways. Of the 26.6 g TEQP-WHO98 released 

in 2000, 14.6 g were applied to land without further processing or stabilization, 4.2 g underwent 

advanced treatment such as composting, 4.7 g were disposed of on the surface or landfilled, and 

the remainder was either used or disposed of in other ways. The PCBs in landfilled sludge were 

not considered releases to the environment under the definition established in this document. 

The other disposal practices were considered releases and were summed to get total land releases, 

as shown in Table 10-2 (above). 

10-25
 
The 1987 and 1995 release estimates are assigned a confidence rating of B, indicating 

high confidence in the production estimate and medium confidence in the emission factor 

estimates. The medium rating was based on the judgment that, although the 74 facilities tested 

by Green et al. (1995) and Cramer et al. (1995) may be reasonably representative of the 

variability in POTW technologies and sewage characteristics nationwide, the sample size was 

still relatively small, and not all dioxin-like PCB congeners were monitored. The 2000 release 

estimates are assigned a confidence rating of A, indicating high confidence in both the 

production estimate and the emission factor estimates. High confidence was placed in the 

emission factors estimated because they were weighted on the basis of wastewater flow rates of 

all POTWs in the United States. 

10.3. CHEMICAL MANUFACTURING AND PROCESSING SOURCES 

In the early 1980s, EPA investigated the extent of inadvertent generation of PCBs during 

the manufacture of synthetic organic chemicals (Hammerstrom et al., 1985). For example, 

phthalocyanine dyes and diarylide pigments were reported to contain PCBs in the milligram-perkilogram range. EPA subsequently issued regulations under TSCA (40 CFR 761.3) that ban the 

distribution in commerce of any products containing an annual average PCB concentration of 25 

mg/kg (50 mg/kg maximum concentration at any time). In addition, EPA requires manufacturers 

with processes that inadvertently generate PCBs and importers of products that contain 

inadvertently generated PCBs to report to EPA any process or import for which the PCB 

concentration is greater than 2 mg/kg for any resolvable PCB gas chromatographic peak. 

10.4. COMBUSTION SOURCES 

10.4.1. Municipal Waste Combustors 

Municipal waste combustors (MWCs) have long been identified as potential PCB air 

emission sources. Stack gas concentrations of PCBs for three MWCs were reported (U.S. EPA, 

1987d); the average test results yielded an emission factor of 18 μg/kg refuse. Stack gas 

emissions of PCBs from the three MWCs were quantified without determining the MWCs’ PCB 

destruction efficiency. 

EPA also analyzed the PCB content of various consumer paper products (U.S. EPA, 

1987d). The results indicated that paper products such as magazine covers and paper towels 

contained up to 139 μg/kg paper. These levels, which were reported in 1981, were attributed to 

the repeated recycling of waste paper containing PCBs. For example, carbonless copy paper 

manufactured prior to 1971 contained PCB levels as high as 7%. This copy paper then became a 

component of waste paper, which was recycled. The PCBs were inevitably introduced into other 

paper products, resulting in continued measurable levels in municipal refuse some four years 

10-26
 

after the PCB manufacturing ban was imposed. Refuse-derived fuel manufactured from these 

paper products had PCB levels of 8,500 μg/kg, indicating that this fuel could be a source of 

atmospheric PCBs. Therefore, it was assumed that municipal refuse does contain detectable 

levels of PCBs and that some of these PCBs may enter the atmosphere when the refuse is 

incinerated (U.S. EPA, 1987d). 

Shane et al. (1990) analyzed fly ash from five MWCs for PCB congener group content. 

Total PCB levels ranged from 99 to 322 μg/kg in the ash, with the tri, tetra, and penta congener 

groups occurring in the highest concentrations. The investigators also analyzed seven bottom ash 

and eight bottom ash/fly ash mixtures for total PCB measured as Aroclor 1254. The DL for this 

Aroclor analysis was 5 μg/kg. Aroclor 1254 was detected in two of the seven bottom ash 

samples (26 and 8 μg/kg) and in five of the eight fly ash/bottom ash mixtures (range, 6 to 33 

μg/kg). 

Sakai et al. (2001) analyzed the PCB levels in fly ash and bottom ash from a newly 

constructed MWC in Japan. The I-TEQ values derived from the data give a total TEQ value of 

31.6 ng/kg for fly ash and 0.85 ng/kg for bottom ash. 

The development of more sensitive analytical methodologies has enabled researchers in 

recent years to detect dioxin-like PCB congeners in the stack gases and fly ash from full-scale 

and pilot-scale MWCs (Sakai et al., 1993a, b, 1994, 2001; Boers et al., 1993; Schoonenboom et 

al., 1993). Similarly, the advances in analytical techniques have enabled researchers to determine 

that dioxin-like PCBs can be formed during the oxidative solid combustion phase of incineration, 

presumably due to dimerization of chlorobenzenes. Laboratory-scale studies have also 

demonstrated that dioxin-like PCBs can be formed from heat treatment of fly ash in air 

(Schoonenboom et al., 1993; Sakai et al., 1994); however, the available data are not adequate to 

support development of a quantitative estimate of a dioxin-like PCB emission factor for this 

source category. MWCs are designated as a Category E source. 

10.4.2. Industrial Wood Combustion 

Emissions of PCB congener groups (but not individual congeners) were measured during 

stack testing at two industrial wood-burning facilities (CARB, 1990d, e). Table 10-13 presents 

the average of the congener group (monochlorobiphenyl [MCB] through decachlorobiphenyl 

[DCB]) emission factors for these two facilities. No tetra- or higher-chlorinated congeners (the 

congener groups containing the dioxin-like PCBs) were detected at either facility at DLs 

corresponding to emission factors in the low range of nanogram per kilogram of wood 

combusted. 

10-27
 

Table 10-13. PCB congener group emission factors for industrial wood 

combustorsa 

Congener group 

Number of 

detections 

Maximum 

concentration 

detected 

(ng/kg wood) 

Mean concentration 

(ng/kg) 

Nondetect set 

to detection 

limit 

Nondetect set 

to zero 

Monochlorobiphenyls 1 32.1 39.4 16 

Dichlorobiphenyls 1 23 50.9 11.5 

Trichlorobiphenyls 1 19.7 42.3 9.8 

Tetrachlorobiphenyls 0 -­ 22.7 --

Pentachlorobiphenyls 0 -­ 17.6 --

Hexachlorobiphenyls 0 -­ 17 --

Heptachlorobiphenyls 0 -­ 17.9 --

Octachlorobiphenyls 0 -­ 15.8 --

Nonachlorobiphenyls 0 -­ 25 --

Decachlorobiphenyls 0 -­ 36.3 -­ a

Two sites for each congener group. 

-- = No information given 

Source: CARB (1990d, e). 

In CARB (1990d), PCBs were measured in the emissions from two spreader stoker woodfired boilers operated in parallel by an electric utility for generating electricity. The exhaust gas 

stream from each boiler was passed through a dedicated electrostatic precipitator (ESP), after 

which the gas streams were combined and emitted to the atmosphere through a common stack. 

Stack tests were conducted both when the facility burned fuels allowed by existing permits and 

when the facility burned a mixture of permitted fuel supplemented by urban wood waste at a ratio 

of 70:30. 

In CARB (1990e), PCBs were measured in the emissions from twin fluidized-bed 

combustors designed to burn wood chips to generate electricity. The air pollution control device 

(APCD) system consisted of ammonia injection for controlling nitrogen oxides and a multiclone 

and an ESP for controlling PM. During testing, the facility burned wood wastes and agricultural 

wastes allowed by existing permits. The available data are not adequate to support development 

of a quantitative estimate of dioxin-like PCB emissions from this source. Industrial wood 

combustion is designated as a Category E source. 

10-28
 
10.4.3. Medical Waste Incineration 

As discussed in Chapter 3, Section 3.3, EPA has issued nationally applicable emission 

standards and guidelines that address CDD/CDF emissions from medical waste incinerators 

(MWIs). Although PCBs are not addressed in these regulations, the database of stack test results 

at MWIs compiled for this rulemaking does contain limited data on PCB congener group 

emission factors. Data are available for two MWIs lacking add-on APCD equipment and for two 

MWIs with add-on APCD equipment in place. The average congener group emission factors 

derived from these test data are presented in Table 10-14. Because data are available for only 4 

of the estimated 1,065 facilities that make up this industry, and because these data do not provide 

congener-specific emission factors, no national estimates of total PCB or dioxin-like PCB 

emissions are being made at this time. Medical waste incineration is designated as a Category E 

source. 

Table 10-14. PCB congener group emission factors for medical waste 

incinerators (MWIs)a 

Congener group 

Mean emission factor (ng/kg) 

(2 MWIs without APCD)

 Mean emission factor (ng/kg) 

(2 MWIs with APCD) 

Nondetects set to 

detection limit 

Nondetects 

set to zero 

Nondetects set to 

detection limit 

Nondetects 

set to zero 

Monochlorobiphenyls 0.059 0.059 0.311 0 

Dichlorobiphenyls 0.083 0.083 0.34 0 

Trichlorobiphenyls 0.155 0.155 0.348 0 

Tetrachlorobiphenyls 4.377 4.377 1.171 0 

Pentachlorobiphenyls 2.938 2.938 17.096 9.996 

Hexachlorobiphenyls 0.238 0.238 1.286 1.078 

Heptachlorobiphenyls 0.155 0.155 0.902 0 

Octachlorobiphenyls 0.238 0.238 0.205 0 

Nonachlorobiphenyls 0.155 0.155 -­ --

Decachlorobiphenyls 0.155 0.155 0.117 0 

a

See Section 3.3 for details on tested facilities. 

APCD = Air pollution control device 

-- = No information given 

10-29
 

10.4.4. Tire Combustion 

As discussed in Chapter 3, Section 3.6, tires are burned in a variety of facilities, including 

dedicated tire burners, cement kilns, industrial boilers, and pulp and paper combustion facilities. 

Emissions of PCB congener groups (but not individual congeners) were measured during stack 

testing of a tire incinerator (CARB, 1991). The facility consisted of two excess air furnaces 

equipped with steam boilers to recover the energy from the heat of combustion. Discarded whole 

tires were fed to the incineration units at rates ranging from 2,800 to 5,700 kg/hr during the three 

testing days. The furnaces were equipped to burn natural gas as auxiliary fuel. The steam 

produced from the boilers drove electrical turbine generators that produced 14.4 megawatts of 

electricity. The facility was equipped with a dry acid gas scrubber and a fabric filter for the 

control of emissions prior to exiting the stack. Table 10-15 presents the congener group (MCB 

through DCB) emission factors for this facility. The emission factor for the total of the tetrathrough heptachlorinated congener groups was about 1.2 μg/kg tire processed. Because these 

data do not provide PCB congener-specific emission factors, no estimates of emissions of dioxinlike PCBs can be made. Tire combustion is designated as a Category E source. 

Table 10-15. PCB congener group emission factors for a tire combustora 

Congener group 

Number of 

detections 

Maximum 

emission factor 

(ng/kg) 

Mean emission factor 

(ng/kg) 

Nondetect set to 

detection limit 

Nondetect set 

to zero 

Monochlorobiphenyls 0 -­ 0.04 --

Dichlorobiphenyls 1 34.8 11.7 11.6 

Trichlorobiphenyls 1 29.5 11.8 9.8 

Tetrachlorobiphenyls 0 -­ 10 --

Pentachlorobiphenyls 2 2,724 1,092 1,092 

Hexachlorobiphenyls 1 106.5 55.9 35.5 

Heptachlorobiphenyls 1 298.6 107.7 99.5 

Octachlorobiphenyls 0 -­ 20.9 --

Nonachlorobiphenyls 0 -­ 17.7 --

Decachlorobiphenyls 0 -­ 41.9 -­ a

Three samples for each congener group. 

-- = No information given 

Source: CARB (1991). 

10-30 
10.4.5. Cigarette Smoking 

Using high-resolution mass spectrometry, Matsueda et al. (1994) analyzed tobacco from 

20 brands of commercially available cigarettes collected in 1992 from Japan, the United States, 

Taiwan, China, the United Kingdom, Germany, and Denmark for the PCB congeners 77, 126, 

and 169. Table 10-16 presents the results of the study. However, no studies examining tobacco 

smoke for the presence of these congeners have been reported. Thus, it is not known whether the 

PCBs present in the tobacco are destroyed or volatilized during combustion or whether PCBs are 

formed during combustion. At least 1,200 tobacco constituents (e.g., nicotine, n-paraffin, some 

terpenes) are transferred intact from the tobacco into the smoke stream by distillation in this area, 

and it is plausible that PCBs present in the unburned tobacco would be subject to similar 

distillation. 

Cigarette consumption and the combustion processes operating during cigarette smoking 

are discussed in Chapter 5, Section 5.5. 

A preliminary rough estimate of potential emissions of dioxin-like PCBs can be made 

using the following assumptions: (a) the average TEQP-WHO98 content of seven brands of U.S. 

cigarettes reported by Matsueda et al. (1994), 0.64 pg/pack (0.032 pg/cigarette), is representative 

of cigarettes smoked in the United States; (b) dioxin-like PCBs are neither formed nor destroyed, 

and the congener profile reported by Matsueda et al. (1994) is not altered during combustion of 

cigarettes; and (c) all dioxin-like PCBs contributing to the TEQ are released from the tobacco 

during smoking. On the basis of these assumptions, the calculated annual emissions would be 

0.018 g TEQP-WHO98, 0.016 g TEQP-WHO98, and 0.014 g TEQP-WHO98 for reference years 

1987, 1995, and 2000, respectively. These estimates are assigned a confidence rating of D 

because the emission factor is clearly not representative of cigarette smoke. 

10.4.6. Sewage Sludge Incineration 

EPA (U.S. EPA, 1996g) derived an emission factor of 5.4 μg PCBs/kg dry sewage sludge 

incinerated. This emission factor was based on measurements conducted at five multiple-hearth 

incinerators controlled with wet scrubbers. However, it is not known what fraction of the 

emissions was dioxin-like PCBs. 

In 1999, stack tests were conducted at a multiple-hearth incinerator in Ohio equipped 

with a venturi scrubber and a three-tray impingement conditioning tower (U.S. EPA, 2000b). Of 

the four test runs conducted, the first test run was aborted and the results from the fourth test run 

were determined to be statistical outliers (p>0.05). The back-half CDD/CDF concentrations for 

test run 4 were 50 to 60% lower than back-half emission concentrations for test runs 2 and 3. 

Because of the problems associated with test run 4, the results were not used to calculate an 

emission factor for dioxin-like PCBs. The average TEQ emission factor (excluding test run 4) 

10-31
 

Table 10-16. Dioxin-like PCB concentrations in cigarette tobacco in brands from various countri

10-32
 


Congener 

IUPAC 

number 

U.S. 

(avg. of 7 

brands) 

Japan 

(avg. of 6 

brands) 

United 

Kingdom 

(avg. of 3 

brands) 

Taiwan 

(1 brand) 

China 

(1 brand) 

Den

(1 b

3,3',4,4'-TCB 77 105.7 70.2 53 133.9 12.6 

3,4,4',5-TCB 81 2,3,3',4,4'-PeCB 105 2,3,4,4',5-PeCB 114 2,3',4,4',5-PeCB 118 2',3,4,4',5-PeCB 123 3,3',4,4',5-PeCB 126 6.2 7.8 6.1 14.5 2.4 

2,3,3',4,4',5-HxCB 156 2,3,3',4,4',5'-HxCB 157 2,3',4,4',5,5'-HxCB 167 3,3',4,4',5,5'-HxCB 169 0.9 0.9 0.9 2.4 0.4 

2,2',3,3',4,4',5-HpCB 170 2,2',3,4,4',5,5'-HpCB 180 2,3,3',4,4',5,5'-HpCB 189 Total TEQP-WHO98 0.64 0.8 0.62 1.49 0.24 

aBlank cells indicate that no measurements of these congeners were made. 

Source: Matsueda et al. (1994). 
was 0.51 ng TEQP-WHO98/kg (see Table 10-17). This emission factor was assigned a low 

confidence rating because it is based on limited surveys that are judged to be possibly 

nonrepresentative. 

Table 10-17. Dioxin-like PCB concentrations in stack gas collected from a 

U.S. sewage sludge incinerator 

Congener 

IUPAC 

number 

Mean emission factor (ng/kg) 

Nondetect set to 1⁄2 

detection limit Nondetect set to zero 

3,3',4,4'-TCB 77 92.37 92.37 

2,3,3',4,4'-PeCB 105 18 18 

2,3,4,4',5-PeCB 114 2.56 2.56 

2',3,4,4',5-PeCB 123 0.82 0.82 

2,3',4,4',5-PeCB 118 38.65 38.65 

3,3',4,4',5-PeCB 126 4.51 4.51 

2,3,3',4,4',5-HxCB 156 4.25 4.25 

2,3,3',4,4',5'-HxCB 157 1.41 1.41 

2,3',4,4',5,5'-HxCB 167 2.55 2.55 

3,3',4,4',5,5'-HxCB 169 3.61 3.61 

2,2',3,3',4,4',5-HpCB 170 7.19 7.19 

2,2',3,4,4',5,5'-HpCB 180 17.79 17.79 

2,3,3',4,4',5,5'-HpCB 189 0.6 0.6 

Total TEQP-WHO98 0.51 0.51 

Source: U.S. EPA (2000b). 

Approximately 0.865 million dry metric tons of sewage sludge were incinerated in 1988 

(Federal Register, 1993a), approximately 2.11 million dry metric tons in 1995 (e-mail dated July 

13, 1998, from K. Maw, Pacific Environmental Services, to G. Schweer, Versar, Inc.), and an 

estimated 1.42 million dry metric tons in 2000 (U.S. EPA, 1999b). Using the above estimated 

amounts of sewage sludge incinerated per year and the average TEQ emission factor of 0.51 ng 

TEQP-WHO98/kg, the estimated annual releases of total PCBs to air were 0.44 g TEQP-WHO98 in 

1987, 1.1 g TEQP-WHO98 in 1995, and 0.72 g TEQP-WHO98 in 2000. These emissions were 

assigned a low confidence rating (Category C) because the emission factor was given a low 

rating. 

10-33
 
10.4.7. Backyard Barrel Burning 

The low combustion temperatures and oxygen-starved conditions associated with 

backyard barrel burning may result in incomplete combustion and increased pollutant emissions 

(Lemieux, 1997). EPA’s Control Technology Center, in cooperation with New York State’s 

departments of health and environmental conservation, conducted a study to examine, 

characterize, and quantify emissions from the simulated open burning of household waste 

materials in barrels (Lemieux, 1997). A representative waste to be burned was prepared on the 

basis of the typical percentages of various waste materials disposed of by New York State 

residents (i.e., nonavid recyclers); hazardous wastes such as chemicals, paints, and oils were not 

included in the test waste. A variety of compounds, including dioxin-like PCBs, were measured 

in the emissions from the simulated open burning. The measured TEQ emission factors for 

waste that had not been separated for recycling purposes were 1.02 × 10-2 μg TEQP-WHO94/kg 

and 5.26 × 10-3 μg TEQP-WHO98/kg waste burned (see Table 10-18). These limited emissions 

data were judged to be inadequate for estimating national emissions (a Category E source). The 

activity level for backyard barrel burning is discussed in Section 6.5.2. 

10.4.8. Petroleum Refining Catalyst Regeneration 

As discussed in Section 5.4, regeneration of spent catalyst used in catalytic reforming to 

produce high-octane reformates is a potential source of CDD/CDF air emissions. In 1998, 

emissions from the caustic scrubber used to treat gases from the external catalyst regeneration 

unit of a refinery in California were tested for CDDs/CDFs as well as PCB congener groups 

(CARB, 1999) (see Chapter 5, Section 5.4 for details). 

All PCB congener groups were detected in each of the three samples collected. The 

average congener group emission factors in units of nanograms per barrel of reformer feed are 

presented in Table 10-19. The total PCB emission factor was 118 ng/barrel. This emission 

factor assumes that emissions are proportional to reforming capacity; emission factors may be 

more related to the amount of coke burned, APCD equipment present, and/or other process 

parameters. 

Because emissions data are available for only one U.S. petroleum refinery (which 

represents less than 1% of the catalytic reforming capacity at U.S. refineries), and because these 

data do not provide congener-specific emission factors, no national estimates of total PCB or 

dioxin-like PCB emissions are being made at this time. This is a Category E source. 

10-34
 

Table 10-18. Dioxin-like PCB emission factors from backyard barrel 

burninga 

Congener 

IUPAC 

number 

Emission factors (:g/kg) 

Test 1 Test 2 Average 

3,3',4,4'-TCB 77 9.3 15.2 12.3 

3,4,4',5-TCB 81 

2,3,3',4,4'-PeCB 105 5.9 4.9 5.4 

2,3,4,4',5-PeCB 114 

2,3',4,4',5-PeCB 118 8.3 14.3 11.3 

2',3,4,4',5-PeCB 123 18.6 28.7 23.7 

3,3',4,4',5-PeCB 126 

2,3,3',4,4',5-HxCB 156 

2,3,3',4,4',5'-HxCB 157 

2,3',4,4',5,5'-HxCB 167 

3,3',4,4',5,5'-HxCB 169 

2,2',3,3',4,4',5-HpCB 170 

2,2',3,4,4',5,5'-HpCB 180 

2,3,3',4,4',5,5'-HpCB 189 

Total TEQP-WHO98 4.21e!03 6.31e!03 5.26e!03 

a

Blank cells indicate that the congener was not detected in either of the two duplicate samples. 

Source: Lemieux (1997). 

Table 10-19. PCB congener group emission factors for a petroleum catalytic

reforming unita 

Congener group 

Mean 

concentration 

(ng/dscm) 

(at 12% O2) 

Mean emission 

rate 

(lb/hr) 

Mean emission 

factor 

(lb/1000 bbl) 

Mean emission 

factor 

(ng/barrel) 

Monochlorobiphenyls 166 5.51e!08 7.11e!09 3.23e+00 

Dichlorobiphenyls 355 1.17e!07 1.52e!08 6.89e+00 

Trichlorobiphenyls 743 2.45e!07 3.17e!08 1.44e+01 

Tetrachlorobiphenyls 849 2.81e!07 3.62e!08 1.64e+01 

Pentachlorobiphenyls 914 3.02e!07 3.88e!08 1.76e+01 

Hexachlorobiphenyls 780 2.57e!07 3.30e!08 1.50e+01 

Heptachlorobiphenyls 1,430 4.73e!07 6.01e!08 2.73e+01 

Octachlorobiphenyls 698 2.32e!07 2.95e!08 1.34e+01 

Nonachlorobiphenyls 179 5.99e!08 7.59e!09 3.44e+00 

Decachlorobiphenyls 41.3 1.39e!08 1.76e!09 7.98e!01 

Total PCBs 6,155.3 2.04e!06 2.61e!07 1.18e+02 

a

Three samples and three detections for each congener group. 

Source: CARB (1999). 

10-35 
10.5. NATURAL SOURCES 

This section discusses biotransformation and photochemical transformation of other 

PCBs. Although there is some evidence that these processes occur, the data were considered 

insufficient for developing release estimates. 

10.5.1. Biotransformation of Other PCBs 

Studies show that under anaerobic conditions, biologically mediated reductive 

dechlorination to lower-chlorinated congeners, followed by slow anaerobic and/or aerobic 

biodegradation, is a major pathway for destruction of PCBs in the environment. Research 

reported to date and summarized below indicates that biodegradation should result in a net 

decrease rather than a net increase in the environmental load of dioxin-like PCBs. 

Laboratory studies (e.g., Bedard et al., 1986; Pardue et al., 1988; Larsson and 

Lemkemeier, 1989; Hickey, 1995; Schreiner et al., 1995) have revealed that more than two dozen 

strains of aerobic bacteria and fungi that are capable of degrading most PCB congeners with five 

or fewer chlorines are widely distributed in the environment. Many of these organisms are of the 

genus Pseudomonas or Alcaligenes. The major metabolic pathway involves addition of oxygen 

at the 2,3-position by a dioxygenase enzyme, with subsequent dehydrogenation to the catechol 

followed by ring cleavage. Several bacterial strains have been shown to possess a dioxygenase 

enzyme that attacks the 3,4-position. 

Only a few strains have demonstrated the ability to degrade hexachlorobiphenyl (HxCB) 

and the higher-chlorinated biphenyls. The rate of aerobic biodegradation decreases with 

increasing chlorination. The half-lives for biodegradation of tetrachlorobiphenyls (TCBs) in 

fresh surface water and soil are 7 to 60+ days and 12 to 30 days, respectively. For 

pentachlorobiphenyls (PeCBs) and the higher-chlorinated PCBs, the half-lives in fresh surface 

water and soil are likely to exceed 1 year. PCBs with all or most chlorines on one ring and PCBs 

with fewer than two chlorines in the ortho position tend to degrade more rapidly. For example, 

Gan and Berthouex (1994) monitored over a 5-yr period the disappearance of PCB congeners 

applied to soil with sewage sludge. Three of the tetra- and pentachlorinated dioxin-like PCBs 

(IUPAC Nos. 77, 105, and 118) followed a first-order disappearance model, with half-lives 

ranging from 43 to 69 months. A hexa-substituted congener (IUPAC No. 167) and a heptasubstituted congener (IUPAC No. 180) showed no significant loss over the 5-yr period. 

Prior to the early 1990s, little investigation focused on anaerobic microbial dechlorination 

or degradation of PCBs, even though most PCBs eventually accumulate in anaerobic sediments 

(Abramowicz, 1990; Risatti, 1992). Environmental dechlorination of PCBs via losses of meta 

and para chlorines has been reported in field studies for freshwater, estuarine, and marine 

anaerobic sediments, including those from the Acushnet Estuary, the Hudson River, the 

10-36
 

Sheboygan River, New Bedford Harbor, Escambia Bay, Waukegan Harbor, the Housatonic 

River, and Woods Pond (Brown et al., 1987; Rhee et al., 1989; Van Dort and Bedard, 1991; 

Abramowicz, 1990; Bedard et al., 1995; Bedard and May, 1996). The altered PCB congener 

distribution patterns found in these sediments (i.e., different patterns with increasing depth or 

distance from known sources of PCBs) have been interpreted as evidence that bacteria may 

dechlorinate PCBs in anaerobic sediment. 

Reported results of laboratory studies confirm anaerobic degradation of PCBs. Chen et 

al. (1988) found that “PCB-degrading” bacteria from the Hudson River could significantly 

degrade the MCB, dichlorobipheyl (DiCB), and trichlorobiphenyl (TrCB) components of a 20 

ppm Aroclor 1221 solution within 105 days. These congener groups make up 95% of Aroclor 

1221. No degradation of higher-chlorinated congeners (present at 30 ppb or less) was observed, 

and a separate 40-day experiment with TCB also showed no degradation. 

Rhee et al. (1989) reported degradation of mono- to penta-substituted PCBs in 

contaminated Hudson River sediments held under anaerobic conditions in the laboratory (N2 

atmosphere) for 6 months at 25°C. Amendment of the test samples with biphenyl resulted in 

greater loss of PCBs. No significant decreases in the concentrations of the higher-chlorinated 

congeners (more than five chlorines) were observed. No evidence of degradation was observed 

in samples incubated in CO2/H2 atmospheres. Abramowicz (1990) hypothesized that this result 

could be an indication that, in the absence of CO2, a selection is imposed favoring organisms 

capable of degrading PCBs to obtain CO2 and/or low-molecular-weight metabolites as electron 

receptors. 

Risatti (1992) examined the degradation of PCBs at varying concentrations (10,000 ppm, 

1,500 ppm, and 500 ppm) in the laboratory with “PCB-degrading” bacteria from Waukegan 

Harbor. After nine months of incubation at 22°C, the 500 ppm and 1,500 ppm samples showed 

no change in PCB congener distributions or concentrations, thus indicating a lack of degradation. 

Significant degradation was observed in the 10,000 ppm sediment, with at least 20 congeners 

ranging from the TrCBs through the PeCBs showing decreases. 

Quensen et al. (1988) also demonstrated that microorganisms from PCB-contaminated 

sediments (Hudson River) dechlorinated most TrCBs through HxCBs in Aroclor 1242 under 

anaerobic laboratory conditions. The Aroclor 1242 used to spike the sediment contained 

predominantly TrCBs and TCBs (85 mol percent). Three concentrations of the Aroclor, 

corresponding to 14, 140, and 700 ppm on a sediment dry-weight basis, were used. 

Dechlorination was most extensive at the 700 ppm test concentration: 53% of the total chlorine 

was removed in 16 weeks, and the proportion of TCBs through HxCBs decreased from 42 to 4%. 

Much less degradation was observed in the 140 ppm sediment, and no observable degradation 

was found in the 14ppm sediment. These results and those of Risatti (1992) suggest that the 

10-37
 

organism(s) responsible for this dechlorination may require relatively high levels of PCBs as a 

terminal electron acceptor to maintain a growing population. 

Quensen et al. (1990) reported that dechlorination of 500 ppm spike concentrations of 

Aroclor 1242, 1248, 1254, and 1260 by microorganisms from PCB-contaminated sediments in 

the Hudson River and Silver Lake occurred primarily at the meta and para positions; orthosubstituted MCBs and DiCBs increased in concentration. Significant decreases over the 

incubation period (up to 50 weeks) were reported for dioxin-like PCBs 156, 167, 170, 180, and 

189. Of the four dioxin-like TCBs and PeCBs detected in the Aroclor spikes (IUPAC Nos. 77, 

105, 114, and 118), all decreased significantly in concentration, with the possible exception of 

PeCB 114 in the Aroclor 1260-spiked sediment. 

Nies and Vogel (1990) reported similar results with Hudson River sediments incubated 

anaerobically and enriched with acetone, methanol, or glucose. Approximately 300 ppm of 

Aroclor 1242 (31 mol percent TCBs, 7 mol percent PeCBs, and 1 mol percent HxCBs) were 

added to the sediments prior to incubation for 22 weeks under an N2 atmosphere. Significant 

dechlorination was observed, primarily at the meta and para positions on the higher-chlorinated 

congeners (TCBs, PeCBs, and HxCBs), resulting in the accumulation of lower-chlorinated, 

primarily ortho-substituted mono- through tri-substituted congeners. No significant 

dechlorination was observed in the control samples (samples containing no added organic 

chemical substrate and samples that were autoclaved). 

Bedard and May (1996) also reported similar findings in the sediments of Woods Pond, 

which was believed to be contaminated with Aroclor 1260. Significant decreases in the sediment 

concentrations of PCBs 118, 156, 170, and 180 (relative to their concentrations in Aroclor 1260) 

were observed. No increases or decreases were reported for the other dioxin-like PCBs. 

Bedard et al. (1995) demonstrated that it is possible to stimulate substantial microbial 

dechlorination of the highly chlorinated PCB mixture Aroclor 1260 in situ with a single addition 

of 2,6-dibromobiphenyl. The investigators added 365 g of 2,6-dibromobiphenyl to 6-ft-diameter 

submerged caissons containing 400 kg sediment (dry weight) and monitored the change in PCB 

congener concentrations for a period of one year. At the end of the observation period, the 

MCBs through HxCBs decreased by 74% in the top of the sediment and by 69% in the bottom. 

The average number of chlorines per molecule dropped 21%, from 5.83 to 4.61, with the largest 

reduction observed in meta chlorines (54% reduction) followed by para chlorines (6%). The 

dechlorination stimulated by 2,6-dibromobiphenyl selectively removed meta-chlorines positioned 

next to other chlorines. 

The findings of these latter studies are significant, because removal of meta and para 

chlorines from the dioxin-like PCBs should reduce their toxicity and bioaccumulative potential 

10-38
 

and also lead to the formation of lower-chlorinated congeners that are more amenable to aerobic 

biodegradation. 

Van Dort and Bedard (1991) reported the first experimental demonstration of biologically 

mediated ortho-dechlorination of a PCB and stoichiometric conversion of that PCB congener 

(2,3,5,6-TCB) to lower-chlorinated forms. In that study, 2,3,5,6-TCB was incubated under 

anaerobic conditions with unacclimated methanogenic pond sediment for 37 weeks, with 

reported dechlorination to 2,5-DiCB (21%); 2,6-DiCB (63%); and 2,3,6-TrCB (16%). 

10.5.2. Photochemical Transformation of Other PCBs 

Photolysis and photo-oxidation may be major pathways for destruction of PCBs in the 

environment. Research reported to date and summarized below indicates that ortho-substituted 

chlorines are more susceptible to photolysis than are meta- and para-substituted congeners; thus, 

photolytic formation of more toxic dioxin-like PCBs may occur. Oxidation by hydroxyl radicals, 

however, apparently occurs preferentially at the meta and para positions, resulting in a net 

decrease rather than a net increase in the environmental load of dioxin-like PCBs. 

On the basis of the data available in 1983, Leifer et al. (1983) concluded that all PCBs, 

especially the higher-chlorinated congeners and those that contain two or more chlorines in the 

ortho position, photodechlorinate. In general, as the chlorine content increases, the photolysis 

rate increases. More recently, Lepine et al. (1992) exposed dilute solutions (4 ppm) of Aroclor 

1254 in cyclohexane to sunlight for 55 days in December and January. Congener-specific 

analysis indicated that the amounts of many higher-chlorinated congeners, particularly monoortho-substituted congeners, decreased, whereas those of some lower-chlorinated congeners 

increased. The results for the dioxin-like PCBs indicated a 43.5% decrease in the amount of 

PeCB 114, a 73.5% decrease in the amount of HxCB 156, and a 24.4% decrease in the amount of 

HxCB 157. However, TCB 77 and PeCB 126 (the most toxic of the dioxin-like PCB congeners), 

which were not detected in unirradiated Aroclor 1254, represented 2.5% and 0.43%, respectively, 

of the irradiated mixture. 

With regard to photo-oxidation, Atkinson (1987) and Leifer et al. (1983), using assumed 

steady-state atmospheric OH concentrations and measured oxidation rate constants for biphenyl 

and MCB, estimated atmospheric decay rates and half-lives for gas-phase PCBs. Atmospheric 

transformation was estimated to proceed most rapidly for those PCB congeners containing either 

a small number of chlorines or those containing all or most of the chlorines on one ring. Kwok et 

al. (1995) extended the work of Atkinson (1987) by measuring the OH radical reaction rate 

constants for 2,2'-, 3,3'-, and 3,5-DiCB. These reaction rate constants, when taken together with 

Atkinson’s measurements for biphenyl and MCB and the estimation method described in 

Atkinson (1991), were used to generate more reliable estimates of the gas-phase OH radical 

10-39
 

reaction rate constants for the dioxin-like PCBs. The persistence of the PCB congeners increased 

with increasing degree of chlorination. Table 10-20 presents these estimated rate constants and 

the corresponding tropospheric lifetimes and half-lives. 

Table 10-20. Estimated tropospheric half-lives of dioxin-like PCBs with 

respect to gas-phase reaction with the OH radical 

Congener group 

Dioxin-like 

congener 

Estimated 

OH reaction 

rate constant 

(10-12

 cm3/ 

molecule-sec) 

Estimated 

tropospheric 

lifetime 

(days)a 

Estimated 

tropospheric 

half-life 

(days)a 

Tetrachlorobiphenyls 3,3',4,4'-TCB 

3,4,4',5-TCB 

0.583 

0.71 

20 

17 

14 

12 

Pentachlorophenyls 2,3,3',4,4'-PeCB 

2,3,4,4',5-PeCB 

2,3',4,4',5-PeCB 

2',3,4,4',5-PeCB 

3,3',4,4',5-PeCB 

0.299 

0.383 

0.299 

0.482 

0.395 

40 

31 

40 

25 

30 

28 

22 

28 

17 

21 

Hexachlorobiphenyls 2,3,3',4,4',5-HxCB 

2,3,3',4,4',5'-HxCB 

2,3',4,4',5,5'-HxCB 

3,3',4,4',5,5'-HxCB 

0.183 

0.214 

0.214 

0.266 

65 

56 

56 

45 

45 

39 

39 

31 

Heptachlorobiphenyls 2,2',3,3',4,4',5-HpCB 

2,2',3,4,4',5,5'-HpCB 

2,3,3',4,4',5,5'-HpCB 

0.099 

0.099 

0.125 

121 

121 

95 

84 

84 

66 a

Calculated using a 24-hr, seasonal, annual, and global tropospheric average OH radical concentration of

 9.7 × 105

 molecule/cm3 (Prinn et al., 1995). 

Source: Telephone conversation on November 16, 1995, between Roger Atkinson, Air Pollution Research Center, 

University of California, and Greg Schweer, Versar, Inc. (based on Atkinson, 1991, and Kwok et al., 1995). 

Sedlak and Andren (1991) demonstrated in laboratory studies that OH radicals generated 

with Fenton’s reagent rapidly oxidized PCBs (2-MCB and the DiCBs through PeCBs present in 

Aroclor 1242) in aqueous solutions. The results indicated that the reaction occurs via addition of 

a hydroxyl group to one nonhalogenated site; reaction rates are inversely related to the degree of 

chlorination of the biphenyl. The results also indicated that meta and para sites are more reactive 

than ortho sites due to stearic hindrance effects. On the basis of their kinetic measurements and 

reported steady-state aqueous system OH concentrations or estimates of OH radical production 

rates, the authors estimated environmental half-lives for dissolved PCBs (MCB through 

10-40
 
octachlorobiphenyl) in fresh surface water and in cloud water to be 4 to 11 days and 0.1 to 10 

days, respectively. 

10.6. PAST USE OF COMMERCIAL PCBs 

An estimated total of 1.5 million metric tons of PCBs were produced worldwide 

(DeVoogt and Brinkman, 1989). Slightly more than one-third of these PCBs (568,000 metric 

tons) were used in the United States (Versar, Inc., 1976). Although the focus of this section is on 

past uses of PCBs within the United States, it is necessary to note that the use and disposal of 

PCBs in many countries, coupled with the persistent nature of PCBs, have resulted in their 

movement and presence throughout the global environment. The ultimate sink of most PCBs 

released to the environment is aquatic sediments. Currently, however, large quantities of PCBs 

are estimated to be circulating between the air and water environments or are present in landfills 

and dumps, some of which may offer the potential for re-release of PCBs into the air. Tanabe 

(1988) presented a global mass balance for PCBs that indicated that as of 1985, 20% of the total 

PCBs produced were present in seawater, whereas only 11% were present in sediments (see 

Table 10-21). Nearly two-thirds of total global PCB production was estimated by Tanabe to still 

be in use in electrical equipment or to be present in landfills and dumps. 

As discussed in Section 10.2, an estimated 568,000 metric tons of PCBs were sold in the 

United States between 1930 and 1975 (Versar, Inc., 1976). Table 10-22 presents annual 

estimates of domestic sales by year for each Aroclor from 1957 to 1974. Estimates of PCB usage 

in the United States by usage category from 1930 to 1975 are presented in Table 10-23. Prior to 

voluntary restrictions by Monsanto Corporation in 1972 on sales for uses other than “closed 

electrical systems,” approximately 13% of the PCBs were used in “semi-closed applications,” 

and 26% were used in “open-end applications.” Most of the usage for semi-closed and open-end 

applications occurred between 1960 and 1972 (Versar, Inc., 1976). 

Table 10-24 presents estimates of the amounts of individual Aroclors that were directly 

released to the environment (water, air, or soil) between 1930 and 1974. Because detailed usage 

data were not available for the period 1930 to 1957, Versar, Inc. (1976) assumed that the usage 

pattern for this period followed the average pattern for the period of 1957 to 1959. The basic 

assumption used by Versar in deriving these estimates was that PCBs were released on the order 

of 5% of those used in closed electrical systems, 60% of those used in semi-closed applications, 

and 25% of those used for plasticizers and that 90% of PCBs used for miscellaneous industrial 

uses had escaped. The reliability of these release estimates was assumed to be ±30%. 

Versar, Inc. (1976) also estimated that 132,000 metric tons of PCBs were landfilled. This 

total comprised 50,000 metric tons from capacitor and transformer production wastes, 36,000 

metric tons from disposal of obsolete electrical equipment, and 46,000 metric tons from disposal 

10-41
 

Table 10-21. Estimated PCB loads in the global environment as of 1985 

Environment 

PCB load 

(metric tons) 

Percentage of 

PCB load 

Percentage of 

world production 

Terrestrial and coastal 

Air 

River and lake water 

Seawater 

Soil 

Sediment 

Biota 

Total 

500 

3,500 

2,400 

2,400 

130,000 

4,300 

143,100 

0.13 

0.94 

0.64 

0.64 

35 

1.1 

39 

Open ocean 

Air 

Seawater 

Sediment 

Biota 

Total 

790 

230,000 

110 

270 

231,170 

0.21 

61 

0.03 

0.07 

61 

Total load in environment 

Degraded and incinerated 

Land-stockeda 

World production 

374,000 

43,000 

783,000 

1,200,000b 

100 31 

4 

65 

100 a

Still in use in electrical equipment and other products, and deposited in landfills and dumps.
 
 b

This value is from Tanabe (1988). DeVoogt and Brinkman (1989) estimated worldwide production to have been

 1,500,000 metric tons. 

Source: Tanabe (1988). 

of material from open-end applications. An estimated additional 14,000 metric tons of PCBs, 

although still “in service” in various semi-closed and open-end applications in 1976, were 

ultimately destined for disposal in landfills. 

An estimated 3,702 kg of TEQP-WHO98 were released directly to the U.S. environment 

between 1930 and 1977 (see Table 10-25). These estimates are based on the Aroclor release 

estimates presented in Table 10-22 and the mean TEQP-WHO98 concentrations in Aroclors 

presented in Table 10-3. 

10-42
 

Table 10-22. Estimated domestic sales of aroclors and releases of PCBs, 1957–1974 (metric tons) 

10-43
 


Year 

Estimated domestic sales 

Aroclor 

1016 

Aroclor 

1221 

Aroclor 

1232 

Aroclor 

1242 

Aroclor 

1248 

Aroclor 

1254 

Aroclor 

1260 

Aroclor 

1262 

1957 0 10 89 8,265 807 2,023 3,441 14 

1958 0 7 51 4,737 1,161 3,035 2,713 83 1959 0 115 109 6,168 1,535 3,064 3,002 163 1960 0 47 70 8,254 1,282 2,761 3,325 148 1961 0 43 109 8,993 1,825 2,855 2,966 164 1962 0 64 102 9,368 1,571 2,869 2,991 196 1963 0 164 6 8,396 2,274 2,681 3,459 188 1964 0 270 6 10,692 2,376 2,849 3,871 202 

1965 0 167 3 14,303 2,524 3,509 2,645 253 

1966 0 239 7 17,943 2,275 3,191 2,665 348 

1967 0 200 11 19,529 2,134 3,037 2,911 381 

1968 0 62 41 20,345 2,220 4,033 2,382 327 

1969 0 230 124 20,634 2,563 4,455 2,013 323 

1970 0 670 118 22,039 1,847 5,634 2,218 464 

1971 1,512 1,005 78 9,970 97 2,114 782 0 

1972 9,481 78 0 330 366 1,585 138 0 

1973 10,673 16 0 2,812 0 3,618 0 0 

1974 9,959 26 0 2,815 0 2,805 0 0 

TOTAL 31,625 3,413 924 195,593 26,857 56,118 41,522 3,254 

% of Total 8.8 0.9 0.3 54.2 7.4 15.6 11.5 0.9 

Source: Versar, Inc. (1976). 
Table 10-23. Estimated U.S. usage of PCBs by use category, 1930–1975 

Use class Use category 

Amount used 

(1,000 metric 

tons) 

Percent of total 

usage 

Reliability of 

estimate (%) 

Closed electrical 

systems 

Capacitors 286 50.3 ±20 

Transformers 152 26.8 ±20 

Semi-closed 

applications 

Heat transfer 

fluids 

9 1.6 ±10 

Hydraulics and 

lubricants 

36 6.3 ±10 

Open-end 

applications 

Plasticizer uses 52 9.2 ±15 

Carbonless copy 

paper 

20 3.5 ±5 

Misc. industrial 12 2.1 ±15 

Petroleum 

additives 

1 <1 ±50 

TOTAL 568 100 

Source: Versar, Inc. (1976). 

10-44
 

Table 10-24. Estimated direct releases of Aroclors to the U.S. environment, 

1930–1974a (metric tons) 

Year 

Estimated environmental releases 

Total PCB 

releases 

Aroclor 

1016 

Aroclor 

1242 

Aroclor 

1248 

Aroclor 

1254 

Aroclor 

1260 

1930–56 0 8,486 2,447 2,269 1,614 14,816 

1957 0 903 319 307 423 1,952 

1958 0 649 483 416 355 1,903 

1959 0 1,042 724 518 507 2,791 

1960 0 1,340 556 449 540 2,885 

1961 0 1,852 792 587 611 3,842 

1962 0 1,811 659 554 571 3,595 

1963 0 1,655 935 529 682 3,801 

1964 0 2,085 980 555 755 4,375 

1965 0 2,689 1,025 660 497 4,871 

1966 0 3,180 876 566 472 5,094 

1967 0 3,376 814 525 504 5,219 

1968 0 3,533 853 733 433 5,552 

1969 0 4,165 993 985 452 6,595 

1970 0 4,569 697 1,168 474 6,908 

1971 76 1,466 51 325 121 2,039 

1972 474 22 0 104 9 609 

1973 534 141 0 181 0 856 

1974 498 141 0 140 0 779 

TOTAL 1,582 43,105 13,204 11,571 9,020 78,482 

% of Total 2.0 54.9 16.8 14.7 11.5 100 a

Does not include an additional 132,000 metric tons estimated to have been landfilled during this period. 

Source: Versar, Inc. (1976). 

10-45
 

Table 10-25. Estimated releases of dioxin-like PCB TEQs to the U.S. 

environment, 1930–1977 

Aroclor 

Percent of 

U.S. sales

a 

(1957–1974) 

Estimated 

PCB releases 

(1930–1974)b 

(metric tons) 

Estimated 

mean TEQP ­

WHO98 

concentrationc 

(mg/kg) 

Estimated 

total TEQP ­

WHO98 

released 

(kg) 

1016 12.88 1,582 d d 

1221 0.96 -­ 0.328 -­

1232 0.24 -­ -­ -­

1242 51.76 43,103 7.47 322 

1248 6.76 13,205 16.87 223 

1254 15.73 11,572 125.94 1,457 

1260 10.61 9,019 188.45 1,700 

1262 0.83 -­ -­ -­

1268 0.33 -­ -­ -­

TOTAL 3,702 a

Sales during the period 1957–1974 constituted 63% of all PCB sales during 1930–1977. Sales data for 

 individual Aroclors are not available for years prior to 1957; however, sales of Aroclors 1221, 1232, 1262, and

 1268 were minor even prior to 1957. b

From Table 10-24. c

From Table 10-3 (assumes nondetect values are zero). d

Data are available for only a few samples of Aroclor 1016 where only two dioxin-like PCB congeners were 

detected. The total TEQP-WHO98 released is less than 0.01 kg. 

-- = Indicates that release estimates were not made because of relatively low usage amounts 

Source: Versar, Inc. (1976). 

10-46
 

11. RESERVOIR SOURCES OF CDDs/CDFs AND DIOXIN-LIKE PCBs 

National CDD/CDF source inventories have been conducted in several nations, including 

the United Kingdom (U.K.), the Netherlands, Germany, Austria, and Sweden, to characterize 

emissions from various source categories and estimate annual CDD/CDF emissions to air (and 

sometimes other media). These inventories focused mainly on emissions from primary sources 

(i.e., emissions from the site or process where the CDDs/CDFs are formed). 

The authors of these inventories (Rappe, 1991; Harrad and Jones, 1992; Bremmer et al., 

1994; Thomas and Spiro, 1995, 1996; Eduljee and Dyke, 1996; Jones and Alcock, 1996; DuarteDavidson et al., 1997) indicated that the annual estimates of releases to air provided in these 

inventories may, for several reasons, be underestimates of actual emissions. First, on an 

empirical basis, estimates of the amounts of CDDs/CDFs deposited annually from the 

atmosphere were greater than the estimates of annual CDD/CDF emissions to the atmosphere. 

Second, because the emission test data were limited, the inventories may have underestimated 

releases from known sources or may not have identified all primary sources. Third, the 

investigators were not able to reliably quantify emissions from potential reservoir (secondary) 

sources, including volatilization of CDDs/CDFs from PCP-treated wood, volatilization from soil, 

and resuspension of soil particles. Relatively little research of either a monitoring or a 

theoretical nature has been performed to identify reservoir sources and to quantify the magnitude 

of current or potential future releases from these sources. 

This chapter presents background information on the major reservoir sources of 

CDDs/CDFs and PCBs, including the potential magnitude (mass) of CDDs/CDFs and PCBs in 

each reservoir, the chemical/physical mechanisms responsible for releases of these compounds, 

and estimates of potential annual releases from each reservoir, if such estimates are feasible. 

11.1. POTENTIAL RESERVOIRS 

Chapters 2 through 10 of this document discuss both known and suspected sources of 

releases of newly formed dioxin-like compounds to the environment in the United States. Once 

released into the open environment, CDDs, CDFs, and PCBs partition to air, soils, water, 

sediments, and biota according to both the nature of the release and the contaminant’s chemical 

and physical properties. 

For this analysis, reservoirs are defined as materials or places that contain previously 

formed CDDs/CDFs or dioxin-like PCBs and have the potential for redistributing and circulating 

these compounds into the environment. Potential reservoirs include soils, sediments, biota, 

water, and some anthropogenic materials. The atmosphere is not considered a reservoir, but 

serves as a medium for transporting and distributing CDDs and CDFs over large geographical 

11-1
 

areas. Atmospheric deposition of CDDs/CDFs to the earth’s surface results in an exchange of 

CDDs/CDFs from the atmosphere to surface waters and soils. 

 Dioxin-like compounds are sequestered by a reservoir only until physical processes 

cause these contaminants to become released into the open environment over a defined time and 

space. When this occurs, reservoirs become sources of dioxin-like compounds in the circulating 

environment. Reservoir sources are not included in the quantitative inventory of contemporary 

sources because they do not involve original releases but rather the re-circulation of past releases 

from anthropogenic sources. They can, however, contribute to human exposure and, therefore, 

are important to consider. 

The rate of movement from one environmental medium to another is termed “flux,” and 

it refers to the direction and magnitude of flow and exchange over a reference time period and 

space. Figure 11-1 presents a conceptual diagram of flux and exchange of dioxin-like 

compounds to multiple environmental compartments such as soils, water, air, sediments, and 

biota. This dynamic system consists of fluxes in and out of the atmosphere as well as other 

exchanges between reservoirs and the atmosphere (recall that the atmosphere is not formally 

considered an environmental reservoir, rather it is a transport media for dioxin-like compounds). 

Movement of dioxin-like compounds between media can be induced by the physical processes of 

volatilization, wet and dry atmospheric particle and vapor deposition, adsorption, erosion and 

runoff, resuspension of soils into air, and resuspension of sediments into water. 

Figure 11-1. Fluxes among environmental reservoirs. 

11-2
 
11.2. CHARACTERIZATION OF RESERVOIR SOURCES 

This section is organized according to each principle reservoir, i.e., soil, water, sediment, 

and biota. The discussion of each environmental reservoir provides information in three parts: 

(1) the potential amount or mass of dioxin-like compounds contained within the reservoir, (2) the 

chemical/physical mechanisms responsible for releases of these compounds back into the open 

and circulating environment, and (3) estimates of potential annual releases from the reservoir if 

such estimates are feasible, given the available state of knowledge. Although anthropogenic 

structures (e.g., PCP-treated fenceposts, telephone poles) are potential reservoir sources, they are 

not discussed here because they are covered in Chapter 8 (the most detailed discussion is on 

PCP, Section 8.3.8). 

11.2.1. Soil 

11.2.1.1. Potential Mass of Dioxin-Like Compounds Present 

In estimating burdens for the U.K., Harrad and Jones (1992) and Duarte-Davidson et al. 

(1997) assumed that the majority of CDDs/CDFs in soil is present in the top 5 cm (except 

possibly in cropland, where they may be present at greater depths due to plowing) and that the 

soil density is 1,000 kg/m3. Coupling these assumptions with the rural and urban U.S. surface 

areas and TEQ concentrations yielded soil burden estimates of 1,350 kg TEQDF-WHO98 (1,530 

kg I-TEQDF) in rural soils and 220 kg TEQDF-WHO98 (250 kg I-TEQDF) in urban soils in the 

United States. These calculations are not definitive, and only serve to indicate approximate 

amounts of CDDs/CDFs that may be contained in soils in rural and urban areas of the United 

States. 

Higher concentrations of CDDs/CDFs for background urban and rural soils may be 

present in soils underlain by municipal and industrial waste and in soils at contaminated 

industrial sites. Higher concentrations may also be present in the soils of areas that have been 

treated with pesticides contaminated with CDDs/CDFs. The lack of comprehensive data on 

CDD/CDF concentrations in these soils, as well as the lack of data on the mass of these soils 

nationwide, precludes estimating total national soil burdens of CDDs/CDFs. Because of the lack 

of data, it is not possible to estimate current soil burdens of CDDs/CDFs associated with past 

pesticide use; however, estimates can be made of the total mass of CDD/CDF TEQs that have 

been applied to soil from past use of the pesticides 2,4-dichlorophenoxyacetic acid (2,4-D) and 

2,4,5-trichlorophenoxyacetic acid (2,4,5-T). 

2,4-D (and its salts and esters) are widely used as post-emergence herbicides in 

agricultural and nonagricultural settings in the United States for control of broadleaf weeds and 

brush. Commercial production of 2,4-D in the United States started in 1944 (Esposito et al., 

1980), and it has been in large-scale, large-volume commercial use for many years (U.S. EPA, 

1975). In terms of annual volume, 2,4-D ranks among the top 10 pesticides used in the United 

11-3
 

States (U.S. EPA, 1994d, 1997e). Table 11-1 presents a compilation of domestic production, 

sales, and usage volumes for 2,4-D and its salts and esters. 

Table 11-1. Historical production, sales, and usage of 2,4-dichlorophenoxyacetic acid (2,4-D) (metric tons)a 

Year 

2,4-D, acid 

2,4-D, esters and salts 

(as reported)

b 

Production 

volume 

Sales 

volume 

Domestic usage/ 

disappearance 

Production 

volume 

Sales 

volume 

2000 — — 23,600–28,100c — — 

1998/99 — — 23,600–28,100d — — 

1996/97 — — 23,600–27,200d — — 

1994/95 — — 21,800–26,300e — — 

1993 — — 16,800–20,400f — — 

1992 — — 16,800–20,400f — — 

1991 — — 18,100–29,500g — — 

1990 — — 18,100–29,500g — — 

1989 — — 18,100–29,500h — — 

1988 — — 23,600–30,400i — — 

1987 — — 23,600–30,400j — — 

1986 — — -­ 8,618 12,150 

1985 — — -­ — 0 

1984 — — -­ — 0 

1983 — — -­ 7,702 8,234 

1982 — — -­ 8,762 8,400 

1981 5,859 3,275 -­ 8,987 8,002 

1980 6,164 3,137 -­ 11,313 11,147 

1979 5,763 6,187 -­ 11,874 13,453 

1978 — — -­ 8,958 9,256 

1977 — — -­ 12,552 10,196 

1976 — — 17,418k 10,913 7,813 

1975 — — -­ 16,134 13,414 

1974 — — -­ 6,558 5,991 

1973 — — -­ 13,400 13,698 

1972 24,948b — 21,772k 10,192 10,899 

1971 — 5,619 15,700k — 18,654 

1970 19,766 7,159 -­ — 19,920 

1969 21,354 8,521 -­ 25,854 20,891 

1968 35,953 10,352 -­ 42,690 30,164 

1967 34,990 15,432 -­ 37,988 29,300 

1966 30,927 12,710 28,985l 32,895 25,075 

11-4
 

Table 11-1. Historical production, sales, and usage of 2,4-dichlorophenoxyacetic acid (2,4-D) (metric tons)a

 (continued) 

Year 

2,4-D, acid 

2,4-D, esters and salts 

(as reported)

b 

Production 

volume 

Sales 

volume 

Domestic usage/ 

disappearance 

Production 

volume 

Sales 

volume 

1965 28,721 11,816 22,906l 28,740 21,454 

1964 24,364 11,343 19,958l 24,660 18,263 

1963 21,007 9,446 15,059l 20,178 16,333 

1962 19,503 7,716 16,284l 16,831 13,075 

1961 19,682 7,591 14,107l 16,683 12,533 

1960 16,413 — 14,107l 15,436 13,661 

1959 13,282 7,240 15,468l 12,438 7,070 

1958 14,036 6,234 9,662l 11,295 5,649 

1957 15,536 6,871 -­ 12,392 7,125 

1956 13,079 6,465 -­ 9,635 7,294 

1955 15,656 5,924 -­ 13,390 8,121 

1954 — 4,838 -­ 10,268 6,886 

1953 11,761 — -­ 10,733 8,855 

1952 13,933 — -­ 11,358 9,637 

1951 — — -­ — — 

1950 6,421 4,301 -­ 5,274 3,219 

1949 6,852 2,991 -­ 5,829 3,211 

1948 9,929 4,152 -­ 2,458 1,598 

1947 2,553 2,320 -­ 1,468 1,108 

1946 2,479 2,330 -­ 515 81 

1945 416 286 -­ — — a

All values are from USITC (1946–1994) unless footnoted otherwise. b

No data were reported for domestic usage/disappearance of 2,4-D esters and salts. c

Source: U.S. EPA (1991c). d

Source: U.S. EPA (2000a). e

Source: U.S. EPA (1997e). f

Source: U.S. EPA (1994d). g

Source: U.S. EPA (1992g). h

Source: U.S. EPA (1991d). i

Source: U.S. EPA (1990c). 

j

Source: U.S. EPA (1988d). k

Source: U.S. EPA (1975). l

Source: USDA (1970). 

— = Not reported to avoid disclosure of proprietary data 

-- = No information given 

11-5
 

As described in Section 8.3.8, CDDs/CDFs were detected in several formulations of 2,4­

D and its derivatives during analyses performed to comply with EPA’s 1987 Data Call-In (DCI) 

for CDDs/CDFs. Although the analytical results of these tests indicated that CDDs/CDFs were 

seldom above the regulatory limits of quantification (LOQs) established by EPA for the DCI, 

several registrants detected and quantified CDDs/CDFs at lower LOQs. The results of these 

tests are summarized in Table 8-28 (Chapter 8). The average TEQ in these tests was 1.1 :g 

TEQDF-WHO98/kg (0.7 :g I-TEQDF/kg). Schecter et al. (1997) reported similar concentrations in 

2,4-D samples manufactured in Europe and Russia; lower levels were observed in U.S. products. 

Some of the results from Schecter et al. are presented in Table 8-34 (Chapter 8). 

If it is assumed that the EPA DCI results are typical of CDD/CDF levels in 2,4-D 

pesticides over the past 20 yr and that the average annual use of these pesticides in the United 

States has been approximately 25,000 metric tons, then the estimated CDD/CDF TEQ released 

to the environment from 2,4-D use during the period of 1975 to 1995 was 550 g TEQDF-WHO98 

(350 g I-TEQDF). 

2,4,5-T was used in the United States for a variety of herbicidal applications until the late 

1970s to early 1980s. The major use of 2,4,5-T (about 41% of annual usage) was for control of 

woody and herbaceous weed pests on rights-of-way. The other major herbicidal uses were 

forestry (28% of usage), rangeland (20% of usage), and pasture (5% of usage). Uses of 2,4,5-T 

for home or recreation areas and for lakes, ponds, and ditches were suspended by EPA in 1970; 

rights-of-way, forestry, and pasture uses were suspended by EPA in 1979; and all uses were 

canceled in 1983. 

Table 11-2 presents a compilation of domestic production, sales, and usage volumes for 

2,4,5-T and its salts and esters. As shown in Table 11-2, production and use of 2,4,5-T generally 

increased each year following its introduction in the 1940s until the late 1960s. Production, 

sales, and usage information for the 1970s are generally not available but are reported to have 

steadily declined during that decade (Federal Register, 1979; Esposito et al., 1980). 

Some information is available on the 2,3,7,8-TCDD content of 2,4,5-T, but little 

information is available on the concentrations of the other 2,3,7,8-substituted CDD/CDFs that 

may have been present. Plimmer (1980) reported that 2,3,7,8-TCDD concentrations as high as 

70,000 :g/kg were detected in 2,4,5-T during the late 1950s. In a study of 42 samples of 2,4,5-T 

manufactured before 1970, Woolson et al. (1972) found 500 to 10,000 :g/kg of TCDDs in 7 

samples, and another 13 samples contained 10,000 to 100,000 :g/kg of TCDDs. HxCDDs were 

found in 4 samples at levels between 500 and 10,000 :g/kg and in 1 sample at a concentration 

exceeding 10,000 :g/kg but less than 100,000 :g/kg. The detection limit in the study was 500 

:g/kg. 

11-6
 
Table 11-2. Historical production, sales, and usage of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) (metric tons)a 

Year 

2,4,5-T 

2,4,5-T, esters and salts 

(as reported)

b 

Production 

volume 

Sales 

volume 

Domestic usage/ 

disappearance 

Production 

volume 

Sales 

volume 

1993 — — -­ — — 

1992 — — -­ — — 

1991 — — -­ — — 

1990 — — -­ — — 

1989 — — -­ — — 

1988 — — -­ — — 

1987 — — -­ — — 

1986 — — -­ — — 

1985 — — -­ — — 

1984 — — -­ — — 

1983 — — -­ — — 

1982 — — -­ — — 

1981 — — -­ — — 

1980 — — 900c — — 

1979 3,200–4,100d — -­ — — 

1978 — — 3,200e — — 

1977 — — 4,100d — — 

1976 — — -­ — — 

1975 — — 3,200e — — 

1974 — — 900f — — 

1973 — — -­ — — 

1972 — — -­ — — 

1971 — — 694g

 — 1,675 

1970 — — 3,200e 5,595 3,272 

1969 2,268 — -­ 5,273 2,576 

1968 7,951 1,329 ~7,000h,i 19,297 15,021 

1967 6,601 757 ~7,000h,i 12,333 11,657 

1966 7,026 2,312 7,756h 8,191 4,553 

1965 5,262 — 3,266h 6,131 5,977 

1964 5,186 1,691 4,037h 5,880 3,128 

1963 4,123 1,928 3,266h

 4,543 2,585 

1962 3,796 1,021 3,674h

 4,765 2,543 

11-7
 

Table 11-2. Historical production, sales, and usage of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) (metric tons)a

 (continued) 

Year 

2,4,5-T 

2,4,5-T, esters and salts 

(as reported)

b 

Production 

volume 

Sales 

volume 

Domestic usage/ 

disappearance 

Production 

volume 

Sales 

volume 

1961 3,134 1,196 2,449h

 3,536 2,372 

1960 2,874 — 2,676h

 3,594 1,891 

1959 2,516 1,039 2,495h

 3,644 1,843 

1958 1,668 692 1,724h

 2,372 1,151 

1957 2,419 — -­ 3,098 1,337 

1956 2,345 816 -­ 3,196 1,473 

1955 1,327 662 1,300e 1,720 1,077 

1954 1,223 639 -­ 1,761 615 

1953 2,395 — -­ 2,443 1,817 

1952 1,583 — -­ 1,423 569 

1951 — — 1,100e — — 

1950 852 297 -­ — — 

1949 — — -­ — — 

1948 — — -­ — — 

1947 — — -­ — — 

1946 — — -­ — — 

1945 — — -­ — — 

a

All values are from USITC (1946–1994) unless footnoted otherwise.
 
 b

No data were reported for domestic usage/disappearance of 2,4,5-T esters and salts.
 
 c

Source: Esposito et al. (1980).
 
 d

Source: Federal Register (1979).
 
 e

Source: Thomas and Spiro (1995).
 
 f

Source: U.S. EPA (1977).
 
 g

Source: USDA (1971); reflects farm usage only.
 
 h

Source: USDA (1970); values include military shipments abroad.
 
 i

Source: Kearney et al. (1973) reports slightly lower domestic consumption for the years 1967 and 1968 than for

 1966.
 


— = Not reported to avoid disclosure of proprietary data 

-- = No information given 

The average 2,3,7,8-TCDD concentration in 200 samples of Agent Orange, a defoliant 

containing about a 50/50 mixture of the butyl esters of 2,4,5-T and 2,4-D that was used by the 

U.S. Air Force in Vietnam, was 1,910 :g/kg (Kearney et al., 1973). Of the 200 samples, 64 

(32%) contained more than 500 :g/kg of 2,3,7,8-TCDD, with the highest concentration reported 

to be 47,000 :g/kg. Storherr et al. (1971) reported detecting 2,3,7,8-TCDD at concentrations 

11-8
 

ranging from 100 to 55,000 :g/kg in five samples of 2,4,5-T. Kearney et al. (1973) reported that 

production samples of 2,4,5-T obtained from the three principal 2,4,5-T manufacturers in 1971 

contained 2,3,7,8-TCDD at levels of <100 :g/kg, 100 :g/kg, and 2,300 :g/kg. 

A 1975 survey of 10 lots of a commercial formulation containing 2,4,5-T showed 2,3,7,8­

TCDD concentrations ranging from 10 to 40 :g/kg (Lewert, 1976). Analyses by EPA of 16 

technical-grade 2,4,5-T samples from five different manufacturers revealed 2,3,7,8-TCDD 

contents ranging from <10 to 25 :g/kg (Federal Register, 1979). Schecter et al. (1997) reported 

the analytical results of one sample of 2,4,5-T purchased from Sigma Chemical Co. (product 

number T-5785, lot number 16H3625). The results, presented in Table 11-3, indicate a total 

TEQDF-WHO98 concentration of 3.26 :g/kg (2.88 :g I-TEQDF/kg). 

Because of the wide variability (three orders of magnitude) in the limited available 

information on the 2,3,7,8-TCDD content of 2,4,5-T (particularly the 2,4,5-T used in the 1950s) 

and incomplete information on domestic usage, it is difficult to reliably estimate the amount of 

2,3,7,8-TCDD that was released to the U.S. environment as a result of 2,4,5-T use. A very 

uncertain estimate can be made using the following assumptions: (1) average annual 

consumption during the 1950s, 1960s, and 1970s was 2,000, 4,000, and 1,500 metric tons/yr, 

respectively; and (2) the average 2,3,7,8-TCDD concentration in 2,4,5-T used over these three 

decades was 10,000 :g/kg in the 1950s, 4,000 :g/kg in the 1960s, and 100 :g/kg in the 1970s. 

Based on these assumptions, the very uncertain estimate of 2,3,7,8-TCDD input from 2,4,5-T use 

over the period of 1950 to 1979 was 36,000 g. 

Another contributing source to the soil reservoir is CDD/CDF in sewage sludge applied 

to land (i.e., surface disposal or land farming), estimated to have been 75 g TEQDF-WHO98 (103 

g I-TEQDF) in 1995 (see Section 8.4.1 for details). If this same amount of TEQ had been applied 

each year during the period of 1975 to 1995, the total amount applied would have been 1,500 g 

TEQDF-WHO98 (2,000 g I-TEQDF). 

11.2.1.2. Mechanisms Responsible for Releases from Surface Soils 

The atmospheric deposition of dioxin-like compounds is believed to be the primary 

means whereby surface soils, in general, become contaminated with dioxin-like compounds. 

CDDs/CDFs and PCBs are highly lipid soluble and have low volatility, and they tend to partition 

to soil rather than into air or water. Once present in or on soils, physical/chemical and biological 

mechanisms (photolysis and biodegradation) can slowly alter the composition and amount of 

compound present. Studies indicate that the dioxin-like compounds (particularly the higherchlorinated CDDs/CDFs) exhibit little downward mobility to lower soil depths after they are 

deposited in or on soil (Puri et al. 1989; Freeman and Schroy, 1985; Orazio et al., 1992; 

Paustenbach et al., 1992). However, re-mobilization of the contaminants to the atmosphere is 

possible through volatilization and resuspension of soil particles. 

11-9
 
Table 11-3. CDD/CDF concentrations (:g/kg) in recent sample of 

2,4,5-trichloro-phenoxyacetic acid (2,4,5-T) 

Congener/congener group 2,4,5-T sample 

2,3,7,8-TCDD 

1,2,3,7,8-PeCDD 

1,2,3,4,7,8-HxCDD 

1,2,3,6,7,8-HxCDD 

1,2,3,7,8,9-HxCDD 

1,2,3,4,6,7,8-HpCDD 

OCDD 

1.69 

0.412 

0.465 

2.28 

1.35 

18.1 

33.9 

2,3,7,8-TCDF 

1,2,3,7,8-PeCDF 

2,3,4,7,8-PeCDF 

1,2,3,4,7,8-HxCDF 

1,2,3,6,7,8-HxCDF 

1,2,3,7,8,9-HxCDF 

2,3,4,6,7,8-HxCDF 

1,2,3,4,6,7,8-HpCDF 

1,2,3,4,7,8,9-HpCDF 

OCDF 

0.087 

0.102 

0.183 

1.72 

0.356 

ND (0.012) 

0.126 

2.9 

0.103 

3.01 

Total 2,3,7,8-CDDa 

Total 2,3,7,8-CDFa 

Total I-TEQDF 

a 

Total TEQDF-WHO98 

a 

58.2 

8.59 

2.88 

3.05 

Total TCDD 

Total PeCDD 

Total HxCDD 

Total HpCDD 

Total OCDD 

Total TCDF 

Total PeCDF 

Total HxCDF 

Total HpCDF 

Total OCDF 

-­

-­

-­

-­

-­

-­

-­

-­

-­

-­

Total CDD/CDF -­ a

 Calculated assuming nondetect values were zero. 

ND = Not detected (value in parenthesis is the detection limit) 

-- = No information given 

Source: Schecter et al. (1997). 

11-10
 

Young (1983) conducted field studies on the persistence and movement of 2,3,7,8-TCDD 

from 1973 to 1979 on a military test area that had been aerially sprayed with 73,000 kg of 

2,4,5-T from 1962 to 1970. TCDD levels of 10 to 1,500 ng/kg were found in the top 15 cm of 

soil 14 yr after the last application of herbicide at the site. Although actual data were not 

available on the amount of 2,3,7,8-TCDD originally applied as a contaminant of the 2,4,5-T, best 

estimates indicated that less than 1% of the applied 2,3,7,8-TCDD remained in the soil after 

14 yr. Young suggested that photodegradation at the time of and immediately after aerial 

application was responsible for most of the disappearance; however, once incorporated into the 

soil, the data indicated a half-life of 10 to 12 yr. Similarly, Paustenbach et al. (1992) concluded 

that the half-life of 2,3,7,8-TCDD in soils at the surface might be 9 to 15 yr and the half-life 

below the surface could be 25 to 100 yr. 

Ayris and Harrad (1997) studied the mechanisms affecting volatilization fluxes of several 

PCB congeners (PCB 28, 52, 101, 138, and 180) from soil and found positive correlations 

between flux and soil temperature, soil moisture content, and soil PCB concentration. For PCBs, 

secondary releases from soils (primarily via volatilization) are believed to currently exceed 

primary emissions in the U.K. (Harner et al., 1995; Jones and Alcock, 1996). Lee et al. (1998) 

quantified PCBs in air samples taken every 6 hr over a 7-day period in the summer at a rural site 

in England and found a strong correlation between air temperature and PCB congener 

concentrations. The concentrations followed a clear diurnal cycle, thus providing some evidence 

that rapid, temperature-controlled soil-to-air exchange of PCBs influences air concentrations and 

enables regional/global scale cycling of these compounds. 

CDDs/CDFs and PCBs sorbed to soil and urban dust particles can also be moved from 

the terrestrial environment to the aquatic environment through contaminated storm water runoff 

and surface erosion of contaminated soils. Results of recent research indicate that, for at least 

some water bodies, soil erosion and storm water runoff can be an important way of transporting 

CDDs/CDFs into the surface water. This is especially true for water bodies lacking nonpoint 

sources of dioxin-like compounds. Smith et al. (1995) analyzed CDD/CDF concentrations in 

sediment cores, air, precipitation, soil, and storm water runoff in an effort to determine the 

contributing sources of these compounds to the lower Hudson River. The mass balance 

estimates developed from these data for 1990 to 1993 are as follows: stormwater runoff entering 

tributaries (76% of total CDD/CDF input), anthropogenic wastes (19%), atmospheric deposition 

(4%), and shoreline erosion (less than 1%). The authors projected the percent contribution of 

these same sources for 1970 as anthropogenic wastes (70%), stormwater runoff into tributaries 

(15%), atmospheric deposition (15%), and shoreline erosion (0.1%). 

Studies conducted by Paustenbach et al. (1996), Mathur et al. (1997) and Fisher et al. 

(1999) discerned the relative concentrations of CDDs and CDFs in storm water. Paustenbach et 

al. and Mathur et al. reported that stormwater runoff from 15 sites in the San Francisco area 

11-11
 

contained CDD/CDF TEQ at levels ranging from 0.01 to 65 pg I-TEQDF/L; most samples 

contained less than 15 pg I-TEQDF/L. The sites differed widely in land use; the highest levels 

measured were obtained from an urban but nonindustrialized area. A distinct variability was 

noted in the results obtained at the same sampling location during different rain events. The 

profiles of CDDs/CDFs in the urban stormwater samples were similar, particularly in samples 

collected at the onset of rain events. Stowe (1996) reported similar findings from analyses of 

sediments from three stormwater basins collecting runoff from a military base, a city street, and 

parking lots. 

Fisher et al. (1999) reported that urban runoff samples from eight sites (15 samples) in 

the Santa Monica Bay watershed contained CDD/CDF TEQ at levels ranging from 0.7 to 53 pg 

I-TEQDF/L (all but one sample were in the range of 0.7 to 10 pg I-TEQDF/L). The samples were 

collected in 1988/1989 from continuously flowing storm drains during both dry and storm 

periods. The mean concentration measured during storm events, 18 pg I-TEQDF/L, was higher 

than concentration observed during dry periods, 1 pg I-TEQDF/L. 

Atmospheric deposition of CDDs and CDFs to lakes and the watershed area of water 

bodies is another important means of contaminating bottom sediments. For example, Lebeuf et 

al. (1996) analyzed sediment cores from different locations in the lower St. Lawrence River 

Estuary and the Gulf of St. Lawrence. The congener group profiles found in the samples 

indicated that the input of CDDs/CDFs was primarily from the atmosphere. Comparison of the 

CDD/CDF concentrations in sediments collected from areas where sediment accumulation is due 

primarily to fluvial transport with sediments from areas where sediment accumulation is due 

primarily to direct atmospheric deposition onto the water indicates that the contribution of 

CDDs/CDFs from direct atmospheric deposition represents less than 35% of the sediment 

burden. Thus, the primary source of CDDs/CDFs is emissions to the atmosphere upwind of the 

estuary that are deposited within the watershed and subsequently transported downstream by 

fluvial waters. 

11.2.1.3. Estimated Annual Releases from Soil to Water 

Nonpoint sources of CDDs/CDFs to waterways include stormwater runoff from urban 

areas and soil erosion in rural areas during storms. Approaches for estimating national loadings 

to water for both of these sources are described below. The estimate derived for the potential 

annual national loading of CDDs/CDFs in urban runoff to waterways is uncertain, but it suggests 

that the loading may be comparable to the contribution from known industrial point sources (at 

least 20 g I-TEQDF in 1995). Similarly, the estimate derived for the potential annual national 

loading of CDDs/CDFs in rural eroded soils to waterways is uncertain, but it has a stronger 

analytical base than does the urban runoff estimate. This loading estimate, however, is 

significantly higher than the contribution from known industrial point sources. 

11-12
 
Urban runoff. Few data on CDD/CDF concentrations in urban runoff have been 

reported. The most recent and largest data sets were reported in studies conducted in the San 

Francisco Bay and Santa Monica Bay regions (Mathur et al., 1997; Fisher et al., 1999). These 

studies found a wide range of CDD/CDF levels in samples of stormwater runoff from 23 sites, 

varying from 0.01 to 83 pg I-TEQDF/L. The wide variability and limited geographic coverage of 

these data preclude derivation of a national emissions estimate at this time. However, by making 

a number of assumptions, a preliminary estimate of the potential CDD/CDF magnitude from this 

source can be made. 

In order to estimate the amount of rainfall in urbanized areas of the conterminous United 

States, a Geographic Information System (GIS) analysis was performed to determine the total 

area of every U.S. Census urbanized area and the 30-yr annual average rainfall for each of those 

areas and to calculate the product of the total areas of urbanized areas with the annual average 

rainfall (Lockheed Martin Corp., 1998). This approach yields an estimate of 1.9 × 1014 L/yr. If 

it is assumed that urban runoff in the United States averages 1 pg TEQDF-WHO98/L (1 pg ITEQDF/L) (i.e., approximately the midpoint of the range reported by Mathur et al., 1997, and 

Fisher et al., 1999), this source could contribute a total of 190 g TEQDF-WHO98 or I-TEQDF/yr to 

U.S. waterways. No data were available to make similar estimates for PCBs. 

A similar analysis was conducted using historical precipitation data from the National 

Oceanic and Atmospheric Administration (NOAA, 2004) and metropolitan/urban area statistics 

from the 1990 and 2000 census. The 30-year annual average rainfall for each state was 

calculated for 1987, 1995, and 2000. An approximation of the urban area for each state was 

estimated by summing the acreage for each metropolitan area identified in the 1990 census. 

Assuming that the amount of land classified as urban did not change significantly from 1987 to 

1990, the urban areas for each state in 1990 were assumed to be equal to those in 1987. 

Similarly, an approximation for urban area for each state was estimated by summing the urban 

area acreage available from the 2000 census. An approximation of the 1995 urban area for each 

State was estimated by taking the average of the 1990 and 2000 estimates. Multiplying the 30­

year average rainfall by the urban area for each state and summing the results provides an 

estimated amount of urban runoff for the conterminous United States. The urban runoff was 

1.24 × 1014, 1.33 × 1014, and 1.42 × 1014 L/year for 1987, 1995, and 2000, respectively. 

Applying the emission factors generated above, urban runoff contributed 124, 133, and 142 g I

TEQDF or TEQDF-WHO98 to U.S. waterways in 1987, 1995, and 2000, respectively. These 

numbers are in agreement with the estimate developed using Lockheed Martin Corp. (1998) 

data. 

Rural soil erosion. Using acreage and erosion factors for cropland provided in the 2001 

Annual National Resources Inventory (USDA, 2003), 1.36, 1.07, 0.96, and 0.91 billion metric 

tons of soil and rill erosion were generated in 1987, 1992, 1997, and 2001, respectively. 

­

11-13
 

Likewise, using acreage data for rangeland from USDA (2003) and a soil and rill erosion factor 

of 4.2 tons/acre/yr (USDA, 1995), approximately 1.55 billion metric tons of soil and rill erosion 

were generated in 1987, 1992, 1997, and 2001. For purposes of estimating values for reference 

years 1995 and 2000, it was assumed that the 1995 erosion estimate was the average of soil and 

rill erosion estimates developed for 1992 and 1997, and that the 2001 numbers approximated 

those generated in 2000. The total amount of eroded soil entering waterways was greater than 

this value, because this value did not include soil erosion from construction areas, forests, and 

other non-croplands and non-rangelands. 

The data summarized in U.S. EPA (2000c) suggest that typical concentrations of 

CDDs/CDFs in soils in rural areas is about 2.8 ng TEQDF-WHO98/kg. It is not known how well 

this estimate represents eroded soil from cropland and rangeland. If these soils contain an 

average of 1 ng TEQDF/kg (i.e., a lower value than the background value for all types of rural 

soil), they would have contributed 2,900, 2,600, and 2,500 g TEQDF-WHO98 to the nation’s 

waterways in 1987, 1995, and 2000, respectively. Given the uncertainties in both the amount of 

eroded soil and dioxin levels, these estimates are considered preliminary (i.e., Category D). As 

with urban runoff, no data were available to make similar estimates for PCBs. 

11.2.1.4. Estimated Annual Releases from Soil to Air 

No quantitative estimates of the mass of dioxin-like compounds that may be released to 

the atmosphere annually from U.S. soils have been published in the literature and none were 

developed for this report. As noted above, the vapor flux of these compounds from soil to air is 

dependent on the soil and air concentrations of dioxin-like compounds and the temperature, 

moisture content, and organic carbon content of the soil. Most of these parameters are not 

characterized well enough for the United States as a whole to enable a reliable estimate to be 

made at present. Particle flux is dependent on many factors, including wind speed, vegetative 

cover, activity level, particle size, soil type/conditions, moisture content, and particle density. 

Through use of models and various assumptions, Kao and Venkataraman (1995) estimated the 

fraction of ambient air CDD/CDF concentrations in the upper midwestern section of the United 

States that may be the result of atmospheric re-entrainment of soil particles. Similarly, through 

use of models and various assumptions, Jones and Alcock (1996) and Harner et al. (1995) 

reached tentative conclusions about the relative importance of volatilization of dioxin-like 

compounds from soils in the U.K. 

Modeling re-entrainment of soil to the atmosphere was conducted by Kao and 

Venkataraman (1995). Their model incorporated information on particle sizes, deposition 

velocities, and concentrations of CDDs/CDFs in soils. Smaller particulates, with median 

diameters ranging from about 0.01 :m to 0.3 :m, are primarily formed from combustion sources 

when hot vapors condense and through accumulation of secondary reaction products on smaller 

11-14
 

nuclei. Particles at the upper end of this size range will deposit to the ground in several days. 

Large or coarse particles, having median diameters of about 8 :m, are generated from windblown dust, sea spray, and mechanically generated particles. CDDs/CDFs absorbed onto 

re-entrained soil would be included in this larger particle size. These larger particles have a 

lifetime in the atmosphere from a few to many hours. 

The fraction of ambient air concentration of CDDs/CDFs that results from soil 

re-entrainment was established on the basis of the contribution of crustal sources to the ambient 

aerosol. Data on typical crustal soil concentrations in air (15 to 50 :g/m3

 for rural areas and 5 to 

25 :g/m3

 for urban areas) were combined with data on the average concentrations of 

CDDs/CDFs in soils (73 ng/kg for rural, 2,075 ng/kg for urban, and 8,314 ng/kg for industrial 

soils) published by Birmingham (1990) for Ontario, Canada, and several U.S. midwestern states. 

This analysis estimated the concentrations of CDDs/CDFs in the ambient aerosol that originate 

from soils to be 1 × 10-3 to 4 × 10-3 pg/m3 in rural areas and 0.01 to 0.05 pg/m3 in urban areas. 

These particulate dioxin concentrations were compared with average total particulate dioxin 

levels of 1.36 pg/m3 in Eitzer and Hites (1989) to arrive at the conclusion that soil re-entrainment 

could account for only 1 to 4% of the particulate dioxins in the atmosphere in urban areas and 

0.1 to 0.3% of those in rural regions (Kao and Venkataraman, 1995). 

This information on the size distribution of ambient aerosols and relative CDD/CDF 

concentrations in different particle size fractions was integrated with particle size deposition 

velocities to estimate the relative contribution to the total mass deposition flux for small and 

large particle sizes. Even though re-entrained soil may constitute only a small fraction of the 

atmospheric levels of CDDs/CDFs, the contribution of dioxins in re-entrained surface soil to the 

total deposition flux could be significant because coarse particles dominate in dry deposition. 

Soil re-entrainment could possibly account for as much as 70 to 90% of the total dry deposition 

of CDDs/CDFs in urban areas and 20 to 40% in rural regions (Kao and Venkataraman, 1995). 

Two approaches were used by Jones and Alcock (1996) to assess the potential 

significance of CDD/CDF volatilization from soils: the fugacity quotient concept and a simple 

equilibrium partitioning model. The fugacity quotient model compares the fugacity (the 

tendency of a chemical to escape from a phase) of individual CDD/CDF compounds in different 

environmental media to determine the tendency for these compounds to accumulate in particular 

environmental compartments (McLachlan, 1996). Fugacities for individual compounds, by 

media, were estimated by Jones and Alcock on the basis of physical/chemical properties of the 

compounds as well as the concentrations in the media. In this instance, fugacity quotients were 

calculated for air and soil by dividing each compound’s fugacity for air by that of soil. 

Quotients near 1 indicate equilibrium conditions between media; values greater than 1 represent 

a tendency for flux (volatilization) from soil to air, and values less than 1 indicate a net flux to 

the soil from the air. The equilibrium partitioning model used by Jones and Alcock predicts the 

11-15
 

maximum (possible “worst case”) flux of CDDs/CDFs from soil to the atmosphere. Air phaseto-soil partition coefficients were calculated using the ratios of soil and air fugacity capacities. 

Equilibrium air concentrations were then calculated using typical U.K. soil concentrations for 

both urban and rural settings. 

From the fugacity quotient model, Jones and Alcock (1996) concluded that the lowerchlorinated CDDs/CDFs may be close to soil-air equilibrium in the U.K., whereas for other 

congeners, soil is a sink rather than a source to the atmosphere. The authors reported that the 

equilibrium partitioning model predicted that 0.15 kg I-TEQ volatilizes annually from soil in the 

U.K. However, they discounted this estimate and concluded that soil volatilization is unlikely to 

be a significant contributor to emissions. The likelihood that these estimates were high was 

attributed to the fact that assumptions were made that the concentrations of CDDs/CDFs in air 

were zero and the model does not consider the resistance of CDDs/CDFs to volatilize from soil. 

Harner et al. (1995) developed a model to predict the long-term fate of PCBs in soils, 

with emphasis on soil-to-air exchanges. Using data on levels of PCBs in air, soil, and vegetation 

in the U.K., the investigators developed a mass balance model to simulate the fate of PCBs in 

U.K. soils from 1935 to 1994. Specifically, monitoring data and physical/chemical property data 

were compiled to calculate fugacities for PCB congeners 28, 52, 138, and 153. The model was 

designed to provide an order-of-magnitude level of accuracy, due in part to the inherent 

variability in the input data. The mass balance equations in the model included a bell-shaped 

function for rates of emissions of PCBs, with the maximum emission rate occurring in 1967. 

From these emissions rates, fluxes between air and soil over several decades were estimated. 

Table 11-4 summarizes the calculated fluxes. 

Table 11-4. PCB 138 fluxes predicted by Harner et al. (1995) 

Year 

Concentration 

in air 

(pg/m3) 

Fugacity 

in air 

(Pascals × 10-9) 

Fugacity 

in soil 

(Pascals × 10-9) 

Concentration 

in soil 

(ng/g) Net flux/direction 

1950 4 0.24 1.1 -­ air 6 soil (444 kg/yr) 

1965 280 1.5 12 -­ air 6 soil (1,000 kg/yr) 

1975 -­ -­ 16 -­

1980 49 -­ -­ -­ soil 6 air (820 kg/yr) 

1994 6 -­ 8.3 -­ soil 6 air (700 kg/yr) 

-- = No information given 

During the 1960s and 1970s, levels of total PCBs in U.K. soils reached average levels of 

approximately 300 :g/kg as a result of atmospheric deposition. Because of restrictions on PCB 

11-16 
use during the last two decades, air concentrations have fallen, and the primary source to the 

atmosphere is now believed to be volatilization from soils. The mass balance model estimated a 

net flux of 700 kg/yr of total PCBs from soils to the atmosphere in 1994. However, this estimate 

is presented with the caveat that the model tends to underestimate the rate of reduction of PCB 

concentrations in recent years, which could be attributed to other mechanisms such as 

biodegradation, photolysis, and other degradation processes. 

11.2.2. Water 

11.2.2.1. Potential Mass of Dioxin-Like Compounds Present 

The surface area of inland waters (including the Great Lakes) in the United States is 

about 359,000 km2 (U.S. DOC, 1995b). Assuming that the mean depth of inland water is 10 m 

(Duarte-Davidson et al., 1997), the total inland water volume is approximately 3,600 billion m3. 

No compilation of CDD/CDF measurements in inland surface waters was made for this report; 

however, if it is assumed that the “typical” value used by Duarte-Davidson et al. for rivers in the 

U.K., 38 pg I-TEQDF/m3, is representative of U.S. waters, then the burden is calculated to be 137 

g I-TEQDF. 

11.2.2.2. Mechanisms Responsible for Supply to and Releases from Water 

As discussed in Section 11.2.1.2, dioxin-like compounds enter surface water from 

atmospheric deposition, stormwater runoff erosion, and discharges of anthropogenic wastes. 

Volatilization is the primary mechanism for release of dioxin-like compounds from the water 

column to the atmosphere. Several studies have addressed the water-air exchange of dioxin-like 

PCBs through volatilization in the Great Lakes (Achman et al., 1993; Hornbuckle et al., 1993; 

Swackhamer and Armstrong, 1986; Baker and Eisenreich, 1990). No similar body of literature 

has been developed to address volatilization of CDDs/CDFs from water. 

Most studies that have addressed PCB water-air exchange have used the two-film model 

developed by Whitman (1927) and made popular by Liss and Slater (1974). When assessing gas 

exchange between air and water, the interface between the two phases can be considered as a 

two-layer (film) system consisting of well-mixed gas and liquid films adjacent to the interface; 

the rate of transfer is controlled by molecular diffusion through the stagnant boundary layer 

(Achman et al., 1993). Liss and Slater (1974) applied the model to assess the flux of various 

gases, specifically in the air-sea systems, and indicated the possibility of its use at any air-water 

interface in the environment if the necessary data are available. Hornbuckle et al. (1993) 

concluded that the two-film model is the best available tool for estimating regional and local flux 

of PCBs from natural waters. The following paragraph, from Achman et al. (1993), succinctly 

summarizes the model. 

11-17
 
The basic equation used to describe the rate of transfer across the interface is 

F = Kol(Cw – C*) (11-1) 

where F is the flux (mol/m2

-day), Cw (mol/m3) is the dissolved PCB concentration in the bulk 

water, and C* is the air concentration expressed as a water concentration in equilibrium with the 

air (P/H, mol/m3). The variable P is the vapor-phase air concentration measured (mol/m3) and 

converted to units of pressure using the ideal gas law; H is Henry’s Law constant (atm-m3/mol). 

The overall mass-transfer coefficient, Kol, has units of velocity (m/day). The concentration 

gradient determines the direction of flux and drives the mass transfer, whereas Kol is a kinetic 

parameter that quantifies the rate of transfer. The value of Kol is dependent on the physical and 

chemical properties of the compound as well as environmental conditions. The reciprocal of Kol 

is the total resistance to transfer expressed on a gas (RT/Hka)- and liquid (1/kw)-phase basis: 

1/Kol = 1/kw + RT/Hka (11-2) 

where kw is the water-side mass transfer coefficient (m/day) and ka is the air-side mass transfer 

coefficient (m/day). R is the universal gas constant (8.2057 × 10-5 atm-m3/mol K), T is the 

absolute temperature, and H is Henry’s Law constant. 

Achman et al. (1993) and Hornbuckle et al. (1993) calculated the volatilization rates of 

PCBs from Green Bay on Lake Michigan on the basis of air and water samples simultaneously 

collected over a 14-day period above and below the air-water interphase and analyzed for 85 

PCB congeners. Air samples collected over nearby land were also analyzed for the 85 PCB 

congeners. The direction and magnitude of flux for each congener were then calculated using 

Henry’s Law and meteorological and hydrological parameters in the “two-film” model (see eq 

11-1). 

The net total PCB transfer rate (i.e., the sum of all congener transfer rates) was found to 

be from water to air (i.e., volatilization). However, during cool water temperature periods 

(October), the direction of transfer reversed for many congeners. Calculated transfer rates to air 

ranged from 15 to 300 ng/m2

/day at low wind speeds (1 to 3 m/sec) to 50 to 1,300 ng/m2/day at 

higher wind speeds (4 to 6 m/sec). On a congener basis, the lower-chlorinated congeners 

dominated total fluxes. The summary of flux calculations is presented in Table 11-5. The most 

important factors influencing the magnitude of volatilization were the water concentration of 

PCBs, wind speed, and water temperature. In addition, Achman et al. (1993) and Hornbuckle et 

al. (1993) found that atmospheric PCB concentrations were higher over contaminated water than 

over nearby land, atmospheric PCBs over water tended to increase with increasing dissolved 

11-18
 

Table 11-5. Summary of flux calculations for total PCBs in Green Bay 

Date Site 

Fluxa 

(ng/m2-day) 

6-4 18 40 

6-5 18 40 

6-6 10 95 

6-7 10 155 

6-10 4 325 

6-11 10 13 

7-28 18 330 

7-29 21 70 

7-30 14 225 

7-31 10 90 

8-1 4 800 

10-21 14 555 

10-22 10 1,300 

10-23 4 30 a

Numbers indicate water-to-air transfer of total PCBs. They represent the sum of individual PCB congener fluxes 

and are described as “daily” fluxes because they correspond to air samples collected over 5–10 hr and water samples 

collected over ~1 hr. 

Source: Achman et al. (1993). 

PCB concentrations, and the congener distribution in the atmosphere correlated linearly with the 

congener distributions in the adjacent water. 

Achman et al. (1993) also summarized the PCB volatilization rates reported by other 

researchers (Baker and Eisenreich, 1990; Swackhamer and Armstrong, 1986; Strachan and 

Eisenreich, 1988; Swackhamer et al., 1988) for Great Lakes water bodies. The results of these 

other studies, presented below, also show net flux of PCBs from water to air. 

Total PCB
 

volatilization rate
 

Water body (ng/m2/day) Reference
 

Lake Superior 141 Baker and Eisenreich (1990) 

Lake Michigan 240 Strachan and Eisenreich (1988) 

Lake Superior 63 Strachan and Eisenreich (1988) 

Siskiwit Lake 23 Swackhamer et al. (1988) 

Lake Michigan 15 Swackhamer and Armstrong (1986) 

11-19
 
11.2.3. Sediment 

11.2.3.1. Potential Mass of Dioxin-Like Compounds Present 

EPA conducted congener-specific measurements of CDDs/CDFs in the sediments from 

11 U.S. lakes located in areas relatively unimpacted by nearby industrial activity. The mean 

TEQ concentration in the uppermost sediment layers from these 11 lakes was 5.3 ng TEQDFWHO98/kg (5.3 ng I-TEQDF/kg) dry weight. For most of the lakes, the uppermost layer 

represents about 10 years worth of sedimentation. CDD/CDF concentrations in lakes impacted 

by industrial activity may have higher concentrations. For example, Duarte-Davidson et al. 

(1997) reported a TEQ concentration of 54 ng I-TEQDF/kg for urban sediments in the U.K. 

As noted above, the surface area of inland waters in the United States is approximately 

359,000 km2 (U.S. DOC, 1995b). In their calculations of sediment burdens in the U.K., DuarteDavidson et al. (1997) assumed that (1) the sediment surface area equals the water surface area, 

(2) the majority of CDDs/CDFs were located in the top 5 cm of sediment, and (3) sediment 

density was 0.13 g dry weight/cm3. Applying these assumptions to the water surface area and 

background TEQ concentration for U.S. sediments yields a burden of at least 120 kg TEQDFWHO98 (120 kg I-TEQDF). 

11.2.3.2. Mechanisms Responsible for Supply to and Releases from Sediment 

Because sediment is closely connected to the water column above it, evaluating the 

potential for sediment to act as a reservoir of dioxin-like compounds is complex and likely to be 

more difficult than studying dioxin-like compounds in a single medium, such as water or soil. 

Volatilization and sedimentation are two mechanisms whereby persistent chemicals such as 

CDDs/CDFs and PCBs are lost from water bodies/columns. Numerous authors (Swackhamer 

and Armstrong, 1986; Muir et al., 1985; Ling et al., 1993) have noted that sediments are a likely 

sink for persistent hydrophobic organic compounds because these compounds are likely to be 

strongly bound to organic particles in the sediment. 

For example, Muir et al. (1985) radiolabeled 2,3,7,8-TCDD and studied its dissipation 

from sediments (collected from a farm pond and a lake) to the water column in laboratory 

studies under static aerobic conditions at 10EC. After 675 days, more than 80% of the labeled 

TCDD was still present in the pond sediment and 87% was still present in the lake sediment. 

Aeration had little effect on the dissipation rates. 

The concept of fugacity is a useful way to estimate the behavior of dioxin-like 

compounds in sediments. Fugacity is expressed in units of pressure (pascals, or Pa) and is the 

partial pressure exerted by the chemical in each medium. Fugacity models estimate equilibrium 

concentrations in specific media at given chemical concentrations in the environment. Clark et 

al. (1988) suggested evaluating contaminant concentrations in multiple environmental media by 

comparing fugacity of adjoining media (e.g., comparing sediment fugacity with water column 

11-20
 

fugacity to determine a chemical’s tendency to move from one to the other). The authors 

evaluated fugacities of certain organochlorine compounds, including PCBs, in air, water, 

sediment, fish, and fish-eating birds and their eggs. The authors presented PCBs fugacities 

developed from data collected in a study of the Lake Ontario region. The fugacities of PCBs in 

various media can be ranked as birds>fish>water>bottom sediment, indicating that PCBs and 

other similar chemicals are likely to remain in bottom sediment and are less likely to re-enter the 

water column. 

11.2.3.3. Releases from Sediment to Water 

Given the lack of data, no quantitative estimates of annual releases can be made. Ling et 

al. (1993) evaluated the fate of various chemicals, including PCBs, in Hamilton Harbour, located 

in Ontario, Canada, using a modified version of the Quantitative Water Air Sediment Interaction 

(QWASI) fugacity model. Among the processes evaluated were diffusion between air and water 

and sediment and water; sediment deposition, resuspension, and burial; and sediment 

transformation. Three primary compartments were studied: air, water, and bottom sediments. 

The sediment was treated as a simple, well-mixed surface layer of active sediment and the buried 

sediment underneath. Chemicals in the active sediment were assumed to be able to exchange 

with the overlying water; chemicals in the buried sediment were assumed to be isolated from the 

sediment-water exchange. Sediment was assumed to be homogenous rather than heterogenous. 

The epi- and hypolimnetic compartments of the water column were defined on the basis of a 

thermocline, and the atmosphere was defined as a semi-infinite medium of constant, defined 

composition. 

Ling et al. estimated rates of PCB movement on the basis of 1987 loadings using two 

models: one with a thermocline and one without a thermocline. The results for the watersediment transfer using the model with a thermocline were ~32 kg/yr entering the hypolimnion 

from the epilimnion, ~27 kg/yr entering the surface sediment from the hypolimnion, and ~18 

kg/yr (>50%) going to burial. For sediment-to-water transfer, ~7 kg/yr transferred to the 

hypolimnion and then 12.5 kg/yr transferred to the epilimnion. Similar numbers were found in 

the single water column model (the model without a thermocline). 

Both the model with a thermocline and the model without a thermocline predicted 

volatilization from the water to the atmosphere—1.6 kg/yr and 1.8 kg/yr, respectively. 

However, the actual contribution of PCBs from sediment to air was not determined. A 

comparison of estimated concentrations with observed values are presented in Table 11-6. For 

PCBs, 68% were buried in the sediment, 20% were exported to Lake Ontario, 5.4% degraded in 

the water and sediment, and 6% volatilized. The authors noted that these percentages were 

uncertain. At the sediment-water exchange, more than 90% of each chemical was contained in 

the sediment because of particle deposition and the high affinity of the chemical for sediment. 

11-21
 

Table 11-6. Comparison of model-estimated PCB concentrations with 

observed values 

Variable PCBs 

Observed concentration 

Sediment 

Water 

0.23–1.04 :g/g 

<20 :g/m3 

Estimated concentration from model without thermocline 

Sediment 

Water 

Amount in sediment 

Amount in water 

Total mass 

0.518 :g/g 

8.33 :g/m3 

74.9 kg 

2.33 kg 

77.2 kg 

Estimated concentration from model with thermocline 

Sediment 

Hypolimnion 

Epilimnion 

Amount in sediment 

Amount in hypolimnion 

Amount in epilimnion 

Total mass 

0.527 :g/g 

8.48 :g/m3 

7.93 :g/m3 

76.3 kg 

1.28 kg 

1.02 kg 

78.6 kg 

Source: Ling et al. (1993). 

There was no indication that contaminants buried in the bottom sediments are transferred 

through diffusion mechanisms back to the surface sediments; however, episodic release of these 

chemicals from surface sediments can occur through mechanisms such as resuspension during 

flooding or lake inversions and uptake/ingestion by benthic biota. 

11.2.4. Biota 

11.2.4.1. Potential Mass of Dioxin-Like Compounds Present 

The mass of CDDs/CDFs in biota in the United States was not estimated as part of this 

report. However, to place perspective on the potential magnitude of this reservoir, 82 g I-TEQDF 

have been estimated to be present in biota in the U.K. (50 g in humans and 32 g in vegetation), 

which is about three orders of magnitude less than the mass estimated to be present in U.K. 

surface soils (Duarte-Davidson et al., 1997; Eduljee and Dyke, 1996). No data are available to 

estimate the biota burden in the United States. 

11.2.4.2. Mechanisms Responsible for Supply to and Releases from Biota 

Apparently, very little of the dioxin-like compounds contained in contaminated soil is 

ultimately taken up by the vegetation growing in the soil. Kjeller et al. (1991) analyzed 

11-22
 
concentrations of CDDs/CDFs in archived soil and grass samples collected from the mid-1840s 

to 1989 at an English experimental station and found that only 0.006 to 0.02% of the soil burden 

of CDDs/CDFs was taken up by the grass. In addition, scientists generally agree that, once taken 

up by plant tissue, CDDs/CDFs are not translocated to other parts of the plant (e.g., fruits or 

shoots) (Bacci and Gaggi, 1985; Hülster and Marschner, 1993, 1994; Nakamura et al., 1994). 

Researchers have found that the concentration of dioxin-like compounds in a plant should 

reach equilibrium with the vapor phase concentrations of dioxin-like compounds in the 

surrounding air (Bacci et al., 1990a, b; Frank and Frank, 1989; Horstmann and McLachlan, 

1992; McCrady and Maggard, 1993; McLachlan et al., 1995; Paterson et al., 1991; Simonich and 

Hites, 1994; Tolls and McLachlan, 1994; Welsch-Pausch et al., 1995). Horstmann and 

McLachlan (1992) stated that the leaf-air transfer of volatile compounds is a reversible process 

governed by concentration gradients. If CDD/CDF concentrations are higher in the surrounding 

air than they are in the air spaces within plant tissue, CDDs/CDFs should diffuse into the plant. 

Once equilibrium is reached and CDD/CDF concentrations in the plant equal that of surrounding 

air, no more CDDs/CDFs should be taken into the plant. When CDD/CDF concentrations in 

surrounding air begin to decrease, CDDs/CDFs should diffuse (probably at a slow rate) out of 

the plant tissue. Apparently, CDDs/CDFs are not bioconcentrated to a significant extent in the 

lipid portion of the leaf cuticle (Gaggi et al., 1985). The CDDs/CDFs present in the leaf tissue 

are predominantly released from the plant through leaf fall onto soil. Therefore, vegetation is 

not likely to be a long-term reservoir of dioxin-like compounds. 

Research suggests that dioxin-like compounds in animal tissue, unlike in vegetation, 

seldom, if ever, reach equilibrium with vapor phase concentrations in the surrounding 

atmosphere (or water column concentrations in the case of aquatic life). Rather, animals 

exposed to dioxin-like compounds are known to bioaccumulate these compounds, primarily in 

body fat (U.S. EPA, 1993g, h). Nonetheless, animals, unlike plants, can metabolize certain 

chlorinated hydrocarbons after they enter the body (Carlberg et al., 1983). Dioxin-like 

compounds can be released from an animal’s body (at congener-specific rates) through 

metabolic processes or through weight loss, breast-feeding, or sweating. McLachlan (1996) 

reported the half-life for the clearance of 2,3,7,8-TCDD from humans to be 7 yr. As a result, 

animal life has a greater potential than does vegetation for being a long-term reservoir source of 

CDDs/CDFs. The majority of the dioxin-like compounds released by animals in the form of 

waste materials will be released to water or soil. Similarly, upon death, the dioxin-like 

compounds remaining in the body will be deposited onto soil or aquatic sediments or will be 

ingested by other animals. 

11-23
 

11.2.4.3. Approaches for Measuring and Estimating Releases from Biota 

Researchers have investigated the uptake and release of CDDs/CDFs by vegetation 

through measurement of actual concentrations during uptake and release by vegetation grown in 

closed systems (greenhouses). Bacci et al. (1992) conducted uptake and release studies of 

1,2,3,4-TCDD by plant foliage in a closed system (a specially constructed greenhouse). 

Concentrations of TCDD vapor in the greenhouse air were maintained during the 370-hr uptake 

phase at a mean concentration of 0.0062 ng/L (air concentration varied slightly from 0.005 to 

0.0075 ng/L). To begin the release phase, the TCDD vapor source (amended sand) as well as the 

greenhouse walls were removed, and release of CDDs/CDFs from the leaves was measured for 

500 hr. The authors concluded that, during uptake, TCDD concentration in the leaves varied as a 

function of time and was dependent on the concentration of vapor-phase TCDD in the 

surrounding air. They estimated the release of TCDD from the vegetation to be relatively slow, 

with a half-life of TCDD of 3,300 hr. 

McCrady and Maggard (1993) conducted a mass balance study of uptake and release of 

dioxin in grass foliage. The results indicated a half-life of dioxin in grass of 128 hr. These 

researchers also noted that, in addition to volatilization, photodegradation of dioxins on the 

foliage appeared to be a significant removal mechanism. They calculated the photodegradation 

half-life to be 44 hr. 

Interpretation of uptake and release data over variable exposure times and contaminant 

concentrations has led to the development of models describing air-to-vegetation equilibrium 

and kinetics controlling the behavior of dioxin in vegetation. Some earlier fugacity modeling 

attempts described the leaf of a plant as behaving as a single compartment. One-compartment 

models were described by Bacci et al. (1990a, b), Trapp et al. (1990), Schramm et al. (1987), and 

Tolls and McLachlan (1994). Researchers presenting most of the recently developed models 

claim that the available data better support the concept of a leaf behaving as two compartments 

(Riederer, 1990; Paterson et al., 1991; Horstmann and McLachlan, 1992; McCrady and 

Maggard, 1993; Tolls and McLachlan, 1994; McLachlan et al., 1995). Input parameters 

considered by most models include critical chemical characteristics of the contaminant, 

characteristics of the plant, exposure times, and contaminant concentrations measured within the 

plant. Riederer (1990) suggested treating a leaf as multiple compartments having different 

accessibilities to the atmosphere and different diffusion resistances. 

Input parameters for the two-compartment model are octanol-water coefficients, cuticlewater partition coefficients, aqueous solubility, and saturation vapor pressure of the chemical of 

concern. Outputs of the model are prediction of equilibrium concentration in different leaf 

tissues, estimates of air-to-vegetation bioconcentration equilibria, and identification of leaf 

compartments in which compounds are likely to accumulate. Riederer (1990) also presented an 

11-24
 

approach for using the model to semiquantitatively assess the potential for revolatilization of 

dioxins from vegetation. 

One advantage of the model presented by Riederer (1990) is that it considers critical 

plant characteristics in the release of dioxins. A plant is an active organism, responding to 

changes in its environment and acting accordingly to ensure its survival. Certain plant 

characteristics, such as the action of stomata (specialized cells usually on the lower leaf surface 

that open and close to control passage of vapors into and out of the leaf interior) and total leaf 

volume, are important factors that affect the release rates of vapor phase contaminants from 

vegetation. 

Paterson et al. (1991) also presented a two-compartment model for release of dioxin-like 

compounds from vegetation. This model describes a plant as being made up of compartments in 

terms of volume fractions of air, water, and nonpolar (lipid-soluble, or octanol-equivalent) 

organic matter. Paterson et al. attempted to show that leaf-air equilibrium and kinetics can be 

correlated with chemical properties of the contaminant and properties of the leaf. The authors 

suggested that the clearance rate constant (k2) can be correlated with the bioconcentration factor. 

This model does not consider critical plant characteristics, such as action of the stomata, and for 

this reason it may be less reliable than models that do consider plant characteristics, such as the 

model presented by Riederer (1990). 

Horstmann and McLachlan (1992) developed a fugacity model to describe release of 

semivolatile organic compounds from the surface of a solid (spruce needles). Their approach 

was slightly different in that their goal was instrument/method development, but their data 

supported the behavior of a leaf as a two-compartment system. 

McCrady and Maggard (1993) also collected data supporting the importance of viewing a 

leaf as a two-compartment system. They used a two-compartment model similar to the one 

described by Paterson et al. (1991) that also does not consider critical plant characteristics and 

thus may be less reliable than models that do (e.g., Reiderer, 1990). 

Tolls and McLachlan (1994) exposed grass cultures for up to 240 hr to several 

semivolatile organic compounds and then measured the release of contaminants from the grass. 

They developed a two-compartment partitioning model based on the data they collected. The 

model consists of a small surface compartment (the leaf cuticle) and a large interior reservoir (air 

spaces within the leaf). Their model assumes that the flux of a chemical is the product of the 

fugacity difference (surface fugacity minus reservoir fugacity) and the conductance between the 

leaf compartments. 

In an attempt to validate this model, McLachlan et al. (1995) compared concentrations of 

semivolatile organic compounds measured in grass grown under field conditions with 

concentrations predicted by their previous laboratory work with a fugacity meter. The 

11-25
 

concentrations measured in the grass cultures agreed with results predicted by the mathematical 

model described by Tolls and McLachlan (1994). 

11.3. SUMMARY AND CONCLUSIONS 

As depicted in Figure 11-1 a set of complex relationships exists among reservoirs and 

between reservoirs and contemporary formation sources. The significance of reservoirs for 

human exposure is more dependent on their ability to affect the concentration of dioxin-like 

compounds in other media than on their size or net release rate. This section first summarizes 

and draws conclusions from the limited information available regarding the character and 

magnitude of reservoir sources. This information is then used to discuss the implications of 

reservoir sources to human exposure. 

11.3.1. Reservoir Sources 

The following are summary statements about soil reservoir sources. 

C	 	 Soil is likely to be the reservoir source with the greatest potential for release of 

CDDs/CDFs to other environmental media, particularly to water. This is due in part 

to its relatively large mass of stored CDDs/CDFs, but more importantly, it is due to 

the existence of demonstrated transport mechanisms for intermedia exchange, e.g., 

soil erosion to surface waters and particle resuspension to air. 

C	 	 The preliminary estimates of CDD/CDF runoff from urban areas to waterways is 

comparable to known industrial point source releases, and runoff from agricultural 

areas to surface waters is more than 100 times greater. It is unclear how much of the 

soil erosion and runoff represents recently deposited CDDs/CDFs from primary 

sources or longer-term accumulation. Much of the eroded soil comes from tilled 

agricultural lands, which would include a mix of CDDs/CDFs from various 

deposition times. The age of CDDs/CDFs in urban runoff is less clear. 

C	 	 Based on the limited information currently available (i.e., primarily fugacity 

modeling), volatilization of CDDs/CDFs from soils is not believed to significantly 

alter ambient air concentrations. However, volatilization of PCBs from soil may be a 

significant process. 

C Based on the limited information currently available, resuspension of soil may 

account for a small fraction (~4%) of CDD/CDF concentrations in air. This 

resuspended soil may, however, constitute a more significant portion of dry 

deposition. 

The following are summary statements about water reservoir sources. 

11-26
 
C It is unclear whether volatilization of CDDs/CDFs from water can significantly alter 

air concentrations. For PCBs, however, the water-air exchange appears to be 

significant, and for some water bodies results in a net transfer from water to air. 

C Water is the major media contributing CDDs/CDFs and PCBs to sediment. Note that 

most of the CDDs/CDFs in sediments originally came from soils. For specific water 

bodies, however, the CDDs/CDFs and PCBs in sediments may have been dominated 

by local industrial discharges to water. 

The following are summary statements about sediment reservoir sources. 

C	 	 It is important to distinguish between surface and deep sediments. Surface sediments 

are commonly resuspended and introduced back into the water; deep sediments 

generally do not interact with the water column. Surface sediments can contribute 

significantly to the CDD/CDF and PCB concentrations in water, whereas deep 

sediments do not. 

C	 	 There is little, if any, movement of dioxin-like compounds once they are buried in the 

bottom sediments. Bottom sediments may be considered as sinks. 

The following are summary statements about biota reservoir sources. 

C	 	 The mass of CDDs/CDFs in vegetation at any given time is likely to be small when 

compared with the mass in soil. Vegetation does play an important role in 

transferring CDDs/CDFs from the air to the soil via the decay of plant biomass. 

C	 	 Release by volatilization from vegetation has been studied and modeled using the 

fugacity approach, and half-lives have been estimated. Based on these results, 

volatilization is not believed to be a significant mechanism for release of 

CDDs/CDFs and PCBs except possibly during forest/brush fires. 

C	 	 The mass of CDDs/CDFs in animals at any given time is likely to be small when 

compared with the mass in soil. Similarly, releases are small and occur primarily by 

excretion and decomposition of dead biomass. 

11.3.2. Implications for Human Exposure 

Although, the ability to make quantitative estimates of releases from reservoir sources is 

limited at present, it is reasonable to conclude that the contribution of reservoir sources to human 

exposure may be significant. Diet accounts for more than 95% of human exposure. Although 

the size of the biota reservoir is small when compared with the soil and sediment reservoirs, it is 

clearly the key contributor to human exposure. The potential contribution of the other reservoirs 

to human exposure is discussed below. 

11-27
 
PCB reservoir releases. Because current sources of newly formed PCBs are most likely 

negligible, human exposure to the dioxin-like PCBs is thought to be derived almost completely 

from current releases of old PCBs stored in reservoir sources. Key pathways involve releases 

from both soils and sediments to both aquatic and terrestrial food chains. One-third of general 

population TEQDFP exposure is due to PCBs. Thus, at least one-third of the overall risk to the 

general population from dioxin-like compounds comes from reservoir sources. 

CDD/CDF releases from soil and sediments to water and exposure via the aquatic 

pathway. The earlier discussion has shown that soils can have significant inputs to waterways 

via soil erosion and runoff. Similarly, the sediment reservoir contributes significantly to 

CDD/CDF concentrations in water. These releases appear to be greater than those from the 

primary sources included in the inventory. Dioxins in waterways bioaccumulate in fish, and fish 

consumption causes human exposure. Fish consumption makes up about one-third of the total 

general population CDD/CDF TEQ exposure. This suggests that a significant portion of the 

CDD/CDF TEQ exposure could be due to releases from the soil and sediment reservoir. 

CDD/CDF releases from soil to air and exposure via the terrestrial pathway. 

Potentially, soil reservoirs could have vapor and particulate releases that deposit on plants and 

enter the terrestrial food chain. The magnitude of this contribution, however, is unknown. EPA 

plans future studies in agricultural areas that will compare modeled air concentrations from 

primary sources with measured levels as a way to gain further insight into this issue. 

11-28
 

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